Water Research 205 (2021) 117658
Available online 11 September 2021
0043-1354/© 2021 Elsevier Ltd. All rights reserved.
Review
Organic pollutants in deep sea: Occurrence, fate, and ecological implications
Edmond Sanganyado
a,b,*, Kudakwashe E. Chingono
c, Willis Gwenzi
e, Nhamo Chaukura
d, Wenhua Liu
a,baGuangdong Provincial Key Laboratory of Marine Biotechnology, Institute of Marine Science, Shantou University, Shantou, Guangdong 515063, China
bSouthern Marine Science and Engineering Guangdong Laboratory, Guangzhou 511458, China
cSchool of Chemical and Process Engineering, University of Leeds, Leeds, LS2 9JT, UK
dDepartment of Physical and Earth Sciences, Sol Plaatje University, Kimberley, South Africa
eDepartment of Soil Science and Agricultural Engineering, Biosystems and Environmental Engineering Research Group, University of Zimbabwe, Harare, Zimbabwe
A R T I C L E I N F O Keywords:
Deep sea
Persistent organic pollutants Biological pump
Marine pollution Polychlorinated biphenyls Organochlorine pesticides
A B S T R A C T
The deep sea - an oceanic layer below 200 m depths – has important global biogeochemical and nutrient cycling functions. It also receives organic pollutants from anthropogenic sources, which threatens the ecological function of the deep sea. In this Review, critically examined data on the distribution of organic pollutants in the deep sea to outline the role of biogeochemical and geophysical factors on the global distribution and regional chemo- dynamics of organic pollutants in the deep sea. We found that the contribution of deep water formation to the influx of perfluorinated compounds reached a maximum, following peak emission, faster in young deep waters (<10 years) compared to older deep waters (>100 years). For example, perfluorinated compounds had low concentrations (<10 pg L−1) and vertical variations in the South Pacific Ocean where the ocean currents are old (<1000 years). Steep geomorphologies of submarine canyons, ridges, and valleys facilitated the transport of sediments and associated organic pollutants by oceanic currents from the continental shelf to remote deep seas.
In addition, we found that, even though an estimated 1.2–4.2 million metric tons of plastic debris enter the ocean through riverine discharge annually, the role of microplastics as vectors of organic pollutants (e.g., plastic monomers, additives, and attached organic pollutants) in the deep sea is often overlooked. Finally, we recom- mend assessing the biological effects of organic pollutants in deep sea biota, large-scale monitoring of organic pollutants, reconstructing historical emissions using sediment cores, and assessing the impact of deep-sea mining on the ecosystem.
1. Introduction
Deep sea (layers below 200 m depth) ecosystems are currently threatened by multiple anthropogenic disturbances such as chemical pollution, overfishing, resource extraction, and climate change (Rogers, 2015). It is the largest ecosystem on Earth and is characterized by a 1.2 billion km3 volume, 434 million km2 seafloor, and a 4.2 km average depth (Rogers, 2015). Deep sea ecosystems play a crucial role in main- taining planetary health and function through supporting unique biodiversity essential for the biogeochemical cycling of nutrients, metals, carbon, and nitrogen. They control nutrient and energy flux across environmental compartments and ecosystem services such as
water circulation, CO2 exchange, biological control, in situ primary production, climate regulation, carbon sequestration, and waste detox- ification (Da Ros et al., 2019). However, information on the structure, function, and composition of the deep sea ecosystem remain scant despite its vast size, partly due to the putative high costs of exploring such ecosystems (Danovaro et al., 2020). Since less than 0.0001% of the ocean floor has been explored, it is difficult to predict the ecological effects of anthropogenic disturbances in the deep sea (Rogers, 2015).
Deep sea ecosystems were previously considered pristine and free from anthropogenic disturbances due to their remoteness. However, organic pollutants enter the deep sea through atmospheric deposition, oceanic transport, and sometimes sea-based anthropogenic activities (e.
* Corresponding author at: Guangdong Provincial Key Laboratory of Marine Biotechnology, Institute of Marine Science, Shantou University, Shantou, Guangdong 515063, China
E-mail address: [email protected] (E. Sanganyado).
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https://doi.org/10.1016/j.watres.2021.117658
Received 2 June 2021; Received in revised form 4 September 2021; Accepted 7 September 2021
g., seabed mining, fishing, ship traffic, and accidental spillage) (Kal- lenborn, 2006). Several studies detected organic pollutants such as polychlorinated biphenyls (PCBs), organochlorine pesticides (OCPs), and polybrominated diphenyl ethers (PBDEs) in seawater, sediments, and biota from the deep sea (Dasgupta et al., 2018; Lohmann et al., 2006). Out of the 17 Sustainable Development Goals set by the United Nations as targets for 2030, eight are directly or indirectly impacted by deep sea pollution (Fig. 1). For example, the presence of organic pol- lutants in the deep sea poses great harm to the ecosystem; thus, threatening the conservation and sustainable use of the ocean resources for sustainable development as required by SDG 14. Although deep sea ecosystems are protected from pollution by global conventions (e.g., Convention for the International Convention for the Prevention of Pollution from Ships, United Nations Convention on the Law of the Sea, and Convention on the Prevention of Marine Pollution by Dumping of Wastes and Other Matter) and regional conventions (e.g., Protection of the Marine Environment in the North-East Atlantic, Convention on the Protection of the Marine Environment in the Baltic Sea, Convention for the Protection of Marine Environment and the Coastal Region of the Mediterranean, Convention for the Protection of the Black Sea), organic pollutants are continually discharged into the environment, and subse- quently reach the deep sea directly or indirectly (Tornero and Hanke, 2016). Monitoring and characterizing anthropogenic disturbances and their effect on key species are essential for a robust assessment of the health of the deep sea ecosystem. This review seeks to examine the literature on the distribution, fate, transport, and trophic transfer of organic pollutants in deep sea ecosystems to determine their sources and transport mechanisms. The key factors influencing the accumulation and transport of organic pollutants in deep sea ecosystems will be dis- cussed. Such knowledge is essential for assessing the environmental risk of organic pollutants and developing conservation and management strategies for deep sea ecosystems.
2. Occurrence of organic pollutants in deep sea ecosystems Rapid industrialization and urbanization compounded by increased demand for food security and quality life have contributed to the rise in environmental emissions of organic pollutants (Table 1). Before their ban or their production restrictions were effected by the Stockholm Convention, it is estimated 2.79 million metric tons of DDT and 1.51 million metric tons of PCBs were produced globally (Fiedler et al., 2019). About 96,000 tons of chemicals based on perfluorooctane sul- fonate (PFOS) were produced globally between 1970 and 2002, while between 1951 and 2015, up to 21,400 tons of perfluoroalkyl carboxylic acids (PFCAs) were produced (Wang et al., 2014, 2013). In 2018, it was estimated that about 400,000 and 100,000 tons of BDE209 (a PBDE congener) were in use or disposed of in waste stocks, respectively (Abbasi et al., 2019). About 30,000 tons of HBCDs are produced glob- ally, with 238,000 tons estimated to be used in China before they are phased out (Li et al., 2016). Due to their extensive application in con- sumer and industrial products and a high potential for long-range at- mospheric (oceanic) transport, organic pollutants are frequently detected in deep sea ecosystems.
