1 Introduction

Nanomaterials (NMs) refer to particles with at least one dimension < 100 nm [1]. Such particles have always been present in the environment, derived from natural and in more recent times, anthropogenic sources [2]. In recent years, ecological concerns have been raised in regards to the environmental fate and behaviour of engineered NMs. These materials are observed to behave differentially to their bulk counterparts [3], and unique physicochemical properties has led to their widespread application, for example; as antibacterial agents in the biomedical industry [4]; catalysts [5]; conductive adhesives or pastes [6]; agricultural fertilisers [7, 8], cosmetics [9]; and as agents for biosensing and imaging (Table 1) [10]. The production of NMs has increased exponentially and is predicted to rise [2, 11]. Ever-increasingly, NMs are being incorporated into commercial products and > 1800 are now listed as containing nano-components [11, 12]. With this in mind, it is increasingly likely that NMs will enter the natural environment, where the ocean represents the sink for the majority of contaminants entered into aquatic systems. Marine ecosystems play a fundamental role in regulating global climatic and biogeochemical cycles, as well as, significantly contributing to food security, world economics and human health.

Table 1 Global production and uses of common engineered nanomaterials [13,14,15]
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The available research examining the potential effects of NMs upon marine organisms has grown considerably in recent years, and a number of reviews are now available e.g. References [1, 32, 33]. Due to their small size, NMs are bioavailable to a great range of biota and adverse impacts of NMs have been recorded in marine species occupying various trophic levels, including; fish [34,35,36,37], zooplankton [38], marine bivalves [39,40,41], and phytoplankton [42,43,44]. NMs have been recorded to exert a wide variety of toxic effects, not limited to but including; genotoxicity and chromosomal damage [45,46,47], neurotoxicity [48], damages to the functioning of the immune system [49], alterations to organismal behavioural (e.g., burrowing and feeding behaviour of marine invertebrates) [39, 50], and growth inhibition of phytoplankton associated with reduced photosynthetic performance and oxidative stress [51,52,53,54,55,56,57].

The ocean faces an era of environmental change, characterised by increases in average sea temperature and ocean acidification (OA) driven by global climate change, as well as the unprecedented introduction of emerging pollutants of anthropogenic origin [58, 59]. Whilst evidence for the likely ecotoxic effects of NMs upon marine biota is increasing, the impact of co-occurring environmental stress on NM toxicity remains uncertain [60,61,62,63]. With the extent of environmental change taking place in the ocean, it will be increasingly important to understand the potential interactions between anthropogenic contaminants and environmental stressors. In this review, the current evidence available on the effects of (i) ocean warming, (ii) OA, and (iii) co-exposure on NM ecotoxicity towards marine biota is discussed, in order to consolidate such information and provide recommendations for future research. A clearer understanding of the likely interaction between environmental stressors and emerging pollutants is key to evaluating their long-term risk and the implementation of effective monitoring and management strategies.

2 Impacts of environmental change on the ecotoxicity of nanomaterials towards marine biota

2.1 Ocean warming

Due to anthropogenic climate change driven by increased CO2 emissions, average surface seawater temperatures are predicted to increase 1–6 °C by 2100 [64]. As a result, the ecophysiology of marine biota is likely to be altered due to changes to metabolic function and changes in seawater chemistry and nutrient availability [65,66,67]. Such changes have the potential to exert damaging effects on ecosystem functioning. For example, it has been estimated that phytoplankton biomass within the global ocean will be reduced ~ 6% by 2100 due to increases in sea surface temperatures and nutrient stratification [68]. The occurrence of extreme variations temperature, such as marine heatwaves, is also predicted increase [69]. This may be particularly evident in coastal zones where seawater temperatures appear more dynamic. Given their close proximity to anthropogenic activity, coastal areas are also often where pollutants of anthropogenic origin are most abundant, thus highlighting the need to understand the combined impact of these two stressors which are likely to occur simultaneously.

A number of studies have been conducted to investigate the interaction between ocean warming and ecotoxicity of NMs (Table 2), and this number appears to have increased in recent years. Research has been carried out on organisms spanning various trophic levels, including: fish [70, 71], zooplankton [71], bivalves [60, 72] and phytoplankton [52, 71]. In particular, bivalves such as marine mussels being have been most commonly studied (Table 2). Increased temperature appears to alter the manner by which biota interact with NMs, in turn enhancing toxic responses such as disruption to photosynthetic processes, altered metabolic function and weakened immune responses [51, 52, 60, 70, 72, 73].

