Ecophysiological perspectives on engineered nanomaterial toxicity in fish and crustaceans

    Ecophysiological perspectives on engineered nanomaterial toxicity in fish and crustaceans Neal Ingraham Callaghan, Tyson James MacCormack PII: DOI: Reference:

S1532-0456(16)30177-6 doi:10.1016/j.cbpc.2016.12.007 CBC 8272

To appear in:

Comparative Biochemistry and Physiology Part C

Received date: Revised date: Accepted date:

31 August 2016 1 December 2016 20 December 2016

Please cite this article as: Callaghan, Neal Ingraham, MacCormack, Tyson James, Ecophysiological perspectives on engineered nanomaterial toxicity in fish and crustaceans, Comparative Biochemistry and Physiology Part C (2016), doi:10.1016/j.cbpc.2016.12.007

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ACCEPTED MANUSCRIPT Title: Review: Ecophysiological perspectives on engineered nanomaterial toxicity in fish and crustaceans

Institute of Biomaterials and Biomedical Engineering

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University of Toronto

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Toronto, ON

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Authors: Neal Ingraham Callaghan1, Tyson James MacCormack2*

Department of Chemistry and Biochemistry

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Mount Allison University

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Sackville, NB

*Corresponding author Tyson MacCormack

Department of Chemistry and Biochemistry Mount Allison University Sackville, NB, E4L 1G8 Canada Phone: 506-364-2371 FAX: 506-364-2313 Email: [email protected]

Keywords: nanotoxicology, nanoparticles, environmental, colloidal behavior, wastewater

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ACCEPTED MANUSCRIPT Abstract Engineered nanomaterials (ENMs) are incorporated into numerous industrial, clinical,

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food, and consumer products and a significant body of evidence is now available on their toxicity to aquatic organisms. Environmental ENM concentrations are difficult to quantify, but

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production and release estimates suggest wastewater treatment plant effluent levels ranging from 10-4 to >101 µg L-1 for the most common formulations by production volume. Bioavailability and

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ENM toxicity are heavily influenced by water quality parameters and the physicochemical properties and resulting colloidal behavior of the particular ENM formulation. ENMs generally

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induce only mild acute toxicity to most adult fish and crustaceans under environmentally relevant exposure scenarios; however, sensitivity may be considerably higher for certain species

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and life stages. In adult animals, aquatic ENM exposure often irritates respiratory and digestive epithelia and causes oxidative stress, which can be associated with cardiovascular dysfunction and the activation of immune responses. Direct interactions between ENMs (or their dissolution

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products) and proteins can also lead to ionoregulatory stress and/or developmental toxicity.

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Chronic and developmental toxicity have been noted for several common ENMs (e.g. TiO2, Ag), however more data is necessary to accurately characterize long term ecological risks. The

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bioavailability of ENMs should be limited in saline waters but toxicity has been observed in marine animals, highlighting a need for more study on possible impacts in estuarine and coastal systems. Nano-enabled advancements in industrial processes like water treatment and remediation could provide significant net benefits to the environment and will likely temper the relatively modest impacts of incidental ENM release and exposure.

1 Engineered nanomaterials Nanoparticles, usually defined as materials having at least one dimension between 1-100 nm, occur naturally (as colloids, sands, dusts, ashes, viruses, etc.) and are manufactured for specific applications (Handy et al., 2011; Klaine et al., 2008). The latter have a wide variety of compositions, including metals (e.g. Ag) and metal oxides (e.g. TiO2), metalloids (Si) and ceramics (e.g. SiO2), carbon-based materials (e.g. carbon nanotubes, dendrimers), and composites (e.g. carbon nanotube metal matrix composites). Dedicated research into nanoscience

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ACCEPTED MANUSCRIPT and nanotechnology has been underway for roughly 40 years, though toxicological research in the area did not begin in earnest until relatively recently and has struggled to keep pace. Engineered nanomaterials (ENMs), the culmination of the former efforts, are now being used in

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a multitude of consumer, clinical, food, and industrial applications (Piccinno et al., 2012; Project on Emerging Nanotechnologies, 2013; Vance et al., 2015). The ENM industry is projected to

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grow to nearly $80 billion by 2022 (RNCOS E-Services Private Limited, 2015). Despite the

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continued proliferation of the industry, most agencies remain in an information gathering phase and few new regulatory measures have been implemented specifically regarding the use, pre-

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release treatment, and release of ENMs and their by-products into the environment. For example; in the USA, there are no unilateral regulations concerning ENM usage and environmental

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release. Depending on the circumstances of use and release, certain ENMs can be regulated under multiple laws including but not limited to the Federal Insecticide, Fungicide, and Rodenticide Act, Safe Drinking Water Act, Clean Water Act, Clean Air Act, Comprehensive

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Environmental Response, Compensation, and Liability Act, Resource Conservation and

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Recovery Act, and various OSHA and NIOSH regulations; carbon nanotubes are also explicitly regulated under the Toxic Substances Control Act (United States Environmental Protection

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Agency, 2014). Existing regulatory frameworks are arguably appropriate for most ENMs, but the rapid increase in use, propagation of novel formulations, and the relative dearth of information on projected environmental burdens and potential for chronic toxicity warrants some level of concern.

Companies are understandably reluctant to release information on the volume of ENMs used in their products or even the rationale for their inclusion (Piccinno et al., 2012). The Project on Emerging Nanotechnologies (www.nanotechproject.org) aims to advance responsible and effective ENM use by tracking all available consumer products using ENMs in their construction. As of 2014, over 1600 consumer products containing some quantity of ENMs have been identified. In ~210 of these products, ENM use is not advertised by the manufacturer in any way. In ~950 products, ENM use is advertised but no documentation is provided on abundance or characteristics, and minor to extensive documentation on quantity or characteristics provided for only ~180 products (Project on Emerging Nanotechnologies, 2013; Vance et al., 2015). This review focuses on the aquatic toxicity of the most common ENM compositions (Project on Emerging Nanotechnologies, 2013), addressed roughly in order of their production levels (Keller 3

ACCEPTED MANUSCRIPT et al., 2013): titanium dioxide (nTiO2), silicon dioxide (nSiO2), iron (nFe) and iron oxide (nFexOy), zinc oxide (nZnO), silver (nAg), carbon nanotubes (CNTs), cerium oxide (nCeO2), and copper (nCu) and copper oxide (nCuO). Our goal is to consolidate ENM release estimates and

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reconcile these data with available information on toxicity to fish and crustaceans. Bioavailability and toxicity will also be discussed in relation to the projected behavior of ENMs

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under environmentally-relevant conditions, where data is available to support such speculations.

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required to characterize potential ecological risks.

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By doing so, we hope to highlight areas where more mechanistic or quantitative insight is

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2 Quantifying ENM release into aquatic systems

Coating products composed of ENMs make up ~42% of total ENM usage, yet contribute 89-97% of all ENM release to water; of this release, up to 25% may pass through a wastewater

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treatment plant (WWTP; Keller et al., 2013). Direct ENM losses to the environment of 0-2% can

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be expected during initial production, with a standard worst-case scenario of up to 6% release to waterways without WWTP processing (Gottschalk and Nowack, 2011). Global ENM release to

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water was recently estimated at between 1,100 and 29,200 metric tons year-1, most of which originates from coatings and cosmetics (Keller et al., 2013). ENMs such as nZnO, nTiO2, and nSiO2 can be used in paints, pigments, and other wear- and weather-resistant coatings, where they are prized for their colours or transparency as per their application, as well as their toughness and resistance to UV damage (Gottschalk and Nowack, 2011; Keller et al., 2013). Other coatings, such as iron and aluminum oxide ENMs, are used to prevent steel corrosion (Keller et al., 2013). From these sources, approximately 90-95% of ENMs present will eventually be released to the environment via the air, water (both through and bypassing WWTPs), and landfills. Greater than 4% of this waste will likely be introduced to waterways without any treatment (Keller et al., 2013). Coatings tend to exhibit high relative levels of environmental release compared to academic/research, medical, and structural applications (including plastics and packaging), which release roughly 20-25% of their constituent ENMs to the environment (Keller et al., 2013). Catalysts, electronics, and similar applications have the lowest projected release estimates of less than 5% (Keller et al., 2013).