The presence of organic pollutants in the global oceans, especially the deep sea, is linked to various sources, which can be classified as land- based and sea-based activities (Fig. 2) (Sanganyado et al., 2020). Their main routes of entry include inputs from coastal regions transported by oceanic currents, atmospheric deposition of free or particle-associated organic pollutants released from land-based activities, unintentional shipping discharges (e.g., accidental oil spill and litter), and deep sea resource extraction (e.g., seabed mining, fishing, and mariculture), waste disposal (Tornero and Hanke, 2016). The entry route of the organic pollutants influences their profile (i.e., concentration and type) and the mechanism by which they transfer from the seawater column to the seafloor. For example, semivolatile persistent organic pollutants (POPs) are often transported to the deep sea by long-range atmospheric transport (Wania and Mackay, 1996). The semivolatile organic
Fig. 1.. The impact of deep sea pollution by organic pollutants on the United Nations Sustainable Development Goals. Icons used with permission from the United Nations (https://www.un.org/sustainabledevelopment).
pollutants deposit on the ocean surface following changes in atmo- spheric conditions. After a while, some of the organic pollutants are revolatilized and carried further by atmospheric transport. This type of long-range transport is called the grasshopper effect (Jurado and Dachs, 2008). A study in Levantine Basin found sediment PCB (12–190 ng g−1) and PAH (<0.3–7.7 ng g−1) concentrations were linked to nearby gas well drilling and dumping sites for sediment dredge (Astrahan et al., 2017). Tables S1 and 2 show the distribution of organic pollutants in deep sea water and sediments, respectively.
Organic pollutants in deep sea ecosystems can also be of natural origin. For example, several studies found that marine bacteria produced organobromines such as hydroxylated PBDEs, methoxylated PBDEs, polybrominated dibenzo-p-dioxins, and polybrominated pyrroles (Agarwal et al., 2017). Agarwal et al. (2017) found that marine sponges, an important biodiversity hotspot in deep sea ecosystems, produced various PBDEs using metagenome-mining techniques. In the western Pacific Ocean, the concentration of methoxylated PBDEs in extractable organic matter from marine sponges (i.e., Haliclona sp. and Callyspongia Table 1.
Examples of organic pollutants found in marine environments and their uses.
Class Compounds Uses
POPsa Organochlorine
pesticides Aldrinb, chlordaneb, chlordecone, DDTb, dicofol, dieldrinb, endrinb, heptachlorb, hexachlorobenzenes (α- and β-HCB)b, lindane, mirexb, pentachlorobenzene, pentachlorophenols, endosulfan, toxapheneb
Controlling pests in agriculture production and public health.
Brominated flame
retardants Decabromodiphenyl ether, hexabromobiphenyl, hexabromocyclododecane
(HBCDD), hexachlorobenzene, hexachlorobutadiene, tetrabromodiphenyl ether Flame retardant additives in sealants, adhesives, textiles, paints, and upholstery.
Halogenated industrial
chemicals Short-chain chlorinated paraffins (SCCPs), perfluorooctane sulfonate (PFOS) and its
salts, perfluorooctane sulfonyl fluoride, polychlorinated biphenyls (PCBs)b Lubricants, plasticizers, adhesives, flame retardants
Unintentional
production Polychlorinated naphthalenes, polychlorinated dibenzo-p-dioxins (PCDD)b,
polychlorinated dibenzofurans (PCDF)b Byproducts of industrial processes
Other organic
pollutants Perfluorinated
compounds Perfluorocarboxylic acids (PFCAs) (e.g., perfluorobutanoate (PFBA), perfluorohexanoate (PFHxA), perfluorooctanoate (PFOA), perfluorononanoate (PFNA), perfluorodecanoate (PFDA)), perfluoroalkylsulfonates (PFSAs) (e.g., perfluorobutanesulfonate (PFBS), perfluorohexanesulfonate (PFHxS), perfluorodecanesulfonate (PFDS))
Water-, stain-, and soil-resistant coatings, insecticides, surfactants, lubricants, and flame retardants
Brominated flame
retardants Other polybrominated diphenyl ethers (PBDEs), besides deca-BDE, Flame retardant additives in sealants, adhesives, textiles, paints, and upholstery.
Unintentionally
produced compounds Polycyclic aromatic hydrocarbons (PAHs) Byproducts of industrial processes or
combustion
aRefers to all the persistent organic pollutants listed in the Stockholm Convention for elimination, restriction, or reduction of unintentional release.
b Represents legacy POPs that were initially listed under the Stockholm Convention; the remaining POPs were later listed and are referred to as emerging POPs.
Fig. 2.. Land- and sea-based sources of organic pollutants and their potential fate and transport pathways. From Sanganyado et al. (2020). Used with permission of Taylor and Francis.
sp.) were 63,500 ng g−1 and 36,500 ng g−1, respectively (Haraguchi et al., 2011). Four methoxylated PBDEs were detected in endemic am- phipods collected from hadal trenches (6,000–11,000 m depth) in the western Pacific Ocean (Cui et al., 2020). However, none were detected in sediments or suspended particulate matter (Cui et al., 2020). Am- phipods play an important role in regulating nutrient and energy flux in marine environments; hence, their species density around marine sponges is often high (Amsler et al., 2009). Gut analysis of nearshore amphipods in the Antarctica Peninsula revealed that some amphipods feed on marine sponges (Amsler et al., 2009). Besides amphipods, organobromines can transfer indirectly to necrophagivore fish (e.g., Paraliparis bathybius, Notoliparis kermadecensis, Pachycara sp., and Bas- sozetus sp.), which feed primarily on amphipods and directly to mega- faunal croppers (e.g., Notacanthus chemnitzii, Barathrites parri, and chimaerids) which feed on marine sponges (Drazen and Sutton, 2017). A study in the western Pacific Ocean found methoxylated PBDEs in fish samples (Haraguchi et al., 2011). Although studies on their distribution and impact in deep sea ecosystems remain scarce due to sampling and analytical challenges, previous studies have shown that naturally pro- duced organohalogens can bioaccumulate and biomagnify in marine mammals; thus, threatening the structure and function of the deep sea ecosystems (Sanganyado et al., 2020; Weijs et al., 2009). A study in the Sea of Japan and North Pacific Ocean found halogenated bipyrroles (up to 4,900,000 ng kg−1 wet weight) and methoxylated PBDEs (up to 190, 000 ng kg−1 wet weight) concentrations were detected in cetaceans in the order of killer whales >toothed whales >baleen whales demon- strating the critical role of feeding ecology on natural organobromine biomagnification (Fujii et al., 2018). Overall, although naturally pro- duced organohalogens are often present at concentrations lower than organic pollutants, biosynthesis needs to be seriously considered a key source of potentially toxic organohalogens in deep sea ecosystems.
3. Factors influencing the transport of organic pollutants in the deep sea
3.1. Biological pump
The transport of organic pollutants in deep sea ecosystems is often regulated by a complex ensemble of biophysical processes involved in transporting organic carbon from the ocean surface to the deep sea and sediments (i.e., biological pump) (Fig. 3) (Galb´an-Malagon et al., 2012). ´ The biological pump can be conceptualized as a biogeochemical system comprising primary production, vertical transport, and sedimentation (Lutz et al., 2007). Phytoplankton accumulate organic pollutants from the surface waters during the primary production stage, driving air-water disequilibrium in organic pollutant concentration. A previous study found phytoplankton uptake and gravitational settling of biogenic particles increased the air-water influx of PCBs (Galb´an-Malag´on et al., 2012). Primary productivity is often governed by solar, physical oceanographic, and climatic factors, and this often results in seasonal and regional trends in the efficacy of the biological pump in seques- trating the organic pollutants (Lutz et al., 2007).