Table 2 Summary of works carried out examining the combined impact of ocean warming and nanomaterials upon marine biota
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Phytoplankton represent the base of the marine food web and contribute substantially to global biogeochemical processes, accounting for approximately 50% of primary productivity on Earth [74]. As such, it is of upmost importance to understand the implications of environmental change and exposure to pollutants on this ecologically significant group of species, as well as their likely combined effect. However, despite this need, little evidence if currently available examining the impact of altered temperature upon the ecotoxicity of NMs towards marine phytoplankton. Previously, NMs have been recorded to exert a range of toxic effects upon marine phytoplankton species, including growth inhibition, oxidative stress, phototoxicity, albeit often recorded at NM concentrations far exceeding those predicted in the environment [42, 55, 75,76,77,78,79]. Such effects, often arise due to altered photosynthetic efficiency or metabolic activity due to chemical stress, processes that are believed likely to be altered by changes in temperature [80, 81]. It is therefore likely that ocean warming and NM exposure are likely to interact in some fashion to drive effects in phytoplankton. Indeed, phototoxicity experienced during AgNP exposure towards phytoplankton has been recorded to vary with temperature [52]. Exposures of the green alga, Dunaliella tertiolecta, to AgNPs over 48 h, revealed decreases in photosynthetic performance which were more severe at 31 °C compared to 25 °C [52]. At this higher temperature, disruption of electron transport in Photosystem II was believed to drive the enhanced decline in photosynthetic performance [51, 52]. In a similar fashion, the diatom S. costatum was also recorded to suffer greater adverse effects of nZnO exposure when temperature was increased. Here, toxicity was again associated with a decrease in photosynthetic performance [71]. It is possible that as global seawater temperature rises, populations inhabiting warmer waters may suffer greater adverse effects of NM exposure in terms of photosynthetic performance than those in cooler waters. It is clear that additional research examining the combined impact of warming and NM exposure on phytoplankton, as well as other emerging contaminants, is required.

Marine bivalve species appear the most commonly selected model for combined warming and nano-ecotoxicology studies. These organisms which often occupy benthic zones are of particular relevance to nano-ecotoxicological research within the marine environment, given that NMs are widely reported to undergo sedimentation within saline media due to increased ionic strength, therefore increasing their bioavailability to such organisms [3, 82,83,84,85,86]. A temperature increase of 4 °C (18–22 °C) has been recorded to enhance sensitivity of marine bivalve species towards nTiO2, significantly altering metabolic activity [73]. Here, electron transport activity was significantly higher than controls at both 5 and 50 µg L−1 nTiO2 when exposed at 22 °C, while no such difference was observed at 18 °C [73]. Although, interestingly, for this measure of metabolic capacity this effect was not observed at the highest concentration (100 µg L−1) [73]. However, whilst 28-d exposure to rutile nTiO2 (0–100 µg L−1) revealed a dose-wise adverse effect in the gills and digestive glands of the mussel M. galloprovincialis, no significant impact of varying temperature on these histopathological endpoints were recorded [73]. Notably, accumulation of nTiO2 by mussels during this study was highest at 18 °C compared to 22 °C, attributed to the higher extent of aggregation and hence sedimentation of particles in the warming condition [73].

Adverse impacts of nZnO exposure have been recorded to vary with season, largely attributed to a variation in temperature [60]. By exposing the blue mussel Mytilus edulis to nZnO (0–100 µg L−1) for a period of 21 d under winter and summer conditions, as well as assessing the impact of additional 5 °C warming, Wu and Sokolova observed an alteration in immune response associated with temperature. In this work, nZnO exerted higher toxicity than dissolved zinc. Interestingly, in winter conditions (10 °C) nZnO caused an increase in phagocytosis and a strong immune response was identified via transcriptomics, which was suppressed under summer temperatures (15 °C) and additional warming [60]. It is suggested by the researchers that the combination of nZnO and warmer sea temperatures may have implications for innate immunity of marine mussels during the summer when pathogen concentrations are often higher [60]. In related works, Wu et al. displayed that environmentally relevant concentrations of nZnO (µg L−1) can exert adverse impacts on the bioenergetics of marine mussels which similarly was exacerbated at higher summer temperatures, an important period for such species when reproduction occurs [72].