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ACCEPTED MANUSCRIPT Environmental ENM concentrations rise with distance down a water system, due to increases in both dumping and flow (Blaser et al., 2008), so levels will not be consistent within a single system. Sedimentary ENM concentrations were projected to be growing exponentially

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over the past 15 years (Gottschalk et al., 2009), especially for nTiO2, nZnO, nAg, and CNTs, although stabilization in the use of certain ENM formulations may lessen the rate of increase to a

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linear level (Blaser et al., 2008). The propensity of ENMs to precipitate means that benthic

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species experience a high relative exposure level compared to pelagic species (Handy et al., 2012, 2011; Ju-Nam and Lead, 2008; Kennedy et al., 2008; Klaine et al., 2008; Mwangi et al.,

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2012), particularly in low flow systems. Biological sequestration of ENMs or their degradation products is also possible, and this may impact the distribution and availability of ENMs in the

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environment. Upper trophic level predators may be exposed to high body burdens via bioaccumulation, although ENM accumulation has only been investigated in a few species and for a limited number of ENM formulations (Angelica and Fong, 2013; Ates et al., 2013;

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Barhoumi and Dewez, 2013; Blinova et al., 2010; Coleman et al., 2013; Mwangi et al., 2012;

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Ramskov et al., 2014; Sha et al., 2015; Wu and Zhou, 2013). Several seminal works estimating ENM production and release are cited herein, but to

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our knowledge, only one (Keller and Lazareva, 2014) uses all forms of release to estimate current levels of ENMs released to a model water system; in this case, San Francisco Bay. Estimated current WWTP effluent concentrations of highly-used ENMs in this system such as nSiO2, nZnO and nFe/ nFexOy are between ~10-1-101 µg L-1, with nTiO2 well into the 101 µg L-1 range. Other common formulations, including nAg, nCeO2, and CNTs have upper bounds at ~101

-100 µg L-1 (Keller and Lazareva, 2014). These estimates are likely to exceed quantities found in

many water systems, since the San Francisco Bay area is heavily impacted by anthropogenic activity as an outlet of multiple waterways servicing industrial centres. However, it is important to note that WWTP effluent will be rapidly diluted when released to the environment, leading to generalized estimates of environmental surface water ENM concentrations on the ng L-1 or fg L-1 scale (Gottschalk et al., 2009). The latter study, while critical in assessing the environmental relevance of ENM exposures, carries two important caveats: Firstly, environmental concentrations of many ENMs are likely increasing at exponential levels, as predicted in the same study (Gottschalk et al., 2009), suggesting that certain ENM levels may be significantly higher than those likely to be encountered in 2009. Secondly, the impact of geography and 5

ACCEPTED MANUSCRIPT proportional contribution of WWTP effluent to a body of water relative to uncontaminated sources were unable to be used to estimate local concentrations in areas of high relative WWTP output (Gottschalk et al., 2009). Therefore, areas of high population and industrial activity (such

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as San Francisco Bay), are likely to exhibit much higher ENM levels than predicted in studies that are unable to incorporate large contributions by WWTP effluent. In contrast, waterways with

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little anthropogenic impact are likely to exhibit much lower ENM burdens. Unfortunately,

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accurate estimates of environmental ENM concentrations are rare, as monitoring studies often provide industrial or WWTP effluent ENM concentrations without accompanying data on total

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flow or flux through the waterway. This has lead to a call for more involved, relevant ENMrelease studies to improve the accuracy of environmental risk assessments (Gottschalk and

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Nowack, 2011). A compiled overview of production and total aquatic release estimates of

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relevant ENMs and their behavior in aqueous systems is provided in Table 1.

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3 The ‘ABCs’: Availability, Bioactivity, and Colloidal stability Environmental inputs, routes of exposure, and relevant factors modulating ENM toxicity

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are summarized in Figure 1. ENMs generally form colloidal suspensions in aquatic systems and as such, their bioavailability and bioactivity are not strictly concentration dependent. The colloidal behavior and subsequent toxicity of ENMs is dictated by their intrinsic physicochemical properties (size, shape, coating, composition), environmental conditions (pH, salinity, temperature, light availability, dissolved (in)organics), and the route and duration of exposure (Baker et al., 2016; Blinova et al., 2010; Furtado et al., 2016, 2015, 2014; Handy et al., 2012; Heinlaan et al., 2008; Klaine et al., 2008; O’Brien and Cummins, 2010; Stark et al., 2015; Sun et al., 2016; Z. Wang et al., 2015). A comprehensive review of the physicochemical factors influencing ENM behaviour and fate in aquatic environments can be found elsewhere (Peijnenburg et al., 2015). Although many have called for standardized nanotoxicology testing, the heterogeneous behaviour of ENMs and diversity of relevant toxicity metrics have been obstacles in developing a unified assessment framework. As of 2016, the International Organization for Standardization is still preparing recommendations for such standardization (International Organization for Standardization, 2016). Few nanotoxicity studies report all pertinent physicochemical parameters, or for logistical reasons, characterize their ENMs in 6

ACCEPTED MANUSCRIPT different conditions from those used experimentally, making between-study comparisons problematic. Dynamic changes in colloidal stability and/or ENM behavior throughout the exposure are also critically important, but challenging to quantify under experimental conditions.

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Supplemental information of any kind on aggregation kinetics, partitioning in the experimental system, or degradation/dissolution is often key to explaining variation between studies or

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between apparently similar ENM formulations.

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Protocols relating to the preparation and delivery of ENMs for aquatic toxicity testing also warrant discussion when considering between-study variability. Many studies, including our

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own, use some form of sonication in the preparation of ENM stocks. The goal is typically to disperse aggregates and ensure accurate nominal exposure concentrations; however, depending

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on the formulation, sonication can actually enhance aggregation (Handy et al., 2012; Taurozzi et al., 2011) or destroy ENMs and promote dissolution (Handy et al., 2011). The mechanism of ENM delivery and rate of introduction into an experimental system can have equally important

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impacts on effective exposure doses. Continuous, low-level inputs more effectively disperse

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2016), as discussed below.

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ENMs, potentially by facilitating particle stabilization by dissolved organic carbon (Baker et al.,

Humic substances impact ENM aggregation mechanisms, and can either stabilize or destabilize ENMs in suspension, depending on other water conditions (Figure 1). Generally, humic matter seems to decrease ENM toxicity (Blinova et al., 2010; Gao et al., 2012; Gottschalk and Nowack, 2011; Handy et al., 2012; Ju-Nam and Lead, 2008; Keller et al., 2013; Kennedy et al., 2008; Klaine et al., 2008; Zhu et al., 2009). Seasonal variations in water quality, including the concentration and composition of humic matter species will therefore differentially affect the bioavailability of individual ENM formulations and their resulting degradation products (Stolpe and Hassellöv, 2010). Salinity has pervasive effects on colloidal stability and even low-strength, diluted seawater can substantially enhance ENM precipitation (Kennedy et al., 2008; Stolpe and Hassellöv, 2007; Wehmas et al., 2015; Wong et al., 2010). Very few studies have addressed ENM effects in saline waters but toxicity has been noted in brackish water fish (Della Torre et al., 2015) and marine invertebrates (e.g., Rocco et al., 2015) so clearly additional attention is required. In coastal and estuarine systems experiencing frequent changes in salinity, the bioavailability and potential impacts of ENMs may be particularly difficult to predict. 7

ACCEPTED MANUSCRIPT Finally, different model organisms or their distinct life stages may experience different routes of ENM exposure based on their morphology and interactions with their environment (Figure 1). For example, zebrafish embryos have chorionic pores that are well above the size of

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ENMs at 0.5-0.7 µm (Lee et al., 2009), and may allow their passage in and out of the embryo (Auffan et al., 2014; Lee et al., 2009). In other studies, the chorion has provided an effective

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barrier against certain ENM species (Kim and Tanguay, 2014), although excessive chorionic

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ENM adsorption may block nutrient passage and thus exert indirect toxicity (Duan et al., 2013). Zinc metalloproteases involved in breaking down the chorion membrane may also be impacted

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by ENMs entering through the chorionic pores, delaying or completely preventing hatching, although significant effects were mostly observed at well above environmentally-relevant doses

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interaction, ingestion, and respiration.