Additionally, physicochemical properties of the organic pollutants (e.g., aqueous solubility, hydrophobicity, volatility, and lipophilicity), local environmental conditions in the deep-water (e.g., depth, oxygen content, and temperature), settling particle characteristics (e.g., size, density, and type), climate/geographical factors (e.g., latitude, ocean currents), seafloor geomorphology (e.g., topographical features), and biodiversity can further influence the organic pollutant sequestration efficiency of the biological pump. Since phytoplankton readily uptake more hydrophobic compounds, higher air-water influxes have been observed for PCB congeners with higher octanol-water partition co- efficients (log KOW) in high productivity regions such as the North Atlantic, North Pacific and Arctic Oceans (Gioia et al., 2008). PCB congeners with low log KOW are often more abundant in high latitude oceans such as subtropical regions, where biological pump efficiencies are often low (Sobek and Gustafsson, 2004). The air-water flux of
Fig. 3..An overview of the major biophysical processes governing the fate and transport of organic pollutants in deep sea ecosystems.
organic pollutants and changes in their export are critical determinants of the vertical profile of dissolved organic pollutants. Hence, under- standing the sequestration efficiencies and mechanisms of the biological pump is imperative since these processes impact the ‘sink and source’ role of the deep sea.
3.2. Physical pump
Water masses originating from continental margins transport organic pollutants to the deep ocean through lateral diffusion, eddy currents, intrusion, and hyperpycnal flows/thermohaline currents (Puig et al., 2014). A study in the continental margin of the Gulf of Mexico found a positive correlation between particle-attached PAHs (0.9 and 7.0 ng L−1) and particulate organic carbon 4000–131,000 ng L−1), although both variables negatively correlated with salinity (Adhikari et al., 2019).
The results suggested riverine discharge was the primary source of PAHs in the deep sea (Adhikari et al., 2019). In the South Atlantic, upper Equatorial, and Indian Oceans, lateral ocean circulation accounted for 20–48% PCB influx (Wagner et al., 2019). At the same time, upwelling contributed up to 10% in removing PCBs from the deep sea to the ocean surface (Wagner et al., 2019). In the Fram Strait, a deep-water channel that transports organic pollutants into the Arctic Ocean, the West Spitsbergen and the East Greenland Currents contributed a total mass flux for α-HCH of 29 tons y−1 and 67 tons y−1, respectively (Ma et al., 2018). However, the ocean currents moved in the opposite direction.
The West Spitsbergen Current moved northwards from the Atlantic to the Arctic Ocean and the East Greenland Currents moving southwards along Greenland. In contrast, there was a net influx of PCBs (i.e., PCB 28, 101, 153, and 180) into the Arctic Ocean transported through the Fram Strait between 1930 and 2015 (80 tons y−1) (Ma et al., 2018). In the Aegean Sea, cyclonic surface circulation and deep-water outflows from the Cretan Strait were shown to be a crucial advective source of PAHs and contributed to the accumulation of PAHs in sediments (Hatzianestis et al., 2020). In stratified oceans, ocean currents and the seafloor interact over short timescales lasting (<5 years) through horizontal, vertical, and eddy currents (i.e., internal solitary waves). By driving processes such as sediment resuspension and transport in the seafloor, internal solitary waves contribute to the outflow and inflow of organic pollutants into the deep sea (Jia et al., 2019). Hence, the role of oceanic currents in the total flux of organic pollutants in the deep sea is often in a state of competition whereby the net flow may be influenced by geophysical, seasonal, and climatic factors.
Global thermohaline currents driven by temperature and salinity gradients have been shown to move surface water together with organic pollutants to the deep sea in a process called deep water formation (Lohmann et al., 2006). A study on the Norwegian, Ross, Weddell, and Labrador Seas found deep water formation was a major route of entry of PCBs to the deep sea (870 kg yr−1) compared to particle settling (320 kg yr−1) (Lohmann et al., 2006). A study on the distribution of organic pollutants in sediments from the Mediterranean Sea found that the Western Mediterranean Deep Water formation increased the deposition of POPs with a low air-water partition coefficient (log KAW) of -4.0 to -2.0 such as lindane, tri-PCBs, tetra-PCBs, and methylphenanthrenes (Salvad´o et al., 2019). In contrast, there was a strong association be- tween POPs with an intermediate log KAW (-2.0 to 0) and high octanol-water partition coefficient (log KOW >6) and sediment particles suggesting horizontal transport in the continental shelf (Salvad´o et al., 2019). Perfluorinated compounds reached 3500 m depths in young deep water masses (<10 years) in the Labrador Sea (North Atlantic Ocean) (Yamashita et al., 2008) but were not detected below the permanent thermocline in old deep waters (>450 years) in the Central Arctic Ocean Basins (Yeung et al., 2017). The intrusion of perfluorinated compounds to deep waters is often negligible in old deep waters such as those found in the Central Arctic Ocean Basins (Yeung et al., 2017) and Fram Strait (Joerss et al., 2020) because the estimated time for deep water forma- tion in these regions is usually more than 100 years yet perfluorinated
acids were first produced and used in the 1960s (Yamashita et al., 2008).
Hence, deep water formation can contribute to the depletion of per- fluorinated compounds from the global ocean surface, particularly in the North Atlantic and Southern Oceans (Lohmann et al., 2006). However, PFAS contamination in the South Pacific Ocean was shown to be negligible (below detection limit to <10 pg L−1) in ocean surface and the deep water compared to other oceans (Yamashita et al., 2008). This was probably because there was no direct discharge of PFAS in the South Pacific Ocean since the water masses comprise of ~1000-years old ocean currents and Antarctic circumpolar water masses. Previous mass flow estimates showed that PFOS and PFOA flux in the deep water layer of the Labrador Sea was four and ten times higher than PCB flux, respectively (Lohmann et al., 2006; Yamashita et al., 2008). The results suggested highly soluble and less volatile organic pollutants transferred from the surface water to the deep water faster than the semi-volatile and rela- tively more hydrophobic organic pollutants. Even though deep water flux of PFAS in the Labrador Sea is relatively fast, transferring the PFOAs (emitted globally to date) into global oceans requires at least 4500 years (Yamashita et al., 2008). In the North Atlantic Ocean, PFOS concen- trations were higher and more variable in the surface water (0–10 m) compared to the subsurface water (365–510 m) and the water layer around the permanent thermocline (985–1335 m) (Zhang et al., 2017).
The lag time between peak PFOS emission and peak concentrations in subsurface and permanent thermocline layer water was 2–3 years and more than 30 years, respectively. This was probably because vertical mixing in the North Atlantic is intense during winters and can reach 600 m depths while subduction and ventilation of water in the surface mixed layer to the permanent thermocline is slow (Zhang et al., 2017). For that reason, the distribution of perfluorinated acids in oceans is often used as chemical tracer to assess the transport of anthropogenic pollutants by oceanic currents.