Ocean warming, as well as sub-optimal temperatures, were found to exacerbate nZnO toxicity towards larvae of the sea urchin, Tripneustes gratilla [62]. This toxicity was associated with the release of ionic Zn2+ and was apparent at nZnO concentrations of > 0.001 mg L−1 at temperatures of 25 °C and 29 °C, whilst larvae were able to resist concentrations of up to 1 mg L−1 at optimal temperature (27 °C) [62]. It is possible that this effect results from varied release of ionic Zn2+ from nanoparticles at the varied temperatures, combined with the alteration in metabolic and enzymatic processes related to temperature [62]. Interestingly, at all tested temperatures, low concentrations of nZnO (0.001 mg L−1) were beneficial to T. gratilla larvae, speculated to be due to the requirement for zinc for essential enzymes involved in the calcification process [62, 87, 88].

An increase in temperature from 20 to 25 °C increased toxic effects of AuNPs (5 nm) towards juvenile fish, Pomatoschistus microps [70]. Individuals were recorded to increase gold uptake approximately two-fold and this was associated with a chemical stress response [70]. As with nTiO2 in works by Leite et al., at the higher temperature AuNPs suspensions were observed to be less stable and displayed rapid aggregation of non-spherical aggregates. However, in this case such processes are thought to have enhanced gold uptake by P. microps due to increasing bioavailability [70]. Previously, it has been reported that dissolution is expected to be higher at elevated temperature, and this has been observed in number of NMs such as AgNPs and nCu [89,90,91]. Often toxic effects of NMs are attributed to the release of toxic ionic species, rather than the NM directly, for example for AgNPs [92]. Therefore, it is possible that toxic effects mediated by such release of ionic species may increase under ocean warming conditions, however this is likely to be material-specific.

Whilst most studies have focused upon one species of interest, exposure of multiple species representing various trophic group can reveal species-specific impacts of combined exposure to NM and warming. Wong and Leung examined the toxicity of nZnO towards species representing marine phytoplankton, zooplankton and fish under a range of temperatures varying 10–30 °C. During this time the fate and behaviour of nanoparticles within artificial seawater was also assessed. Such information is of vital importance to fully understand any synergistic or antagonistic relationship between NMs and environmental stressors. Here, it was recorded that dissolution of Zn2+ from nanoparticles was reduced as temperatures increased, whereas no clear pattern could be seen between temperature and nZnO aggregation [71]. For the diatom Skeletonema costatum and the amphipod Melita longidactyla negative effects of nZnO exposure were enhanced with increased temperature; the 96 h IC50 value for S. costatum was 17 mg L−1 at 15 °C, and dropped considerably to 3 mg L−1 in the 28 °C treatment. Similarly, for M. longidactyla 96 h LC50 values were > 3.3 mg L−1 when exposed to nZnO at 10 °C, falling to 0.08 mg L−1 in the 30 °C treatment [71]. Given that nZnO toxicity is generally attributed to release of dissolved zinc, which was found to be greatest at low temperatures, such findings do not appear to follow trends seen in previous work [71]. For the amphipod this was proposed to be attributed to an alteration in behaviour at lower temperatures which reduced their uptake of nZnO and prevented negative effects [71]. Once more, this highlights the complexity in evaluating the response of biota to both environmental stressors and contaminants which may individually cause altered behaviour and physiology. Growth of Oryzias melastigma fish larvae improved with increased temperature and no significant interaction between temperature and nZnO was recorded [71].

The evidence available reveals a number of potential synergistic adverse effects of combined exposure to NMs and ocean warming in marine species. Whilst limited, research examining this interaction in phytoplankton species shows the potential for increased phototoxicity during NM-exposure to a range of materials under warming [52, 71]. Such effects may have implications of primary productivity and local ecosystem function. Due to altered resilience of various taxa to increases in temperature and NM exposure respectively [77, 93, 94], as well as in combination, adverse impacts of growth and photosynthesis may act to drive alterations in the composition and performance of the phototrophic community. As a result, higher trophic levels may experience decreased prey availability and/or quality. The physiology of marine invertebrate species has been recorded be negatively affected by combined exposure to warming and NMs. Damages to metabolic functioning, innate immunity and oxidative stress were recorded in marine mussels and sea urchins alike [60, 62, 72, 73], with potential impacts for reproduction due to the seasonality of such effects [72]. Marine invertebrates represent a keystone species within the marine environment and also represent an important group of species for aquaculture. Warming has also shown evidence to enhance uptake of metal ions released by NMs [70], which could facilitate trophic transfer and biomagnification through the food chain. This is concerning and it is of importance to study this feature in greater detail, as well as to investigate whether similar findings are recorded with other pollutants. Typically, ocean warming is believed to enhance toxic effects of NMs by altering their fate and behaviour in seawater (i.e., aggregation and dissolution), subsequently altering bioavailability of NMs or their toxic products such as released ionic species [62, 70, 73].