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(Ong et al., 2014b). In adult fish, on the other hand, exposure will be largely limited to dermal

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4 Toxicity of environmentally-relevant ENM formulations The bioactivity of ENMs frequently stems from the promotion of ROS production or

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from direct interactions between ENMs (or their component parts) and proteins or membranes, as seen in Figure 2 (Ali et al., 2015; Bessemer et al., 2015; Callaghan et al., 2016; Canesi and Corsi, 2015; Dieni et al., 2013, 2014; Gormley et al., 2016; Handy et al., 2011; Heinlaan et al., 2008; Remya et al., 2015; Rundle et al., 2016; Singh et al., 2010; Wehmas et al., 2015). These can lead to secondary responses, which can include tissue remodeling (Bessemer et al., 2015; Callaghan et al., 2016; Hao et al., 2009; Singh et al., 2010; T. Wang et al., 2015; Wu and Zhou, 2013), activation of various immune system components (Della Torre et al., 2015; Hussain et al., 2012; Kononenko et al., 2015; Luo et al., 2015), and subsequent decreases in scope for physiological performance (Callaghan et al., 2016; Powers et al., 2011). Due to the heterogeneity of ENM behaviour based on constituent material, surface coating, and the surrounding medium, as well as the disparity of ENM characterization between studies, it is difficult to directly compare the toxicity of different formulations. Instead, we will individually examine eight of the most commonly produced and released ENMs to evaluate their potential for both acute and subacute toxicity, in addition to complicating or mitigating factors that should be considered.

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ACCEPTED MANUSCRIPT 4.1 Titanium dioxide Prized for their structural and spectral properties, nTiO2 are produced in large quantities,

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c.a. 88 kt yr-1 based on 2010 estimates (Keller et al., 2013), in anatase and rutile mineral forms, or mixtures thereof. Both polymorphs have approximately the same density and hardness, but

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their crystal structures vary. The rutile form is more common, but anatase shows a trend of higher toxicity and proclivity to accumulate in tissues (Sha et al., 2015). nTiO2 is used in paints,

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pigments and other coatings, sunscreens, filtration applications, plastics, and cements (Piccinno et al., 2012; Project on Emerging Nanotechnologies, 2013; Vance et al., 2015). As likely one of

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the two most highly-produced ENMs, estimated aquatic release rates are very high relative to most other ENM formulations, on the order of 105 tons per year worldwide (Keller et al., 2013).

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In a highly-polluted water system such as the San Francisco Bay, effluent levels likely approach or exceed 50 µg L-1 (Keller and Lazareva, 2014).

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The toxic potential of nTiO2 has been debated but several studies have demonstrated

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bioactivity at relatively high doses, at least to fish. Respiratory distress and changes to metal metabolism were observed in rainbow trout chronically (14 day) exposed to 0.1-1.0 mg L-1

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nTiO2 of unspecified surface coating (Federici et al., 2007). Rainbow trout also showed kidneyand brain-localized oxidative stress when exposed to 1.0 mg L-1 25 nm uncoated nTiO2, but not a bulk control (Boyle et al., 2013). In the latter study, both nanoscale and bulk forms induced gill histopathology, and the nanoscale form especially affected swimming behaviour (Boyle et al., 2013), ostensibly due to a hypoxic response and/or cerebral damage as evidenced by brainspecific markers of oxidative stress and abnormal vasculature. Unlike a number of other ENM formulations, nTiO2 on its own does not appear to be particularly harmful to crustaceans. Neither nanoscale nor bulk TiO2 was acutely toxic at any environmentally-relevant concentration in Daphnia magna and Thamnocephalus platyurus (LC50s above 20 g L-1) (Heinlaan et al., 2008). Several other Daphnia studies with nTiO2 show similar results (reviewed in O’Brien and Cummins, 2010). There are mixed results concerning the developmental toxicity of nTiO2 as well. In one study, neither nanoscale (≤ 20 nm uncoated) nor bulk TiO2 affected the 96 h survival rate of zebrafish embryos (Zhu et al., 2008); however, a separate 120 h incubation found minor developmental anomalies resulting from nTiO2 exposure (Zhu et al., 2008). In another study, one 9

ACCEPTED MANUSCRIPT formulation of nTiO2 (10.5-11.6 nm uncoated) induced more jaw and snout malformations than the corresponding bulk TiO2 control, but this was not the case for all nTiO2 formulations tested (Wehmas et al., 2015). These effects were ablated when ultrapure water was replaced with

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embryonic media as the dispersant, likely due to ENMs complexing with organic and inorganic matter (Wehmas et al., 2015). Furthermore, and most alarmingly, environmentally-relevant

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concentrations (5-40 µg L-1) of 5-6 nm nTiO2 dispersed with 0.5% w/v

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hydroxypropylmethylcellulose K4M over 45 days delayed body growth, increased relative heart and liver size, and decreased relative brain size in zebrafish (Sheng et al., 2016). Sequestration of

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Ti in the liver, heart, and brain increased proportionally with dose, and decreased locomotor activity was observed in correlation with increasing glial cell proliferation. Similarly, there was a

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decrease in the monoamine transmitters norepinephrine, dopamine, and serotonin with increasing dose. Transcriptome analysis revealed activation of various repair and memory compensatory processes in the brain (Sheng et al., 2016). Together, these findings suggest that long-term,

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subacute exposure to environmentally-relevant concentrations of nTiO2 can lead to impaired

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swimming behaviour, likely due to neurotoxic effects. The studies by Zhu et al. (2008), Wehmas et al. (2015), and Sheng et al. (2016) used similar nTiO2 sizes but different exposure times, and

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their combined results emphasize the importance of chronic exposure studies; the duration of nTiO2 exposure may be critical in determining its developmental toxicity (at least in this model). Acute studies showing few effects with environmentally-relevant doses of nTiO2 may miss indications of severe toxicity associated with chronic exposures. nTiO2, along with other ENM formulations such as nZnO, is used extensively in paints and cosmetics such as sunscreen for its ability to lower the energy of inbound ultraviolet (UV) radiation. It is not surprising then, that nTiO2 toxicity can be greatly potentiated by UV light, as the energy absorbed by the ENM promotes ROS generation and physical interactions. In simulated natural light, LD50s for zebrafish exposed to 21 nm uncoated nTiO2 decreased roughly 2 orders of magnitude compared to those measured under standard laboratory lighting (Ma et al., 2012). The impact of UV light was even more dramatic for Daphnia magna, decreasing the LD50 by ~5 orders of magnitude to 30 µg L-1, well within estimates for potential environmental loads (Keller and Lazareva, 2014). UV light is rapidly attenuated with depth but these results could have profound implications for surface dwelling organisms.

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ACCEPTED MANUSCRIPT Acute exposures to nTiO2 at environmentally-relevant concentrations seem unlikely to elicit severe toxicity on their own. However, if toxicity is potentiated with UV light, organisms near the water surface may be negatively impacted by environmentally-relevant concentrations

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of nTiO2 (Figure 1). Evidence of chronic and developmental nTiO2 toxicity at sublethal concentrations merits continued investigation. As we highlight throughout this review, such sub-

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acute ENM toxicity may reduce the capacity of organisms to cope with routine environmental

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stressors. The ecological effects of nTiO2 exposure may therefore manifest as increases in sensitivity to temperature fluctuations or in oxygen or food availability (Figure 2).