PFAS profiles and concentrations significantly vary between surface and deep water oceans, except in seas that experience intense vertical mixing (Wei et al., 2008; Yeung et al., 2017). A study on per- fluorooctanesulfonate (PFOS), perfluorooctanoate (PFOA) and per- fluorobutanesulfonate (PFBS) in global oceans found distinct stratification in the sources of the compounds; the North Atlantic Cur- rent was the major source of the perfluorinated acids in the surface waters, the Labrador current in the subsurface waters, and the Denmark Strait Overflow Water in the deep layers (below 2000 m depth) (Yamashita et al., 2008). In Western Mediterranean Sea, PFAs concen- trations in the deep water (141 pg L−1) were about two times lowers than those in the surface water samples (357 pg L−1) (Brumovský et al., 2016). The Western Mediterranean deep water has short renewal cycles since at least two intense dense shelf water cascading and open-sea convection events were observed between 2004 and 2012. The verti- cal profile of PFAS in the Mediterranean and Japan Seas substantially differed even though both are semi-enclosed seas with deep waters isolated from the open ocean (Yamazaki et al., 2019). Yamazaki et al.
(2019) found that the PFAS concentration in the Mediterranean Sea were an order of magnitude higher than in the Japan Sea. The PFAS concentrations steadily decreased with an increase in depth in the Japan Sea, with only PFDA and PFHxS having a maximum concentration in the mid-deep waters (1000–1500 m). In contrast, PFAS vertical profile was highly variable in the Mediterranean Sea, although at some locations higher concentrations of PFDA, PFNA, PFHxA, and PFHpA were detec- ted in the deepest water columns sampled (2300–3600 m) (Yamazaki et al., 2019). This is probably because PFNA, PFHxA, and PFHpA are legacy PFAS that were extensively used in fluoropolymer products (Lin et al., 2020). Overall, the PFAs profiles in young deep waters such as the Mediterranean Sea are indicative of the PFAS profiles in surface seawater of a decade earlier, while those of old deep waters are asso- ciated with legacy PFAs. However, currently missing are data on PFAS profiles in deep sea waters surrounded by low-income countries, where several banned PFAs are likely to be still in use.
3.3. Settling particle characteristics
Biogenic, inorganic, and anthropogenic particles are important vectors for the vertical transport of organic pollutants in oceans. As the (non)biogenic particles sink, organic pollutants may be released to the deep-water column via (bio)transformation or dissolution of particles.
However, the profile of organic pollutants transported can be influenced by the nature of the settling particle. For example, a study in the Indian, Pacific, and Atlantic Oceans found that the vertical flux of PFAS via zooplankton fecal matter was one to two orders of magnitude higher than that of PFAS associated with phytoplankton (Gonz´alez-Gaya et al., 2019). Settling of large particles (>50 μm) in deep marine ecosystems contributes significantly to the vertical transport of pollutants through the water column, while small particles are carried by water masses and may also contribute to the advective transport of the organic pollutants (Martí et al., 2001). The primary source of organic pollutants in small particles is partitioning from water particles. In contrast, organic pol- lutants in large biogenic particles come directly from the food chain (Martí et al., 2001). A study in Mediterranean deep waters found that the concentration of halogenated organic contaminants on the small parti- cles (<0.7 μm) decreased significantly with depth compared to large particles (<50 μm) (Martí et al., 2001). Hence, particle size can influ- ence the accumulation rate of organic pollutants in the water column and sediments.
Microplastics are complex particles comprising of a dynamic mixture of monomers, additives, and processing agents (Galloway et al., 2017).
They often host microorganisms and bind organic pollutants and organic material, which alters the microplastic density and surface charge resulting in a change in bioavailability of the organic pollutants (Galloway et al., 2017). An estimated 1.2–4.2 million metric tons of plastic debris enter the ocean through riverine discharge, of which 14.4–236 thousand metric tons are floating in the global ocean (Jam- beck et al., 2015; Weiss et al., 2021). Around 99.8% of the plastic debris discharged into the oceans since the 1950s are estimated to be below the ocean surface, in the subsurface, deep water layer, or seafloor (Koel- mans et al., 2017). Plastic particles are transferred to the deep water and seafloor through accumulation in biota following ingestion, thermohaline-driven currents, and gravitational settling following loss of buoyancy or incorporation into fecal matter (Kvale et al., 2020). A recent study showed that Bathochordaeus stygius, one of the most abundant zooplankton, effectively transported microplastics from the ocean surface to the deep sea by ingesting and then excreting micro- plastics as part of their fecal pellets (Katija et al., 2017). During physi- cochemical and biological weathering of plastic debris to form microplastics, plastic monomers, additives, and plastic attached organic pollutants inadvertently leach to the surrounding water column; thus, contributing to the flux of organic pollutants in the deep sea (Dasgupta et al., 2021). This is greatly concerning since around 10,500 compounds are used worldwide as plastic monomers, additives (e.g., antioxidants, biocides, flame retardants, plasticizers, light stabilizers, and colorants), and processing aids (e.g., heat stabilizer, crosslinking agent, lubricant, solvent, and antistatic agent), of which 2486 are substances of potential concern while no hazard data was available for 4100 substances (Wie- singer et al., 2021). A previous study found polychlorinated biphenyl (127–142 ng g−1), organochlorine pesticides (4280–5350 ng g−1), and chlordane (1080–1260 ng g−1) concentrations in microplastics and plastic pellets found in the Xisha Trough in the South China Sea (Das- gupta et al., 2021). The concentration of organochlorine pesticides in Xisha Trough was higher than those found in coastal beaches (2.2–1970 ng g−1) on the South China Sea, while PAH concentrations were an order of magnitude lower than those found on the beach (11.2–7710 ng g−1) (Shi et al., 2020). Organochlorine pesticides such as DDT are more hy- drophobic than PAHs and tend to have lower desorption and degrada- tion rates during transport to the deep sea from coastal environments (Bakir et al., 2014). However, long-term desorption of organic pollutants from microplastics is highly influenced by the polymer-water partition
coefficients of the compounds. For example, a previous study found light PCB congeners desorbed faster (e.g., PCB-8 half-life = 14 days) than heavier congeners (e.g., PCB-209 half-life =210 years) (Endo et al., 2013). Hence, it is expected that microplastic-associated PCBs in the deep sea would be dominated by heavier congeners. However, the less recalcitrant and more water-soluble tri-PCBs were the most dominant PCB congeners (>92%), and the composition was consistent with the technical mixtures produced in China (Dasgupta et al., 2021). Low molecular weight PCB congeners like tri-PCB have more extended at- mospheric reach; they are sometimes generated during the microbial transformation of high molecular weight PCBs and are preferentially exchanged during ocean-air interactions due to their lightweight (Das- gupta et al., 2021, 2018). In vitro digestive models have revealed microplastic can act as a vector or cleaner (extracting organic pollutants from the gut) in aquatic organisms (Mohamed Nor and Koelmans, 2019).
However, comprehensive data on the flux of organic pollutants into the deep sea through microplastic transport are still lacking.