2.2 Ocean acidification

During the last century, an estimated 1.8 ppm year−1 of pCO2 has been added to the atmosphere due to human activity [95]. The ocean plays a key role in mitigating this increase in atmospheric pCO2, taking up around a quarter of emissions [96]. However, CO2 quickly undergoes dissolution in seawater, resulting in decreased pH and causing OA [97]. On average the pH of seawater is predicted to decrease from ~ 8.1 to 7.8 by the year 2100 [98]. In recent years, concerns have increased regarding the likely impact of OA on marine biota, particularly calcifying organisms [99,100,101]. OA has been reported to impact a variety of marine species [99]. For example, corals are predicted to reduce calcification activity, while alteration to phytoplankton community structure and physiology has also been recorded, with potential impacts on carbon cycling [97, 99, 102]. In addition, alterations in water chemistry driven by lowered pH is likely to alter the fate and behaviour of NMs, therefore influencing their bioavailability and interaction with biota in a material-specific fashion [90, 103, 104]. Given the potential stress that OA may exert independently, alongside possible altered bioavailability of NMs in a more acidic ocean, it is therefore important to consider the potential impact of acidification on the ecotoxicity of nano-pollutants.

As in studies examining ocean warming, primarily the impact of OA on NM toxicity has been investigated using marine invertebrate species (Table 3), many of which are economically important for aquaculture [105]. Here, a number of synergistic adverse effects of NM exposure and have been recorded [40, 63]. A combined impact of ZnO NPs at high concentrations (10 mg L−1) and low pH (7.3) caused a long-lasting adverse effect upon the mussel, Mytilus coruscus, even after stressors were removed [106]. This effect was greater than either stressor applied individually, which also caused significant negative impacts on haemocytes, including increased haemocyte mortality and reactive oxygen species content [106]. Interestingly, individuals were still able to survive under all conditions [106]. Similar results have been recorded in response to nTiO2 and OA conditions in the same species [107]. In this study, low pH resulted in greater aggregation of nTiO2, enhancing their uptake and likely increasing toxic effects [107]. In both of these studies, the influence of NMs on toxicity was greater than that of low pH [106, 107]. When combined with OA, nTiO2 has also been found to impair normal function of digestive enzymes of the marine mussel M. cortuscus [63]. However, this interaction was only investigated at values exceeding those predicted in the environment (2.5 and 10 mg L−1) [63]. It is proposed that ingestion of nTiO2 may cause damage to the digestive gland of the mussels, whose digestive function is already compromised by low pH [63]. Notably, presence of nTiO2 drove an increase lysozyme activity, believed to be an immune response of the digestive gland against nTiO2 [63, 108].

Table 3 Summary of works carried out examining the combined impact of ocean acidification and nanomaterials upon marine biota
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Interestingly, OA and NM exposure have also been recorded to act synergistically to alter feeding and behaviour of marine mussels due to physiological changes [110, 111]. For example, decreased feeding rate and metabolic activity of the mussel M. coruscus in response to nTiO2 was found to be significantly exacerbated by low pH [110]. Additionally, combined exposure of low pH and nZnO at high concentrations (10 mg L−1) caused a reduction in predator avoidance in M. coruscus due to a reduction in adhesive strength of byssus thread used to attach to substrates [111].