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4.2 Silicon dioxide

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Silica nanomaterials (nSiO2) are used in various coatings, pigments, cosmetics, and textiles (Project on Emerging Nanotechnologies, 2013; Vance et al., 2015) and may be one of the most heavily-produced ENMs, with effluent concentration estimates on the order of 10-1-101 µg

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L-1 in San Francisco Bay (Keller and Lazareva, 2014). Few quantitative or mechanistic

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toxicological studies have been carried out on nSiO2, and most of the available literature on the subject is presented in a previous review (Fruijtier-Pölloth, 2012). The relative dearth of studies

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on nSiO2 likely stems from evidence that bulk silica exhibits essentially no toxicity (Martin, 2007). Briefly, the mechanism of bioactivity likely relates to the adsorption of nSiO2 to cells, followed by surface protein abstraction and loss of cell membrane integrity. Crustaceans and zebrafish are minimally affected by medium-to large-diameter nSiO2 spheres (50-200 nm), but small-diameter (≤10 nm) spherical formulations and high aspect ratio nanowires may present a moderate risk (Fruijtier-Pölloth, 2012) due to greater surface area-to-mass ratios. Both algae and bacteria experience mortality and growth inhibition when exposed to various forms of nSiO2 and toxicity is inversely proportional to ENM diameter (Jiang et al., 2009; Van Hoecke et al., 2008). Although most nSiO2 formulations do not appear to exhibit substantial acute toxicity, its relatively high rate of release to the environment and evidence of bioactivity in certain formulations warrants additional study. 4.3 Iron and iron oxides Nanoscale iron (nFe) and iron oxide (nFexOy) are used in a variety of structural and coating applications and as catalysts in environmental remediation and detoxification 11

ACCEPTED MANUSCRIPT applications (Blaise et al., 2008; Pawlett et al., 2013; Project on Emerging Nanotechnologies, 2013; Vance et al., 2015). As with nSiO2, there is relatively little literature assessing aquatic

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toxicity, especially at predicted effluent levels of 100-101 µg L-1 (Keller and Lazareva, 2014). Various iron alloy oxides exert their median toxicity (as assessed by a variety of metrics)

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between 102-104 µg L-1 in fish and aquatic invertebrates (Blaise et al., 2008). Magnetic, uncoated nFe2O3 between 33-90 nm showed LC50s ranging from 50-90 µg L-1 in algae and 21-66 mg L-1

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in Daphnia (depending on ENM age), while coating with dimercaptosuccinate reduced acute toxicity (Y.-Q. Zhang et al., 2016). In terms of developmental toxicity, very high concentrations

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(≥50 mg L-1) of 30 nm uncoated nFe2O3 induce malformations in zebrafish embryos (tissue presumably below (Zhu et al., 2012).

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ulceration, pericardial edema, body arcuation) with seemingly no effects at 10 mg L-1, or

The precise toxicological mechanism of nFe and its oxides seems to be largely oxidative,

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although different species likely have different mechanistic profiles and the potential for physical

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interaction between ENMs and biological surfaces remains. nFe2O3 (Karlsson et al., 2008), nFe3O4, nCo0.2Zn0.8Fe2O4, and nCo0.5Zn0.5Fe2O4 (Barhoumi and Dewez, 2013) all induce

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oxidative damage at high concentrations, however the significance of these results at nearenvironmental levels remains unclear. One of the few mechanistic toxicology studies in fish examined ionoregulation in Indian major carp (Labeo rohita) exposed to a chronic but environmentally-irrelevant concentration of 5 x 105 µg L-1 nFe2O3 (100 x 200 nm spindles). Over up to 25 days, nFe2O3 induced dysregulation of Na+ and K+ levels and a large increase in Na+/K+-ATPase (NKA) activity (Remya et al., 2015). This response is similar to that observed in a study using nZnO (Bessemer et al., 2015), and we suggest that ENMs damaged the gill membrane and triggered a compensatory increase in NKA expression. Whether this response would occur at an environmentally-relevant concentration remains to be seen. Environmental remediation applications incorporating nFe present unique exposure scenarios to aquatic organisms, where high concentrations of ENMs are directly applied to a water system in a short time. Under such conditions, point source nFe concentrations may approach or exceed those shown to elicit toxicity in fish and crustaceans (Ju-Nam and Lead, 2008; Liu et al., 2014; Oberdörster et al., 2005; Pawlett et al., 2013). Given that remediation

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ACCEPTED MANUSCRIPT treatments are applied strictly in environments with pre-existing contaminant loads, the potential for additional modest nFe toxicity is unlikely to be a concern.

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4.4 Zinc oxide nZnO is widely exploited for its desirable photochemical and semiconductor properties.

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Applications include sunscreens, paints, cosmetics, textiles, environmental remediation catalysts,

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and electronics (Project on Emerging Nanotechnologies, 2013; Vance et al., 2015). Effluent levels of nZnO in San Francisco Bay are estimated at between 1-10 µg L-1 (Keller and Lazareva,

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2014). Certain formulations readily dissolve in high-organic content media (Xia et al., 2008) but others exhibit little to no dissolution in freshwater (Bessemer et al., 2015). Such observations

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highlight the need for extensive ENM characterization data in explaining associated toxicity responses. nZnO induces oxidative stress in cells (Xia et al., 2008), at least partially through cellmediated mechanisms (Dieni et al., 2014; Gormley et al., 2016), although these studies have

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used doses well above those to be expected in environmental settings. In the white sucker

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(Catostomus commersonii), acute nZnO exposure at high concentrations (1-10 mg L-1) induces gill damage and remodeling processes, lowering aerobic scope and triggering bradycardia,

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potentially via changes in acetylcholine signalling (Bessemer et al., 2015; Callaghan et al.,

There is conflicting evidence as to whether or not there is a nano-specific effect of ZnO toxicity. Exposure to nZnO, bulk ZnO, or the ionic Zn2+ source ZnSO4*7H2O elicited similar responses in terms of EC values in crustaceans (Heinlaan et al., 2008). However, 4 and 40 mg L-1 bulk ZnO and 14 mg L-1 bulk ZnSO4 both tended to increase metallothionein I and II, superoxide dismutase, and HSP70 expression relative to either 4 or 40 mg L-1 (3.2 or 32 mg L-1 Zn respectively) nZnO in whole-body protein extracts of medaka (Oryzias melastigma) after 4 day exposures. This occurred despite the bulk forms releasing <50% of the Zn2+ that nZnO did (Wong et al., 2010). Wong et al. (2010) also found a lower 96 h LD50 for nZnO than ZnSO4*7H2O in the copepod Tigriopus japonicas but no difference in the amphipod Elasmopus rapax. The authors attributed the toxic effects of nZnO to Zn2+ release (Wong et al., 2010), although the mechanisms discussed previously may also be applicable.

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ACCEPTED MANUSCRIPT Embryological toxicity of ZnO strongly suggests additional effects at the nanoscale. The hatch rate of zebrafish embryos is much lower when exposed to nZnO (beginning above ~1 mg L-1) than an equivalent concentration of Zn2+. At environmentally-relevant levels (0.1-0.5 mg L), tissue ulceration was evident in 10-15% of embryos by 144 h post-fertilization, with incidence

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and severity increasing proportionally with dose (Zhu et al., 2008). In contrast, high levels (~2.0-

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50 mg L-1) of nZnO or Zn2+ both elicited similar rates of mortality and malformations in

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zebrafish embryos (Wehmas et al., 2015). As with other ENMs, the use of embryonic media attenuated these effects due to complexing and/or precipitation of Zn species. The nZnO

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formulation dissolved readily in water, suggesting that toxicity was mainly due to Zn2+ release. Most evidence suggests that nZnO exhibits little toxicity in the range of concentrations

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currently present in the environment, but formulation-specific differences in particle stability and behaviour make it difficult to draw firm conclusions on risk. The acute impacts of nZnO on hatching and cardiorespiratory function in fish, albeit at high doses, merit some concern

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regarding chronic exposures. The presence of dissolved and suspended organic matter in

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environmental conditions will undoubtedly mitigate some bioactivity (Figure 1), but additional

4.5 Silver

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data on chronic toxicity under environmentally-relevant exposure scenarios would be valuable.

nAg is used in myriad coating applications, including medical devices, textiles, and filters for its antibacterial and antifungal properties (Project on Emerging Nanotechnologies, 2013; Vance et al., 2015). It does not seem to exhibit clear patterns of precipitation relating to size or functionalization (e.g., PVP-coated vs. uncoated) (Coleman et al., 2013). In San Francisco Bay, WWTP effluent nAg levels have been estimated on the order of 10-2-10-1 µg L-1 (Keller and Lazareva, 2014). Ionic Ag+ is a potent and well characterized ionoregulatory toxin in freshwater, and to some extent, marine fishes (Wood et al., 2012), so nAg dissolution analysis is important for identifying nano-specific effects (Shaw and Handy, 2011). Ag+ from nAg dissolution may only be present at low levels in ecologically-relevant conditions due to silver’s extensive capacity for thiol binding (Blaser et al., 2008). There is considerable uncertainty regarding the kinetics of nAg dissolution in the environment but the rate of Ag+ release will depend greatly on dissolved oxygen availability and other water chemistry, as well as the physicochemical characteristics of the nAg formulation (McGillicuddy et al., 2017; Zhang et al., 2016). As with 14