4. Factors influencing the accumulation of organic pollutants in the deep sea
4.1. Microbial biotransformation
It is often assumed that the biotransformation mechanisms of organic pollutants in the deep sea and surface water ecosystems are similar despite the drastic differences in temperature and hydrostatic pressure (Louvado et al., 2015). At low temperatures, the enzymatic activity of organohalogen-degrading microorganisms often decreases, resulting in low organic pollutant degradation efficiency. In contrast, high hydro- static pressure in deep sea ecosystems was shown to decrease the fungal degradation of poly-β-hydroxybutyric acid (Gonda et al., 2000). Previ- ous studies showed that bacteria in deep sea ecosystems developed phenotypical characteristics such as cell membrane with a higher amount of unsaturated fatty acids and lack of genes involved in photo- synthetic reactions, which enabled their adaptation to the psychro-peizophilic conditions (Louvado et al., 2015). These phenotypic adaptations may reduce the permeability of the deep sea microorgan- isms, resulting in decreased organic pollutant uptake and intracellular biotransformation. The low temperatures and high hydrostatic pressure conditions can facilitate the role of deep sea ecosystems as the ultimate sink of organic pollutants. However, organohalogen-degrading bacteria may have adapted to the psychro-peizophilic conditions as they devel- oped higher diversity in metabolic activity (Konstantinidis et al., 2009).
Since studies modeling the global fate and transport are often based on the chemodynamics of the organic pollutants in surface water, they may under- or overestimate the global distribution of organic pollutants, considering the deep sea is probably the largest environmental compartment organic pollutants partition.
Microbial transformation of freely dissolved- or particle-attached organic pollutants contributes to removing organic pollutants in the deep sea. Next-generation sequencing analysis of deep sea sediments from the Great Australian Bight detected three genes (i.e., alkB, c23o, and pmoA) associated with aerobic hydrocarbon degradation (van de Kamp et al., 2019). Various studies have shown that bacteria isolated from different deep sea geomorphologies such as hydrothermal vents could degrade organohalogens. Bacteria isolated from hydrothermal vents have been shown to readily degrade PAHs and n-alkanes since they evolved to use hydrocarbons as their primary carbon source (Ma et al., 2021). Gammaproteobacteria (i.e., FJ613315, Pseudomonas stutzeri strain hyss62) and Actinobacteria (i.e., HM222653, Nocardioides sp. 0701C5–1) that oxidize PCBs were previously isolated from deep sea sediments in the Kongsfjorden (Papale et al., 2017). The removal efficiency of P. stutzeri contained the bphA gene, and its removal efficiency was 10–69% at 4 ◦C and 7.3–93% at 15 ◦C (Papale et al., 2017). The di- and tri-PCBs were more efficiently removed at both temperatures than tetra-PCBs suggesting heavier congeners might be more recalcitrant
than lighter congeners in the deep sea sediments. The degradation of organic pollutants by deep sea bacteria might follow pathways different from those observed from coastal, soil, freshwater, and open ocean bacteria. Deep sea bacteria adapted to lower oxygen, nutrient, and temperature, and higher pressure conditions, which probably altered how they use organic pollutants as food sources (Scoma et al., 2019).
Studies on the biotransformation of organic pollutants other than hy- drocarbons in deep sea ecosystems remain scarce.
4.2. Bioavailability
Particle-associated organic pollutants often exhibit low bioavail- ability, the extent to which depends on their hydrophobicity. The bioavailability of organic pollutants often changes as the particle- associated organic pollutants are transported to the deep water layer due to changes in temperature, salinity, oxygenation, and particle characteristics. This is partly because the physicochemical characteris- tics of the organic pollutants (e.g., lipophilicity, dipole-moment, and water solubility) may alter as the water chemistry changes between the surface, subsurface, to deep water (Lyytik¨ainen et al., 2003). Interest- ingly, previous studies have shown that the accumulation of PFAS in biota was driven by phospholipid and protein interactions, processes that are governed by the polarity and hydrophobicity of the compound (Ng and Hungerbühler, 2014). Additionally, changes in sediment char- acteristics (e.g., particle size, polarity, and organic carbon content) due to changes in water chemistry with depth or microbial degradation may alter the bioavailability of organic pollutants (Lyytik¨ainen et al., 2003).
For example, sediments in the Blanes Submarine Canyon had poly- chlorinated dibenzofurans (PCDF) concentrations (754 ng kg−1) strongly correlated with soot (r2 =0.852, p =0.008) but not organic carbon content (Castro-Jim´enez et al., 2013). This was probably because PCDF strongly sorbs to soot and not to organic carbon as demonstrated by the low fOC (KOW/ρOC)/fSC•KSCW ratio (0.2–1), where fOC represents the mass fraction organic carbon, KOW the octanol-water partition co- efficient, ρOC the octanol density (g L−1), fSC the mass fraction of soot content, and KSCW is soot-water partition coefficient (Castro-Jim´enez et al., 2013). The partitioning of PCDFs to soot in deep sea ecosystems might reduce their bioavailability resulting in the sediments acting as an ultimate sink. Furthermore, Hadal sediments are known to have very low organic carbon content (<0.22%), yet high concentrations of PCBs (930–4200 ng kg−1) and PBDEs (245–590 ng kg−1) have been detected in hadal sediments from the Marian Trench (Dasgupta et al., 2018). The hadal sediments had high contents of clay minerals such as illite, cli- nochlore, and nontronite, which are known to promote (i) steric effects between bulky or nonplanar organic pollutants and the clay minerals and (ii) the formation of complexes between weakly hydrated cations in the clay minerals and the negatively charged moieties of the organic pollutants (Liu et al., 2015). Hence, there is need for additional studies on the role of sediment characteristics on the bioavailability of organic pollutants in deep sea.
Compounds with low water solubility and high hydrophobicity often preferentially partition to the rapidly sinking particles while those with high water solubility preferentially remain in the water column. The trend is probably not universal because a previous study found that there was no significant correlation between PFAS flux to the deep water layer and the hydrophobicity of the compounds (Gonz´alez-Gaya et al., 2019).
In fact, long chain PFCAs which had a relatively higher hydrophobicity had low flux to the deep chlorophyll maximum (Gonz´alez-Gaya et al., 2019). PFAS are probably transported from the surface to the deep water through the biological pump since Gonz´alez-Gaya et al. (2019) demonstrated that particle settling flux negatively correlated to the dissolved phase concentration of PFAS in the deep chlorophyll maximum. However, a previous study found that there was no signifi- cant correlation between PFAS accumulation in phytoplankton and PFAS hydrophobicity (Casal et al., 2017). Hence, the adsorption or bioaccumulation of PFAS such as PFOA to sinking particles is not
primarily dependent on compound hydrophobicity but other properties such as polarity (Ng and Hungerbühler, 2014). There is a need for further studies to assess the effect of organic pollutant bioavailability on the efficacy of biological pumps in removing organic pollutants from the sea surface.
4.3. Water column depth
Biogeochemical processes in the water column influence the occur- rence, settling, and fate of organic pollutants in the ocean (Galb´an-Malag´on et al., 2012). In the Arctic Ocean, PBDE concentrations were an order of magnitude higher in the deep waters than in the polar mixed layer suggesting a predominancy of vertical transport of the PBDEs (Salvad´o et al., 2016). The relative contribution of less bromi- nated congeners (i.e., tri- to hepta-BDEs) increased with depth compared to the heavier congeners (Salvad´o et al., 2016). The increase with depth of BDE-71, which are not found in technical mixtures, and other less brominated congeners suggest heavier congeners transformed to lighter congeners during vertical or oceanic transport (Salvad´o et al., 2016).