In terms of oxidative stress, OA had no additional impact during nZnO exposure to M. coruscus, however both stressors caused a number of oxidative responses individually [61, 106]. In contrast, oxidative stress caused by high concentrations of nTiO2 (10 mg L−1) on the same species was exacerbated under acidified conditions [40], highlighting the variability in response to different nano-components. Here nTiO2 was found to aggregate to greater sizes at low pH, believed to increase their uptake by the mussels and may explain the increase in negative effects recorded [40, 112]. Oxidative stress was also enhanced in the polychaete H. diversicolor under OA conditions (pH 7.6) in response to both functionalised and pristine carbon nanotubes (CNTs, 0.01 and 0.1 mg L−1) [113]. Differences in the impact of pH on the toxicity of the two CNTs was recorded. The combined stressors of lowered pH (7.6) and functionalised CNTs acted to enhance lipid peroxidation, however such an effect was not observed with pristine particles [113]. Neurotoxicity was a feature of exposure, and as more apparent under acidified conditions [113]. In addition, the combined effects of acidification and CNT exposure compromised the ability of H. diversicolor to utilise energy reserves, no such effect was observed in the control pH condition (8.0) even in the presence of NMs [113]. Similarly, to other studies, the synergistic toxicity of OA and NMs was believed to be attributed to varied fate and behaviour of particles under lowered pH. A decrease in aggregation rate and hence increased stability of the CNT suspension was proposed to increase bioavailability of particles and their resultant toxicity [113]. In a similar fashion, during carbon nanotube exposures towards marine bivalve species, low pH reduced aggregation and enhanced stability of nanotube suspensions, in turn increasing their bioavailability and exacerbating toxic effects characterised by decreases in respiration and alterations to metabolic function [105]. Xia et al. also found that OA increased the inhibitory effects of nTiO2 growth of marine microalga Chlorella vulgaris due to stabilisation of the nTiO2 suspension, thought to increase the internalisation of particles into cells, inducing oxidative stress [115]. This result highlights the material-specific variability in response to experimental conditions, given that OA enhanced aggregation of nTiO2 in the study by Huang et al., described above.

Nanoplastics (NPs) represent an emerging contaminant whose occurrence in the environment is believed to be underrepresented by current analytical capabilities due to their small size. These particles can either be manufactured to be in the nano-range, or are believed to occur following the degradation of plastic debris within the aquatic environment [116,117,118,119,120]. For example, it is estimated that one microplastic particle sized 5 mm may break down to produce 1014 nanoplastic particles sized 100 nm [121]. Both OA and plastic exposure have previously been recorded to be detrimental to the development of Antarctic krill, a species which plays a key ecological role [122, 123]. Rowlands et al. examined the combined effects of ocean acidification (pH 7.7) and PS NPs (160 nm, 2.5 mg L−1) on the embryonic development of Antarctic krill following 6 d exposure. A key feature of exposure was the extensive aggregation of NPs, which reached sizes in the micron-range after just 24 h. The aggregation of negatively charged PS NPs within saline media is widely reported [124,125,126]. Synergistic adverse impacts of acidification and exposure to PS NPs was observed through a significant reduction in krill embryonic development from approximately 22% in the control to 13% in the OA and NP treatment, a response not observed after exposure to either stressor singularly. Here, the researchers propose acidified conditions may have acted to alter leaching of chemical additives present in the PS NPs, enhancing their toxicity, or alteration to metabolic function due to altered homeostatic processes required under varied pH [122].

A range of effects from altering organismal physiology to behaviour have been recorded in co-exposures of NMs and OA. However, a clear limitation of the current literature is the variety of marine taxa that have been examined, thus making it difficult to accurately evaluate the likely impact on the wider ecosystem. It is clear that future research is required in this area, where studies upon marine phytoplankton that play such key ecological roles will be key. Here, in particular work on calcifying organisms such as coccolithophores would be of significant interest. In a similar fashion to studies examining ocean warming described above, the differences in NM toxicity under OA conditions are largely a result of altered fate and behaviour of NMs when pH is lowered, subsequently altering their bioavailability and mode of toxicity.

2.3 Co-exposure

A great number of anthropogenic pollutants enter the marine environment, and hence NMs are unlikely to be the sole contaminant that biota interact with, particularly in coastal regions where anthropogenic pollution appears at its greatest [70, 127,128,129]. NMs have been recorded to interact with a number of co-contaminants in the aquatic environment [2, 38, 130,131,132,133,134,135], and research in this area is beginning to emerge in the literature. Materials such as nTiO2 display high affinity for a number of metals and organic substances, enhancing toxicity of contaminants including TBT, PAHs and heavy metals in marine invertebrate species by increasing their bioavailability and facilitating their uptake [2, 38, 130,131,132,133]. To further evaluate the likely impact of NMs in the natural marine environment, efforts should be made to investigate the likely synergistic effects of individual stressors upon biota, as well as impacts upon the fate and behaviour of each respective contaminant, which may be influenced by one another [135].