ACCEPTED MANUSCRIPT other ENMs, the disposition of nAg can be expected to change when exposed to dynamic environmental conditions (Furtado et al., 2016). For example, 20 nm PVP-coated nAg toxicity was ameliorated (LD50) by 39-fold in a crustacean (Chydorus sphaericus) and by up to 27-fold

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in zebrafish larvae with the addition of humic matter (Wang et al., 2015b). The authors suggest this is mainly due to the prevention of Ag+ release, and that this effect suggests a prolonged risk

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period for any nano-specific interactions (Wang et al., 2015b). In a boreal lake setting using large

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quantities of nAg of multiple coatings and extended measurement periods, the majority of nAg either dissolved or precipitated into the sediment (c.a.10-40% of total nAg added depending on

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starting concentration) or planktonic layers (c.a. 6-18%); the remainder was stabilized in the water column by organic matter but chelation maintained free Ag+ levels well below thresholds

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for toxicity or even detection (Furtado et al., 2015, 2014). Together, these results suggest the potential impact of nAg and its degradation products will be tempered considerably under natural conditions and that toxicity will not be a major concern.

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A number of thorough reviews have directly addressed the aquatic toxicity of nAg (e.g.

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McGillicuddy et al., 2017; Sharma, 2013; Walters et al., 2014), so we only focus on few examples to illustrate key points. After a 14-day nAg exposure, medaka experienced dose-

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dependent (0.05-0.5 mg L-1 30 nm PVP-coated) decreases in antioxidant capacity and increased histopathological evidence of oxidative damage to gill (Wu and Zhou, 2013). Sodium transport was inhibited in a nano-specific manner in juvenile rainbow trout exposed to 1.0 mg L−1 of 5-40 nm citrate-capped nAg (Schultz et al., 2012), suggesting the potential for disruptions in osmotic homeostasis. Moreover, significant whole-body Ag sequestering resulted from exposure to nAg but not Ag+ in tissues of both juvenile and adult sheepshead minnows (Griffitt et al., 2012). It is unlikely that such clear responses will be observed in the wild, where nAg concentrations will be at least an order of magnitude lower, the route of exposure will vary (respiratory and/or dietary), and the presence of dissolved organic matter and UV light (see below) will influence nAg behaviour in the system (Figure 1). Zebrafish embryo mortality resulting from nAg infiltration through chorion pores increased with exposure concentration (c.a. 0.21-3.6 mg L-1 of 1.6 nm nAg) (Lee et al., 2009). Interestingly, malformation incidence peaked at c.a. 40% at 0.98 mg L-1 nAg, and consisted mainly of finfold and tail/spinal cord abnormalities, edema of the head and yolk sac, and cardiac 15

ACCEPTED MANUSCRIPT and eye malformations (Lee et al., 2009). At high concentrations (0.32-10.7 mg L-1), 10 nm citrate-coated nAg induced lower levels of mortality than exposure to an equivalent concentration of Ag+ (Powers et al., 2011), which was also observed when comparing Ag+ to 3

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nm citrate-coated nAg (Auffan et al., 2014). Increasing water salinity markedly decreased chorionic sequestration of nAg, while uptake by the remainder of the embryo was unaffected

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(Auffan et al., 2014). The coating of nAg seems to affect toxicity, as both citrate and fulvic acid

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decreased the mortality of zebrafish embryos exposed to a range of nAg concentrations, likely due to decreasing ionic release (Osborne et al., 2013). 4 and 10 nm oleic acid-coated nAg

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decreased heart rate in zebrafish embryos after 72 h post fertilization proportionally with dose (0.48-3.85 mg L-1) (Xin et al., 2015). Finally, the embryonic uptake of nAg is inversely

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proportional to the diameter of nAg, and has been shown to cause multiple developmental defects and the activation of metal detoxification genes (Xin et al., 2015). The mortality observed in embryos likely has an oxidative component as observed in mature fish, although the

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possibility of direct toxicity at the ENM-biological interface should also be considered.

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In adult Daphnia pulex, ionic Ag+ is over two orders of magnitude more toxic than the nanoparticulate form (20-30 nm, unspecified coating) in terms of 48 h LC50 (Griffitt et al.,

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2008). As there was low dissolution of nAg (0.07% by mass) in this study, it is likely that the nanoparticulate form must impart the majority of the toxicity observed in both D. pulex and zebrafish (Griffitt et al., 2008). In D. pulex, these 48 h LC50s for nAg are at least 2 orders of magnitude higher than the upper range of expected nAg burdens in highly-contaminated areas such as San Francisco Bay. Similarly in zebrafish, 48 h LC50s for nAg are c.a. 4 orders of magnitude higher than the expected upper limits in the environment (Keller and Lazareva, 2014). Given the range of sublethal chronic and developmental responses observed in both fish and crustaceans, it is very likely that subacute toxicity is a much greater concern than lethal exposures in the environment. As with nTiO2, nAg may be susceptible to photophysical effects in the environment (Figure 1). When exposed to light from an illuminating incubator, 9 nm uncoated nAg quickly agglomerated, and as a result exerted lower toxicity to the protozoan Tetrahymena pyriformis, possibly by slowing Ag+ release (Shi et al., 2012). Further investigations are necessary to determine the potential environmental implications of interactions between nAg and light energy. 16

ACCEPTED MANUSCRIPT The environmental risks posed by nAg will likely depend largely on particle dissolution and the release of Ag+. Several studies highlighted above note nano-specific toxicity of nAg, but most employed high concentrations in acute exposure regimes. Whole-lake nAg addition studies

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have illustrated that bioavailability is limited by the same mechanisms which limit the availability, and hence toxicity, of free Ag+ (Furtado et al., 2015, 2014). As with other ENM

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formulations, it is possible that significant toxicity may be observed in extended exposures or

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under a specific set of conditions (Figure 2), but overall risks appear minimal.

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4.6 Carbon nanotubes

Single- and multi-walled carbon nanotubes (SWCNTs and MWCNTs, respectively) are

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used in a wide variety of applications for their unique mechanical, thermal, and electrical properties (Piccinno et al., 2012; Project on Emerging Nanotechnologies, 2013; Vance et al., 2015). Estimates for environmental concentrations from pollution are low, with effluent levels on

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the order of 10-3-10-1 µg L-1 in San Francisco Bay (Keller and Lazareva, 2014). An assortment of

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toxic mechanisms have been proposed for both SWCNTs and MWCNTs, which seem to have similar magnitudes of toxicity for a range of aquatic invertebrates (Mwangi et al., 2012).

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Mechanistic studies for CNTs abound, but there is a paucity of chronic exposure data for fish and crustaceans using realistic concentrations. DNA damage and ROS generation have been observed in a freshwater snail exposed to SWCNTs for 24-96 h at superenvironmental concentrations (50-460 µg L-1; (Ali et al., 2015), so toxicity may be a concern in other species. An important consideration in the design and administration of such studies is the rapid precipitation of suspended CNTs in environmental conditions (Kennedy et al., 2008). The functionalization and coating of CNTs likely has a large effect on their toxicity. A study comparing SWCNTs coated with SDS to those coated with gum Arabic showed the latter coating increased the 48 h LD50 15-fold (to 0.27 mg L-1), which the authors attributed to the antioxidant properties of gum Arabic (Gao et al., 2012). However, the differences in colloidal stability between the formulations were neither measured nor discussed, so further study is necessary to fully explain this phenomenon. A study examining very high levels (100 mg L-1) of unfunctionalized, hydroxylated, and carboxylated MWCNTs found that the two functionalized formulations, unlike the unmodified control, could not be stabilized in either MilliQ or 20% seawater. Upon addition of 100 mg L-1 humic matter, the stable suspended fraction increased by 17

ACCEPTED MANUSCRIPT 3-4 fold in all three formulations, and markedly decreased toxicity (Kennedy et al., 2008). Together, these results suggest that any environmentally-relevant concentrations of MWCNTs in a realistic setting will ultimately precipitate. Moreover, even if they remain suspended due to the

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stabilizing effects of humic matter, their toxic effects may be dramatically decreased, as seen in other ENMs (Blinova et al., 2010; Gottschalk et al., 2009; Jang et al., 2015; Ju-Nam and Lead,

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2008; Kennedy et al., 2008; Klaine et al., 2008; Stolpe and Hassellöv, 2010, 2007; Taurozzi et

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al., 2011; Z. Wang et al., 2015).