Previous studies in the Central Arctic Ocean Basin (Carrizo et al., 2017) and Fram Strait (Ma et al., 2018) found that transformation of DDTs to DDEs contributed to the relative increase in DDEs with depth. In contrast, the concentration of highly chlorinated PCBs increased with depth, suggesting sorption to settling particles contributed to the verti- cal transport of the PCBs (Ma et al., 2018). Passive sampling employed by Ma et al. (2018) has shown to be a valuable tool for assessing the vertical distribution of organophosphate esters (OPEs) and PBDEs in the Fram Strait (McDonough et al., 2018) and PCBs, PAHs, and OCPs in the Irminger Sea (Booij et al., 2014), although little vertical trends were observed. In the Fram Strait, chlorinated OPEs were the most dominant OPEs (34–100%) in the deep water mass compared to alkyl and aryl OPEs (McDonough et al., 2018). The total concentration of the chlori- nated OPEs (0.006–0.430 ng L−1) was an order of magnitude higher than that of the alkyl/aryl compounds (0.0001–0.066 ng L−1) (McDonough et al., 2018). Chlorinated OPEs are more persistent in the deep water mass since they are less degradable via hydrolysis or indirect photolysis than the alkyl/aryl OPEs. Particle, sediment, and water sampling chal- lenges continue to limit our understanding of the role of particle size on the mass influx of organic contaminants in the deep ocean ecosystem.
Passive samplers are more suitable for assessing the vertical and global distribution of organic pollutants since they are cost-effective, easier to handle, and have high analyte enrichment potential. As a result, a global monitoring program for organic pollutants in global oceans, including deep waters, called the Aquatic Global Passive Sampling (AQUA-GAPS) network was established in 2017 (Lohmann et al., 2017).
Accumulation of organic pollutants in the deep sea can be influenced by ocean stratification. Stratification in the St. Lawrence Maritime Es- tuary and Gulf water column occurred due to cold and oxygen-rich waters from the Labrador Current and warm and oxygen-deficient wa- ters from the Atlantic Ocean isolated the surface waters and the deep waters (Picard et al., 2021). The concentration ratios in subsurface to deep water ratios were significantly higher for current-use PFAS (e.g., the ratio for PFBS was 6.1) than legacy PFAS (e.g., the PFOS and PFOA ratios were 2.5 and 1.8, respectively) (Picard et al., 2021). Pharma- ceuticals in the Gulf of Cadiz exhibited a similar trend, probably due to the water column stratification caused by the presence of dense Medi- terranean waters and less saline Atlantic Ocean waters (Biel-Maeso et al., 2018). Since the water masses in stratified seas have different physicochemical properties, it is expected that the (bio)transformation of the organic pollutants and the settling particles will differ with depth.
However, there are few studies on the role of water mass characteristics and the fate of organic pollutants.
4.4. Seafloor geomorphology
Recent advances in exploration and sampling technology revealed
deep sea ecosystems contain distinct geomorphological features such as seamounts, submerged plateaus, submarine canyons, trenches, and hy- drothermal vents (McClain and Rex, 2015). Besides being a hotspot of faunal abundance, diversity, and sometimes endemism, deep sea geomorphological features often alter oceanographic conditions (Rogers, 2015). The interaction between ocean currents and a steep geomorphological feature (e.g., valleys, ridges, and canyons) often promotes the accumulation of organic matter and sediments that origi- nate from the continental shelf zone (Astrahan et al., 2017). For example, Castro-Jim´enez et al. (2013) found that the total 2,3,7, 8-PCDD/Fs concentration of 17 congeners (102–680 ng kg−1 d.w.) were three to six times higher in sediments at the deepest locations of the Blanes Submarine Canyon (1700 m depths) than at the shallow locations (500 m depths) or the adjacent open slope. Ocean current circulations in submarine canyons promote the flushing of sediment particles and their associated organic pollutants down the slope resulting in the accumu- lation of the sediments and organic pollutants in the deepest locations of the canyon. These ocean currents are often initiated by coastal storms and dense shelf water cascading (i.e., movement of water masses following cooling, evaporation, or freezing of the ocean surface in the shelf continent, resulting in dense water formation that sinks as warm and saline water rises). A recent study in the Gulf of Cagliari revealed submarines canyons were the primary physical pathway transporting PAHs and PCBs from land to the deep sea (Tamburrino et al., 2019).
Astrahan et al. (2017) found that the highest amounts of organic matter in bottom sediments corresponded with the locations of ridgelines and canyons. The sediments with high organic carbon content had higher concentrations of highly hydrophobic PCBs and PAHs (Astrahan et al., 2017). In addition, seafloor geomorphology can also affect critical ecosystem features such as population structure, composition, and function, which drive biotransformation and bioaccumulation of POPs in the deep sea (Costello and Chaudhary, 2017). A previous study found seafloor morphologies such as troughs, sediment waves, and furrows increased biodiversity in the deep sea, which in turn influenced ecosystem processes (e.g., biological processes involved in the biological pump mechanisms) at a regional scale (Zeppilli et al., 2016).
4.5. Ecological implications
The fate and transport behavior of organic pollutants in deep sea ecosystems determine their magnitude and duration of exposure to deep sea organisms. Hydrophobic organic pollutants such as DDTs, highly chlorinated PCBs, and highly brominated PBDEs readily bioaccumulate in deep sea organisms due to their high lipophilicity and recalcitrant nature. However, biomagnification and bioaccumulation of organic pollutants in deep sea ecosystems is influenced by the habitat niche (e.g., pelagic, benthic, and demersal), physicochemical properties, and occurrence of the organic pollutants and metabolic capacity. For example, hydrophobic organic pollutants such as PCBs, DDTs, and PBDEs with log KOW of 6.0–8.0 were shown to significantly bio- accumulate in carnivorous fish such as blackbelly lanternshark (Etmop- terus lucifer), Kaup’s arrowtooth eel (Synaphobranchus kaupii), snubnosed eel (Simenchelys parasitica), and eelpout (Lycodes hubbsi) (Takahashi et al., 2010). This suggested that feeding habits contribute significantly to the biomagnification of organic pollutants in deep sea ecosystems. Organic pollutants with log KOW <6.0 (e.g., HCHs, HCB, and HBCDs) equilibrate relatively faster with the aqueous phase than DDTs, PBDEs, and highly chlorinated PCBs. As a result, while trophic magnification can be expected for highly hydrophobic organic pollut- ants, concentrations of HCHs and HCB have smaller variations across trophic levels (Takahashi et al., 2010). In addition, fish from the western North Pacific Ocean and along the Oyashio Current had higher HCH concentrations than those from warmer locations along the Kuroshio Current. Due to their high vapor pressure, HCHs readily partition to the atmosphere under warm conditions resulting in less HCH available for bioaccumulation in fish (Takahashi et al., 2010).