NMs have been found to enhance the toxicity of other particulate materials such as microplastics (MPs) [134, 136]. For example, growth of the diatom S. costatum was inhibited by co-exposure to CuNPs and MPs [134]. Similarly, exposure of the microalga Tetraselmis chuii to AuNPs and MPs significantly reduced specific growth rate, whilst exposure to either contaminant alone had no adverse effect [137]. It is worth noting that this effect was only recorded at AuNP and MP concentrations of 3 mg L−1 and 4 mg L−1, respectively, far exceeding environmental concentrations [137,138,139]. At lower concentrations no such adverse effect was recorded. Whilst, synergistic toxic effects between NMs and other pollutants have been recorded, results vary. Presence of MPs had no impact on toxicity exerted by AuNPs towards juvenile fish, P. microps. Although here, presence of AuNPs decreased MP concentrations in test media significantly, a process enhanced at high temperature [70]. Antagonistic effects of NMs upon co-contaminants have also been observed. Nano-sized cerium was found to slightly reduce the toxic effects caused by exposure to mercury in the marine mussel M. galloprovincialis. However, significant toxicity was still recorded in treated individuals [140].

The presence of multiple contaminants within the environment is likely to occur in localised areas, and their presence is likely to influence one another in a variety of ways, as has been recorded. To date, the limited evidence available does not allow for a detailed conclusion to be reached upon the likely impact of NM and co-contaminant exposure upon marine biota, but it does highlight the importance of such research in the future.

3 Conclusions

As a result of changes in our global climate, environmental change will inevitably occur in the ocean and hence it is important to incorporate this into future ecotoxicological study. In this review, a number of studies have been highlighted to provide examples of such an approach, and present insightful evidence to consider the influence that co-stressors may have on nano-ecotoxicology in the marine environment.

In the majority of studies, the specific toxic effects caused by NM exposure were largely consistent when presented to biota as the sole contaminant, or in combination with warming or OA treatments. The impact of these environmental variables appeared to primarily be an alteration to NM fate and behavior during exposures, which in turn altered their bioavailability to test organisms and hence either mitigated or enhanced toxic effects (Fig. 1). In a number of cases, both warming and OA caused an increase in sedimentation of NM, thus facilitating the rapid transport of NMs through the water column towards deeper areas and sediments. Whilst this process may immediately reduce bioavailability to planktonic organisms, it likely increases exposure to benthic and sediment-dwelling biota, likely enhancing any toxic effects. Indeed, this can be seen in the evidence whereby marine bivalve species experienced greater adverse effects of NM exposure following warming- or OA-mediated increases in NM aggregation, thus enhancing their uptake by these benthic species [73]. Due to the occurrence of such changes to NM behavior, it will be beneficial for researchers in the field to conduct further research on how co-occurring stressors may influence NM fate in the natural environment, so those species at highest risk can be highlighted for ecotoxicological study.

Fig. 1
figure 1

Graphical representation of the manners by which environmental change may influence nano-ecotoxicology in the ocean; a Co-exposure with other marine pollutants (e.g., microplastics) may alter the bioavailability of NMs by increasing/decreasing their stability in the water column; b ocean warming and acidification may enhance aggregation of NMs, facilitating their transport to deeper zones and eventually sediments where they may accumulate (c). In this fashion, the bioavailability of NMs and any related toxic species to benthic and sediment-dwelling species may be enhanced; d Conversely, ocean warming and acidification may increase the stability of NMs in suspension, facilitating their transport via ocean currents and increasing their bioavailability, and hence toxicity, to species occupying higher zones in the water column; e ocean warming and acidification may increase/decrease the release of toxic species from NMs (i.e., reactive oxygen species (ROS) or ionic species), in doing so these environmental changes may enhance or mitigate NM toxicity

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In terms of studies examining nano-ecotoxicology in the presence of a second anthropogenic contaminant, to date, evidence appears too limited to provide any clear conclusion. However, such work is of upmost importance when considering the unprecedented entry of such materials into the ocean, and given that contaminants will almost inevitably co-occur within hotspots of pollution. Therefore, expanding our knowledge in this area will be of high interest, particularly in terms of developing monitoring and management strategies for heavily polluted zones.

In the works included in this study, the vast majority focused upon marine invertebrate species, primarily bivalves. Although such organisms represent a model of high interest, and are important species for aquaculture, increased data on other taxonomic groups is required. In particular, studies on marine phytoplankton would be greatly beneficial given their fundamental role in marine ecosystem functioning and position at the base of the marine food web. Future studies should focus attention on environmentally relevant concentrations of NM to enhance applicability of research to real-world conditions, and consider time-scales that fully capture the fate of NMs in the marine environment.