A final factor worthy of consideration with regard to CNT toxicity is the presence of

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other contaminants; many CNT formulations used in research and industry can carry a nonnegligible fraction of metals or other contaminants. Severe toxicity in amphipods (Hyalella

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azteca), midges (Chironomus dilutes), mussels (Villosa iris), and oligochaetes (Lumbriculus variegatus) exposed to 1000 mg L-1 MWCNTs over 14 days was greatly mitigated (87% increased survival in H. azteca, 5% decreased survival in C. dilutes, and 63% increased survival

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in V. iris) when ENMs were first washed with nitric acid to remove contaminating metals

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(Mwangi et al., 2012). The environmental relevance of these results are questionable, however, given the high concentration of MWCNTs employed in the study. It is worth noting, however,

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that in one case, Cu, Cr, Fe, Mo, and Ni comprised over 13% of the ENM sample by mass (Mwangi et al., 2012). 4.7 Cerium oxide

Cerium oxide nanoparticles (nCeO2) are used in coatings, fuel additives, and automotive catalytic converters (Piccinno et al., 2012; Project on Emerging Nanotechnologies, 2013; Vance et al., 2015) but relatively few studies have addressed their toxicity in aquatic organisms. Effluent nCeO2 levels on the order of 10-2-100 µg L-1 are expected in San Francisco Bay (Keller and Lazareva, 2014). nCeO2 are unusual compared to many other redox-active ENMs, in that they appear to be anti- rather than pro-oxidant in physiological and environmental systems (Rundle et al., 2016; Xia et al., 2008). Despite this, it is likely that nCeO2 will still exert some form of toxicity to aquatic organisms. In white sucker, exposure to 25 nm uncoated nCeO2 (1 mg L-1 for 25 h) raised cortisol levels slightly, but had no effect on heart rate, red blood cells, or various markers

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ACCEPTED MANUSCRIPT of tissue damage (Rundle et al., 2016), unlike nZnO in a similar exposure regimen (Bessemer et al., 2015; Callaghan et al., 2016; Dieni et al., 2014). Relative to bulk Ce, minor developmental defects were observed in zebrafish embryos exposed to uncoated nCeO2 (2.8-2.9 nm) for 120 h

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in water but not in embryo medium, including yolk sac edema and developmental delays (Wehmas et al., 2015). Finally, a study using nCeO2 from 3 different suppliers (10-30, 15-30,

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and 70 nm) found remarkably different toxicities to marine bacteria and algae depending on the

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source (Leung et al., 2015). The disparity in effects may result from differences in coating and size between suppliers. Additional work is clearly required to characterize the potential

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environmental impacts of this rather common ENM.

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4.8 Copper and copper oxide

The toxicity of copper is well-known, with copper salts frequently used as an aquatic biocide, and copper fixtures often identified as the cause for unexpected mortality in aquarium

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systems. Copper and its oxide have a wide variety of ENM applications in cosmetics, paint,

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filtration, electronics, and textiles (Project on Emerging Nanotechnologies, 2013; Vance et al., 2015) due to favourable physical and conductive properties. The effluent concentration of

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copper-based ENMs in San Francisco Bay is in the range of 10-4-5x10-2 µg L-1 (Keller and Lazareva, 2014). The toxic effects of nanoscale copper (nCu) and copper oxide (nCuO) seem to be highly dependent on the form of the ENM (Hua et al., 2014; Ramskov et al., 2014) and the properties of the dispersant; nCu behavior is very susceptible to minor differences in water composition, even between freshwater varieties, as evidenced by differences in invertebrate LC50s (Blinova et al., 2010). The surface functionalization of nCu(O) species may also impact toxicity, but the heterogeneity of ENM characterization data in the literature complicates metaanalyses. Similar animal species can also show remarkable differences in sensitivity to nCu, while maintaining a similar sensitivity to other forms of Cu. For example; the crustaceans Daphnia magna and Thamnocephalus platyurus respond similarly to CuSO4 and bulk CuO, but are likely to display different sensitivity to nCuO (Heinlaan et al., 2008). There may also be an age-dependent component of susceptibility; in adult Daphnia pulex, copper ions are ~0.5 orders of magnitude more toxic than nCu while in Ceriodaphnia dubia neonates, copper ions and nCu showed LD50s in the same order of magnitude (Griffitt et al., 2008). The low-end of LC50s for nCu in these crustaceans are about an order of magnitude over the upper range of expected nCu 19

ACCEPTED MANUSCRIPT levels in effluents entering San Francisco Bay (Keller and Lazareva, 2014), suggesting that acute toxicity is not an immediate concern. Additional chronic toxicity studies using environmentallyrelevant exposure regimes would be valuable to address concerns over potential subacute effects.

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For zebrafish, 48 h LC50s for nCu are over 4 orders of magnitude higher than the highest expected environmental concentrations (Griffitt et al., 2008). Despite this, specific nCu

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formulations may exhibit toxicity at concentrations well within predicted environmental ranges

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(see below).

In zebrafish embryos, 25 nm nCu with an unspecified coating showed a lower LC50 than

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50 or 100 nm forms, but were still less toxic than 400 nm microparticles (bulk scale) or ionic Cu(NO3)2 (Hua et al., 2014). Hatch rate decreased and mortality increased 24-120 h post

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fertilization in proportion with nCu concentration, albeit at a concentration range (0.25-8 mg L-1) well above predicted environmental levels. There was a trend of decreased embryonic translocation across exposures to all forms of Cu, but this effect was greatest in 25 nm nCu,

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followed by ionic Cu(NO3)2 and then all other, larger nCu forms. As gills are not active in

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zebrafish until ~14 d post-fertilization, the authors attribute these behavioural effects to Cuinduced developmental defects, as opposed to respiratory distress (Hua et al., 2014). Although

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not ENM-specific, this interpretation is supported by behavioural perturbations such as impaired circadian rhythm and swimming performance in rainbow trout exposed to dietary Cu (Campbell et al., 2002; Handy et al., 1999). Further work is necessary to determine whether there is a nanospecific effect, or if dissolved Cu2+ alone accounts for these findings. The study by Hua et al. (2014) represents one of the most thorough and valuable comparisons of multiple nCu forms to bulk and ionic controls, but concentrations are still 3 orders of magnitude higher than the upper limits expected for nCu or nCuO levels in highly-polluted waters (Keller and Lazareva, 2014). In general, it seems that acute nCu(O) toxicity may be a potential environmental concern for crustaceans, as LC50 values are within an order of magnitude of predicted wastewater concentrations, while chronic and developmental toxicity are more relevant concerns for fish. There is currently a significant gap in our knowledge of nCu(O) toxicity to brackish and marine fish however, and a few studies have noted bioactivity in these animals (Wang et al., 2015a), albeit at high concentrations (20 µg L-1). As noted in previous sections, additional work is

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ACCEPTED MANUSCRIPT required to assess the potential for chronic toxicity under realistic exposure scenarios,

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particularly in saline waters.