Bioaccumulation and biomagnification of organic pollutants in deep sea biota are influenced by an interplay between interspecies traits such as habitat niche, vertical mobility, respiration rate, lipid composition, metabolic capacity, and food web interconnectivity and oceanographic factors such as ocean currents (Castro-Jim´enez et al., 2013; Romero-R- omero et al., 2017). For example, a study in the Avil´es Submarine Canyon found that PCBs and PBDEs biomagnified in the pelagic but not in the benthic food web (Romero-Romero et al., 2017). PCB-108, 118, 138, 153, and 180 were the most dominant congeners in deep sea biota, probably due to their abundance in technical mixtures. There was high inter-specific variation in the distribution of the in the deep sea fish.
However, considering the limited sampling size of the study, additional studies are required to assess intra-specific variations in organic pollutant bioaccumulation. In the Blanes Submarine Canyon, the highest PCDD/F concentrations were detected in nektobenthic crustaceans (220–795 ng kg−1 l.w.) and fish (110–300 ng kg−1 l.w.) (Castro-Jim´enez et al., 2013). Since pelagic species have high vertical mobility, organic pollutants bioaccumulation in pelagic species tends to be associated with the concentration of organic pollutants in primary producers and the surface water. In contrast, by inhabiting primarily on the seafloor, benthic and nektobenthic species often have organic pollutants profiles linked to deep sea sediments and waters as well as relatively highly hydrophobic contaminants (Castro-Jim´enez et al., 2013).
5. Future directions
Determining the mechanisms, pathways, and rates of fate and transport of organic pollutants in deep sea ecosystems is challenging as it is a multidisciplinary endeavor requiring skills in physical oceanog- raphy, analytical chemistry, biogeochemistry, marine ecology, envi- ronmental microbiology, and even hydrology. Traditional approaches for sampling and detecting organic pollutants in environmental samples based on oceanographic cruises are often time-consuming, labor-inten- sive, complex, and impractical for long-term monitoring due to cost. The depth, high hydrostatic pressure, and low temperatures prevalent in deep sea ecosystems limit the applicability of traditional technologies for collecting, storing, and manipulating samples; hence, the scarcity of deep sea environmental samples (Cario et al., 2019). Since deep sea conditions are difficult and expensive to replicate in the laboratory, microcosm studies to investigate the (bio)transformation and sorption behavior of organic pollutants is challenging. Hence, future studies should answer the research questions described in Table 2 to better understand the fate, transport, and impact of organic pollutants in deep sea ecosystems. We discuss these critical research topics in the subse- quent passages.
Traditional ocean monitoring approaches lack adequate spatial, vertical, and temporal coverage required for a multiscale, multi- compartment, and multidisciplinary assessment of complex biogeo- chemical processes in the cast deep sea ecosystems. Biogeochemical processes controlling the chemodynamics of organic pollutants vary with space and time, hence the need for multi-dimensional coverage (Kaiser and Barnes, 2008). While autonomous platforms such as floats, underwater vehicles, and floats have been successfully used in measuring physical oceanographic processes, gravitational settling of particles, and mechanisms and efficiency of biological carbon pump, very few studies used the platforms to understand the chemodynamics of organic pollutants (Chai et al., 2020). Autonomous underwater vehicles are often used to map seafloor geomorphologies, locate ocean dumping sites, track contaminant plumes and thermocline, and collect environ- mental samples (Jones et al., 2019). Tracking the thermocline can be a powerful tool for assessing the role of deep-water formation or water mass flux in the input of organic pollutants in the deep sea (Table 2).
Coupling autonomous underwater vehicles with photometry can offer a detailed ephemeral view of the seafloor, and it has been previously used to identify ocean dumping of DDT containers (Kivenson et al., 2019). A previous study used a Sentry Precision Robotic Impeller Driven plankton
sampler coupled with autonomous underwater vehicles successfully collected plankton and larva from the Blake Ridge Seep (2160 m depth) on the US Atlantic Margin (Billings et al., 2017). Therefore, autonomous platforms are valuable for assessing the role of biological and physical pumps in the global dynamics of organic pollutants in deep sea ecosystems.
The difficulties in assessing biological responses following organic pollutants exposure in biota at spatial or temporal scale could be alle- viated by using an autonomous underwater vehicle for sampling deep
sea species, linking exposure to biological response using chemical analysis combined with biomarker analysis, transcriptomics, lipidomics, or metabolomics (Deleo et al., 2021), and paleo-ecotoxicology linking historical contamination to deep sea fossils of different biological or- ganisms or species diversity changes using environmental DNA (Korosi et al., 2017). Understanding the biological effects of organic pollutants in deep sea ecosystems is difficult due to the complexity of the mixture effect as biota are exposed to multiple threats (Sanganyado et al., 2020).
Current ecotoxicological assays often disregard the potential mixture Table 2.
A summary of key research areas that require further study in understanding the impact of organic pollutants in deep sea ecosystems.
Research topic Research question Major challenges and limitations Research priorities Key Refs.
Biological
effects Are the concentrations of organic pollutants present harmful to the deep sea ecosystems?
The effects of organic pollutants are complex and species-specific. Most studies focused on: (i) single compound toxicity rather than mixture effects; (ii) single species
Employ integrated chemical and bioanalytical analysis to understand the biological effects at a molecular and cellular level.
(Deleo et al., 2021;
McConville et al., 2018;
Theron et al., 2020) Continental shelf-deep
sea exchange How does continental shelf geomorphology influence the accumulation and transport of organic pollutants in the deep sea?
Seafloor geomorphologies are an interface between the continental shelf and the deep waters. However, their complex and spatially variable physical attributes may govern the accumulation and transport of organic pollutants.
Identify the key characteristics of different geomorphologies (e.g., submarine canyon and trenches) that influence organic pollutants accumulation and transport.
(Haalboom et al., 2021;
Romero-Romero et al., 2017; Salvad´o et al., 2012)
Air-sea exchange Does air-air-sea exchange influence the profile of organic pollutants accumulating or leaving deep sea ecosystems?
Volatile and semivolatile organic pollutants often enter the deep sea via atmospheric deposition and pass through the productive euphotic zone, where the composition may be altered by abiotic and biotic processes before reaching the deep sea.
There is a need for large scale data on the distribution of organic pollutants in multiple environmental compartments such as air, ocean surface, deep waters, and sediments.
(Ge et al., 2021)
Temporal trends in organic pollutants loading
Does changing usage patterns due to regulations influence the distribution of organic pollutants in deep sea ecosystems?
Although regulatory actions resulted in decreased production and fresh discharge of organic pollutants, primarily discharge from stockpiles and old equipment is ongoing. Organic pollutants are also secondarily discharged from electronic waste, melting glaciers, and unintentional production.
Use dated sediment cores to reconstruct the historical inputs and annual fluxes of organic pollutants s in deep sea sediments.
(Combi et al., 2020)
Physical vs biological
pump Is water mass more important than the biological pump, phytoplankton uptake, and other biogeochemical processes controlling the vertical distribution of organic pollutants in deep sea ecosystems?
Tracking water mass origins and the gravitational settling of particles in the deep sea is challenging due to sampling difficulties.
Establish the effect of vertical depth on the biogeochemical processes controlling the fate of organic pollutants.
Conduct large-scale studies on the role of phytoplankton uptake on global dynamics of organic pollutants.
(Dachs et al., 2002; Sun et al., 2016)
Source apportionment What are the sources of POPs in
different deep sea ecosystems? Organic pollutants from primary sources often retain the composition of the commercial mixture while those from secondary sources are altered.
However, the organic pollutants in deep sea ecosystems originate from multiple sources.
Employ chiral signatures, congener ratios, and changes in ratios of parent organic pollutants and their metabolites to determine contaminant sources.
(Jones, 2021)
Bioaccumulation and biomagnification of organic pollutants
What are the mechanisms and pathways of the long-term biouptake of organic pollutants in the deep sea food web? How does it differ from other ecosystems?
Constructing deep sea food webs to determine the long-term biouptake of organic pollutants is challenging due to sampling challenges. Biomagnification across the food web is influenced by the physicochemical properties of the organic pollutants, the local oceanographic conditions, and ecological factors.
Employ stable isotope ratios and opportunistic sampling to construct deep sea food webs.
(Cui et al., 2020;
Webster et al., 2014)
Emerging contaminants and transformation products
Are emerging contaminants and transformation products accumulating in deep sea ecosystems?
More than 1,600 emerging contaminants have been identified as priority pollutants in marine environments.
Emerging contaminants enter the deep sea through multiple input pathways.
Hence, investigating all emerging contaminants is expensive, tedious, and require highly sensitive and selective analytical equipment.
Emerging contaminants transform in the deep sea to form potentially more toxic contaminants.
Prioritization of emerging contaminants based on QSAR and usage patterns.
Identification of major inputs pathways of emerging contaminants in the deep sea.
(Azaroff et al., 2020)
effects caused by the interaction of the organic pollutants with natural stressors such as salinity, cold temperature, and high hydrostatic pres- sure prevalent in deep sea ecosystems (Mestre et al., 2014). In addition, the bulk of the literature on organic pollutants in deep sea ecosystems tend to report total concentrations, while data on bioaccessibility and bioavailability which are more relevant for assessing impact of organic pollutants on ecosystem health, are scarce. Future studies on the bio- logical effects of organic pollutants should mimic the deep sea condi- tions to ensure the ecotoxicological data generated for environmental risk assessment is relevant and accurate.
There is a need for comprehensive studies exploring the impact of sea-based activities such as deep sea mining on organic pollutants che- modynamics and deep sea biogeochemical processes. The discharge of dissolved substances and nutrients due to deep sea mining activities alters the seafloor morphology, benthic geochemistry, deep water chemistry, primary production, and food web structure, which may change the biological and physical pump efficiencies (Levin et al., 2020). Deep sea mining can cause habitat degradation, resulting in loss of biodiversity and species connectivity; thus, altering the food web structure. However, the lack of large-scale studies on the ecological consequences of deep sea mining makes it difficult for regulators such as the International Seabed Authority to develop toxicity thresholds that are adequate for minimizing the adverse effects caused by deep sea mining (Kaikkonen et al., 2018). Baseline studies on the biological, chemical and physical condition of the deep seabed are required to ensure future environmental monitoring programs for organic pollut- ants can adequately determine the effect of exploration and exploitation activities (Ginzky et al., 2020).
Large-scale monitoring of spatial and temporal changes in mass flow of organic pollutants in deep sea ecosystem is essential for under- standing the impact of human activities on the deep sea. While global monitoring systems have been proposed for monitoring current trends in the distribution of organic pollutants in oceans, historical emissions have been widely estimated using historical usage patterns. Linking historical emissions data to changes in deep sea ecosystems is difficult due to uncertainties in estimating organic pollutant flux and linking exposure to biological response. Recent studies demonstrated that combining paleolimnology and ecotoxicology techniques by assessing changes in contaminant profile and biological proxies (e.g., fossils and environmental DNA) down a dated sediment core analysis could provide critical data on historical emission and its associated ecological conse- quences (Korosi et al., 2017). Deep sea sediment cores have been suc- cessfully used to reconstruct past climatic and oceanographic changes (Artemova et al., 2019; Toomey et al., 2013), evaluate recovery of seafloor from bottom trawling (Paradis et al., 2021), assess changes in microbial community structure with sediment depth (Wang et al., 2010), and estimate historical flux of mercury in the Northwest Pacific Ocean (Aksentov and Sattarova, 2020). Reconstructing past emissions of organic pollutants particularly in oceans could help in assessing the impact of chemical regulations on organic pollutant flux in deep seas.
Transformation products should not be overlooked when monitoring the distribution of organic pollutants in deep water ecosystems. In the St.
Lawrence Maritime Estuary and Gulf, the concentration of metolachlor ethanesulfonic acid, a metolachlor transformation product, was nearly thirty times higher than that of the parent compound (Picard et al., 2021). This is particularly concerning since metolachlor ethanesulfonic acid was shown to be more toxic to zebrafish larvae than S-metolachlor as they caused overexpression of genes involved in the regulation of the thyroid system and cell cycle (Rozm´ankov´a et al., 2020). Additionally, determination of transformation product and parent compound ratios together with congener ratios and chiral signatures can be a useful tool for source apportionment in deep seas (Shi et al., 2020). Source appor- tionment can aid in the development of environmental regulations that protect deep sea ecosystems.
6. Conclusion
Although organic pollutants have been detected in deep sea envi- ronments for the past 50 years, more data on biogeochemical, geophysical, oceanographic, and sedimentological processes are required to make meaningful predictions of the global fate of organic pollutants in the deep sea. Such data can be obtained using passive samplers as they have high analyte enrichment factor, low matrix in- terferences, provide concentrations of the bioavailable phase, and can provide time-weighted average concentrations. However, most in- vestigations on the distribution of organic pollutants in deep sea envi- ronments were conducted using active sampling techniques, while passive sampling is rarely used. Therefore, future studies on the che- modynamics of organic pollutants in the deep sea should employ passive sampling.
There were significant variations in the vertical profiles of organic pollutants in particular deep seas or between regions. The variations could be attributed to the pollutants (e.g., physicochemical properties such as log KOW, volatility, aqueous solubility, and use and production patterns), oceanic currents (e.g., deep water formation and eddy diffu- sion), and biogeochemical processes in the water column. Considering much of the deep sea remains unexplored and targeted analysis was predominantly used in these studies, the chemodynamics of several organic pollutants in much of the unexplored ocean remains unknown.
To better understand the ecological risk posed by organic pollutants, future studies should explore the factors influencing their bio- accumulation and biomagnification in deep sea food webs. In addition, since the toxicological effects of organic pollutants on much of the deep sea biota remain poorly understood, there is a need for experimental and in vitro models for understanding the toxicity mechanisms of organic pollutants in various deep sea biota. Autonomous underwater vehicles equipped to deploy passive samplers can help extend the region of deep seas investigated, while suspect and non-target screening techniques could help investigate the behavior of ‘known unknown’ and ‘unknown unknown’ pollutants in the deep sea.
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. The content of this publication has not been approved by the United Nations and does not reflect the views of the United Nations or its officials or Member States.
Acknowledgment
The authors gratefully acknowledge the financial support by the Key Special Project for Introduced Talents Team of Southern Marine Science and Engineering Guangdong Laboratory (Guangzhou) (Grant No.
GML2019ZD0606), and Shantou University Research Start-Up Program (Grant No. NTF20002), Li Ka Shing Foundation Interdisciplinary Research Project (Grant No. 2020LKSFG04E).
Supplementary materials
Supplementary material associated with this article can be found, in the online version, at doi:10.1016/j.watres.2021.117658.
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