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5 Gaps and considerations

Our understanding of the aquatic toxicity of ENMs has grown considerably in the past

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decade but a number of key areas still require additional work. The majority of the mechanisms of ENM toxicity highlighted in Figure 2 were identified and characterized using short-term

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exposures to high concentrations of ENMs under tightly controlled conditions in the laboratory. Although an important first step, such studies fail to accurately represent the chronic, low dose

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exposures (µg L-1 or ng L-1) or dynamic environmental conditions likely to be encountered in natural systems. Chronic exposures may induce subtle, complex effects in animals not apparent in acute studies, including increased susceptibility to infection and risks to reproduction and

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development (Figure 2), both of which have been noted in fish exposed to ENMs (Jovanović et

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al., 2011; Ong et al., 2014b). Considerable variability is also observed in the toxicity of a single ENM formulation between species, or even different life stages of the same species (Griffitt et

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al., 2012, 2008; Heinlaan et al., 2008; Ma et al., 2012; Manusadžianas et al., 2012; Wong et al., 2010). Additional chronic, environmentally-relevant exposures with multiple species and life stages and a shift in focus from acute LC50s to EC50 data on relevant biomarkers will help to contextualize the true long term environmental risks of ENMs. Understanding the long term risks of ENMs to aquatic organisms will also require a better understanding of how sensitive histological and biochemical indicators of ENM toxicity translate into impacts on whole animal performance (Figure 2). Acute exposures to high concentrations of certain ENMs can impact swimming behavior (Boyle et al., 2013) or increase sensitivity to hypoxia (Bilberg et al., 2010) in a few species of freshwater fish but data for such environmentally-relevant endpoints is sparse. Damage to the gill epithelium is often observed following ENM exposures, so reductions in swimming performance and aerobic scope may be common, although rarely measured. The subtle epithelial remodelling observed in the gills of some fish (Callaghan et al., 2016) may even facilitate increases in swimming performance and aerobic scope in species which limit gill surface area under well-oxygenated conditions (e.g.

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ACCEPTED MANUSCRIPT through loss of the interlamellar cell mass; Nilsson et al., 2012). Characterizing how biochemical and histological measures of subacute ENM toxicity affect whole animal performance will allow us to better predict potential long term impacts on wild populations. The logical extension of this

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is to look at how routine environmental stressors (e.g. fluctuations in temperature, oxygen, and food availability) modulate ENM toxicity. Research is just beginning to broach this issue in fish

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and crustaceans (e.g. Ferreira et al., 2016; Wong and Leung, 2014) and much more data is

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needed to complete the picture.

A growing body of literature is addressing ENM toxicity in marine invertebrates (Canesi

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and Corsi, 2015) but data are still severely lacking for brackish and marine fish. Increases in salinity compress the electrical double layer of ENMs (Handy et al., 2008; Hotze et al., 2010),

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reducing the repulsive forces contributing to colloidal stability and promoting ENM aggregation and precipitation. The bioavailability of ENMs is therefore likely to be limited in saline waters, so work thus far has justifiably focused on freshwater systems. However, studies have noted

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ENM toxicity to brackish and marine fish (e.g. Wang et al. 2015a) and numerous marine

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invertebrates (reviewed by Baker et al., 2014; Canesi and Corsi, 2015). Given the noteworthy input of ENMs to estuarine and coastal marine systems, more data is necessary to understand the

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risks posed to animals in saline waters. This knowledge gap was clearly highlighted by a recent analysis of the available literature on nanotoxicology in saltwater (Minetto et al., 2016). Numerous authors have emphasized the need for thorough ENM and water quality characterization in nanotoxicological studies (Boyle et al., 2015; Coleman et al., 2013; Furtado et al., 2014; Klaine et al., 2008; Leung et al., 2015; Mueller and Nowack, 2008; Mwangi et al., 2012; O’Brien and Cummins, 2010; Warheit et al., 2007) and its importance should be clear from the examples presented within. Many studies, including our own, use ENMs curated from high-volume industrial suppliers to better represent environmentally-relevant conditions but manufacturers may not wish to disclose some proprietary information related to ENM properties. While some properties such as core size/shape, surface (zeta) potential, and hydrodynamic diameter are relatively straight-forward to determine, identifying the exact structure of an unknown surface functional group can be more complex. Such information is key in interpreting toxicity data, as small variations in a surface functional group can significantly influence colloidal behavior and bioactivity (e.g. MacCormack et al., 2012). Extending ENM 22

ACCEPTED MANUSCRIPT quantification and characterization techniques to complex water and tissue samples from the environment is a considerable challenge, and is one that must be overcome to move the discipline forward (von der Kammer et al., 2012). A major goal in nanotoxicological research is

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to identify general ENM characteristics associated with toxicity and comprehensive characterization data is necessary to achieve this. A database facilitating the meta-analysis of

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existing information in this area could be quite valuable for identifying bioactive ENM

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characteristics and informing regulatory decisions on novel ENM formulations. At the laboratory bench, consideration should also be given to the compatibility of ENMs with analytical

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instrumentation and common biochemical assays. The unique physical, chemical, and electronic properties of ENMs may interfere with assay mechanisms and lead to erroneous interpretations

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of toxicity (Ong et al., 2014a). The introduction of standards for ENM preparation and quantitative toxicity testing may address many of the remaining uncertainties surrounding ENM

6 Concluding Remarks

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risk assessments (Warheit et al., 2007).

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Overall, the existing data indicate that exposure to most common ENM formulations at environmentally-relevant concentrations is unlikely to induce acute mortality in aquatic animals under natural conditions. It is probable, however, that chronic exposures under specific environmental conditions will negatively impact the physiology of some species, decreasing their capacity to cope with normal stressors (Figure 2). Comparisons between nanotoxicological studies are often complicated by inconsistent ENM characterization, emphasizing the need for standardized and comprehensive investigations into ENM behavioural parameters. Furthermore, the marked differences in ENM behaviour in laboratory and environmentally-relevant conditions, along with environmental release estimates, suggest that future nanotoxicological studies must be undertaken with due consideration to the conditions under which organisms are expected to encounter the ENM of interest (Figure 1). As a consequence of our mandate, we focused considerable attention on a number of negative outcomes associated with ENM exposure, but rapid advances in nanotechnology also hold great promise for improving environmental stewardship. For example; the use of ENMs in

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ACCEPTED MANUSCRIPT various aquaculture applications may greatly increase productivity and decrease environmental impacts (Handy et al., 2011) and the development of nano-enabled catalysts and filtration devices will improve WWTPs and environmental remediation processes. Adopting a balanced

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approach to the regulation and proliferation of the nanotechnology industry could substantially reduce the overall burden of contaminants in the aquatic environment and promote recovery in

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highly impacted systems.

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Acknowledgements

The authors wish to thank two anonymous reviewers for insightful and constructive

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feedback. NIC was supported by a Natural Sciences and Engineering Research Council of Canada (NSERC) Canada Graduate Scholarship-Master’s, an Ontario Graduate Scholarship, a New Brunswick Innovation Foundation scholarship, and a C. David Naylor Fellowship Endowed

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by a Gift from the Arthur L. Irving Foundation. TJM is supported by an NSERC Discovery grant

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Xia, T., Kovochich, M., Liong, M., Madler, L., Gilbert, B., Shi, H., Nel, A., 2008. Comparison

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of the mechanism of toxicity of zinc oxide and cerium oxide nanoparticles based on dissolution and oxidative stress properties. ACS Nano 2, 2121–34.

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Xin, Q., Rotchell, J.M., Cheng, J., Yi, J., Zhang, Q., 2015. Silver nanoparticles affect the neural development of zebrafish embryos. J. Appl. Toxicol. 35, 1481–1492. doi:10.1002/jat.3164 Zhang, C., Hu, Z., Deng, B., 2016. Silver nanoparticles in aquatic environments: Physiochemical behavior and antimicrobial mechanisms. Water Res. 88, 403–427. doi:http://dx.doi.org/10.1016/j.watres.2015.10.025 Zhang, Y.-Q., Dringen, R., Petters, C., Rastedt, W., Köser, J., Filser, J., Stolte, S., 2016. Toxicity of dimercaptosuccinate-coated and un-functionalized magnetic iron oxide nanoparticles towards aquatic organisms. Environ. Sci. Nano. doi:10.1039/C5EN00222B Zhu, X., Tian, S., Cai, Z., 2012. Toxicity Assessment of Iron Oxide Nanoparticles in Zebrafish (Danio rerio) Early Life Stages. PLoS One 7, 1–6. doi:10.1371/journal.pone.0046286 Zhu, X., Wang, J., Zhang, X., Chang, Y., Chen, Y., 2009. The impact of ZnO nanoparticle aggregates on the embryonic development of zebrafish (Danio rerio). Nanotechnology 20, 195103. doi:10.1088/0957-4484/20/19/195103 37

ACCEPTED MANUSCRIPT Zhu, X., Zhu, L., Duan, Z., Qi, R., Li, Y., Lang, Y., 2008. Comparative toxicity of several metal oxide nanoparticle aqueous suspensions to Zebrafish (Danio rerio) early developmental stage. J. Environ. Sci. Heal. Part A, Toxic/hazardous Subst. Environ. Eng. 43, 278–284.

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ACCEPTED MANUSCRIPT Table 1. Estimates of production, aquatic release, and environmental levels in a model system (San Francisco Bay) of commercially relevant ENMs. Data on colloidal behaviour and stability in relevant media is also provided, where available (references in brackets). Production (kt/y)

TiO2

88 [2]

Aquatic release (kt/y) 15.6 [2]

Approximate environmental levels (µg L-1) [1] 4-50

Behaviour in freshwater

≤20 nm uncoated in MilliQ produced aggregates of 30 nm [6]

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1-10 [3]

63 nm, unspecified coating in MilliQ produced aggregates of 300 nm [7]

10.5-11.6 nm uncoated in zebrafish embryo medium produced aggregates of 2110-2550 nm [10]

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5 [4]

Behaviour in brackish/seawater or saline

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ENM

3 [5]

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21 nm uncoated (2 mg L-1) produced aggregates of 323 nm at 0 h and 797 nm after 24 h, while c.a. 24% had precipitated. At 7 mg L-1, initial aggregates of 827 nm increased to 1538 nm and 67% precipitated over 24 h [8] 10-30 nm uncoated (10 mg L-1) in deionized water produced aggregates of 432 nm [9]

95 [2]

2.1 [2]

0.1-11

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SiO2

100 [4] 42 [2]

4.3 [2]

0.7-10

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Fe/FexOy

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10-100 [3]

0.01-0.1 [3]

ZnO

34 [2] 0.1-1 [3]

3.7 [2]

0.8-10

0.25 [1112]*

29 nm Fe2O3 of unspecified coating in MilliQ water produced aggregates of 1580 nm in MilliQ [7] 20-30 nm Fe3O4 of unspecified coating in MilliQ water produced aggregates of <200 nm in MilliQ [7] 20 nm uncoated in MilliQ produced aggregates of 180 nm [6] 71 nm of unspecified coating in MilliQ produced aggregates of 320 nm [7]

0.02 [4] 1.8 [5]

25 nm of unknown coating (10 µg L-1) in MilliQ produced aggregates of 332 nm after 30 min and dissolved by 0.002% after 24 h [13]

CeO2

10 [2] 0.1-1 [3]

0.3 [2]

0.02-1

13 nm uncoated (10 mg mL-1) produced aggregates of 60 nm after 3 h [14] 20 nm uncoated (10 mg L-1) produced 2.7 mg L-1 Zn2+ [15] 25 nm uncoated in MilliQ produced aggregates of 227 nm [18]

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20 nm uncoated in seawater produced aggregates of 300 nm and dissolved by 4.6% after 8 h [12] 13 nm uncoated (10 mg mL-1) in growth media produced 80-90 µmol L-1 Zn2+ after 3 h [16] 8.4 nm uncoated in zebrafish embryo medium produced aggregates of 1280 nm and precipitated by >80% in 3 h [17]

2.8-2.9 nm uncoated in zebrafish embryo medium produced aggregates of 2960 nm [10]

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0.06 [2]

0.5 [2]

0.01 [2]

<10 [3]

0.015 [22]

0.0001-0.05

42 nm CuO, unspecified coating in MilliQ produced aggregates 220 nm [7] 27 nm uncoated Cu in relatively hard water dissolved by 0.03% [19] 27 nm uncoated Ag in relatively hard water dissolved by 0.07% [19]

25, 50, and 100 nm uncoated Cu in egg water dissolved by 10-42%, 623%, and 8-24%, respectively [20]†

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Ag

10 [4] 0.2 [2]

0.01-0.2

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Cu

Stable between 0.21-3.6 mg L-1 at 12 nm (non-aggregating) for >120 h in up to 10 mM NaCl [23]

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0.005 [5] 0.1-0.3-0.8 [21]

3 nm, citrate coated at 5 mg L-1 increased in mean hydrodynamic diameter from 40 nm in freshwater to c.a. 8 µm in 10% seawater, with a corresponding increase in Ag+ release from <0.2% of total mass in freshwater to c.a. 4% in 10% seawater after 48 h [27]

Stable up to 48 h at 100 mg L-1 35 nm in 0.1% Na-citrate [24]

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0.2 [22]

3 [2]

0.03 [2]

0.003-0.3

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24, 30, and 68 nm, PVP coated in dechlorinated tap water produced aggregates of 88, 43, and 74 nm respectively. 46 and 67 nm uncoated in dechlorinated tap water produced aggregates of 69 and 143 nm, respectively [25]

0.01-1 [3] 0.5 [21]

30 nm PVP coated in dechlorinated tap water produced aggregates of 50-63 nm and dissolved by 6.4% after 24 h [26] 110-170 nm x 5-9 µm SWCNT, unspecified coating in MilliQ produced aggregates of 300 nm x 6 µm [7]

100 mg L-1 HO- and COOHfunctionalized MWCNTs (20-30 nm x 10-30 µm) could not be stabilized in MilliQ or seawater while unfunctionalized MWCNTs (10-30 nm x 10-30 µm) could [28]

* release estimate for sunscreen (cosmetic) use only † range of dissolution depended on concentration (1-8 mg L-1); higher concentrations had uniformly lower % dissolution by mass. Dissolution measurements in egg water (0.21 g Instant Ocean in 1 L) 1 (Keller and Lazareva, 2014), 2 (Keller et al., 2013), 3 (Piccinno et al., 2012), 4 (United Nations Environment Programme UNEP, 2007), 5 (Gottschalk et al., 2009), 6 (Zhu et al., 2008), 7 (Karlsson et al., 2008), 8 (Ma et al., 2012), 9 (Angelica and Fong, 2013), 10 (Wehmas et al., 2015), 11 (Danovaro et al., 2008), 12 (Wong et al., 2010), 13 (Bessemer et al., 2015), 14 (Xia et al., 2008), 15 (Zhu et al., 2008), 16 (Xia et al., 2008), 17 (Zhu et al., 2009), 18 (Rundle et al., 2016), 19 (Griffitt et al., 2008), 20 (Hua et al., 2014), 21 (Mueller and Nowack, 2008), 22 (Blaser et al., 2008), 23 (Lee et al., 2009), 24 (Griffitt et al., 2012), 25 (Coleman et al., 2013), 26 (Wu and Zhou, 2013), 27 (Auffan et al., 2014), 28 (Kennedy et al., 2008)

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Figure 1. Aquatic release and fate of engineered nanomaterials (ENMs). ENMs are introduced to the environment in multiple forms (relative abundance indicated by arrow thickness). Once in the aquatic system, colloidal behaviour and stability depend on physicochemical interactions between the ENM and the surrounding environment; ENMs may dissolve into their component ions, precipitate, or be stabilized in the water column by complexing with humic matter or sequestration by algae. Precipitation will produce higher localized ENM concentrations in the substrate or immediate area as a result of agitation (e.g. by benthic organisms). In the upper region of the water, ENMs can be dispersed by aerosolization in the surf and may induce additional toxicity when distributing UV energy that would otherwise be absorbed by the water.

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Figure 2. Potential mechanisms of engineered nanomaterial (ENM) toxicity in aquatic animals. Exposure to ENMs can trigger the generation of reactive oxygen species (ROS) either directly or as a result of immune system responses. The resulting oxidative stress may cause tissue damage and/or epithelial remodeling, particularly in the gill and gut. Direct and secondary interactions between ENMs (or their components) and proteins may elicit immune responses that can contribute to localized or systemic inflammation. Such interactions may also inhibit hatching and lead to developmental defects in embryonic animals. These cellular and organ-level responses can eventually impact whole animal performance and survival.

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