RESEARCH FRONT
CSIRO PUBLISHING
Environ. Chem. 2014, 11, 207–226
http://dx.doi.org/10.1071/EN13221
Review
Aquatic toxicity of manufactured nanomaterials:
challenges and recommendations for future
toxicity testing
Aaron G. Schultz,A,E David Boyle,A Danuta Chamot,A Kimberly J. Ong,A
Kevin J. Wilkinson,B James C. McGeer,C Geoff SunaharaD and Greg G. GossA
A
Department of Biological Sciences, University of Alberta, Edmonton, AB, T6E4W1, Canada.
Department of Chemistry, University of Montreal, PO Box 6128, Succursale Centre-Ville
Montreal, QC, H3C 3J7, Canada.
C
Biology Department, Wilfrid Laurier University, Waterloo, ON, N2L 3C5, Canada.
D
Aquatic and Crop Resource Development, National Research Council of Canada,
6100 Royalmount Avenue, Montreal, QC, H4P 2R2, Canada.
E
Corresponding author. Email: ags2@ualberta.ca
B
Environmental context. The increased use of nanomaterials in industrial and consumer products requires
robust strategies to identify risks when they are released into the environment. Aquatic toxicologists are
beginning to possess a clearer understanding of the chemical and physical properties of nanomaterials in
solution, and which of the properties potentially affect the health of aquatic organisms. This review highlights
the main challenges encountered in aquatic nanotoxicity testing, provides recommendations for overcoming
these challenges, and discusses recent studies that have advanced our understanding of the toxicity of three
important OECD nanomaterials, titanium dioxide, zinc oxide and silver nanomaterials.
Abstract. Aquatic nanotoxicologists and ecotoxicologists have begun to identify the unique properties of the
nanomaterials (NMs) that potentially affect the health of wildlife. In this review the scientific aims are to discuss the
main challenges nanotoxicologists currently face in aquatic toxicity testing, including the transformations of NMs in
aquatic test media (dissolution, aggregation and small molecule interactions), and modes of NM interference (optical
interference, adsorption to assay components and generation of reactive oxygen species) on common toxicity assays.
Three of the major OECD (Organisation for Economic Co-operation and Development) priority materials, titanium
dioxide (TiO2), zinc oxide (ZnO) and silver (Ag) NMs, studied recently by the Natural Sciences and Engineering Research
Council of Canada (NSERC), National Research Council of Canada (NRC) and the Business Development Bank of
Canada (BDC) Nanotechnology Initiative (NNBNI), a Canadian consortium, have been identified to cause both bulk
effect, dissolution-based (i.e. free metal), or NM-specific toxicity in aquatic organisms. TiO2 NMs are most toxic to algae,
with toxicity being NM size-dependent and principally associated with binding of the materials to the organism.
Conversely, dissolution of Zn and Ag NMs and the subsequent release of their ionic metal counterparts appear to represent
the primary mode of toxicity to aquatic organisms for these NMs. In recent years, our understanding of the toxicological
properties of these specific OECD relevant materials has increased significantly. Specifically, researchers have begun to
alter their experimental design to identify the different behaviour of these materials as colloids and, by introducing
appropriate controls and NM characterisation, aquatic nanotoxicologists are now beginning to possess a clearer
understanding of the chemical and physical properties of these materials in solution, and how these materials may
interact with organisms. Arming nanotoxicologists with this understanding, combined with knowledge of the physics,
chemistry and biology of these materials is essential for maintaining the accuracy of all future toxicological assessments.
Additional keywords: nanoparticles, nanotoxicology, silver, titanium dioxide, zinc oxide.
Received 9 December 2013, accepted 22 May 2014, published online 20 June 2014
Introduction
as small materials that are at most 1–100 nm in one dimension
and have unique properties and functions resulting from their
small size.[10,11] NMs have recently been redefined by the
European Commission as ‘a natural, incidental or manufactured
material containing particles, in an unbound state or as an
aggregate or as an agglomerate and where, for 50 % or more of
the particles in the number size distribution, one or more external
The increasingly rapid production and release of a novel type of
synthetic material collectively termed nanomaterials (NMs),
necessitates their investigation as possible threats to ecosystems
(see reviews by Baun et al.,[1] Handy et al.,[2] Klaine et al.,[3]
Moore,[4] Scown et al.,[5] Peralta-Videa et al.,[6] Klaine et al.,[7]
Bondarenko et al.[8] and Ray et al.[9]). These materials are defined
Journal compilation Ó CSIRO 2014
207
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A. G. Schultz et al.
dimensions is in the size range 1–100 nm’.[12] Nanotechnology
and the production and utilisation of NMs is a multibilliondollar per annum commercial industry that is expanding
rapidly. Increased production efficiencies and advances in the
development of NMs have enabled an increased number of
nano-enabled medical, industrial and consumer products.[13]
Currently, the most widely used NMs in consumer products
are engineered metal-containing and metal oxide (MeO) NMs,
Aaron Schultz is a Post-Doctoral Fellow in the laboratory of Prof. Greg Goss at the University of Alberta, Canada. He is a fish
physiologist and aquatic toxicologist and received his B.Sc. (with 1st Class Honours) in 2006 and Ph.D. from Deakin
University, Geelong, Australia, in 2010. His primary research interests focus on studying the adaptive mechanisms employed
by fish and other aquatic organisms that allow them to survive in constantly changing environments and in response to
anthropogenic contaminants such as nanomaterials.
David Boyle is a Post-Doctoral Fellow in the research group of Prof. Greg Goss at the University of Alberta, Canada. He
received his B.Sc. from the University of Bath, UK, in Applied Biology in 2001 and his M.Sc. in 2003 and Ph.D. in 2008 both
from King’s College London, UK. After an appointment as a Research Scientist at the National Institute for Nutrition and
Seafood Research, Bergen, Norway, he has worked as a Post-Doctoral Research Fellow in the field of aquatic nanotoxicology,
firstly at Plymouth University, UK, and since 2012, at the University of Alberta, Canada.
Danuta Chamot is a Senior Research Associate in the laboratory of Prof. Greg Goss in the Department of Biological Sciences
at the University of Alberta. She received her B.Sc. in 1988 and M.Sc. in 1990 from the University of Toronto and her Ph.D. in
plant molecular biology from the University of Bern (Switzerland) in 1993. She has spent most of her career studying stressregulated gene expression in aquatic microbes. Since joining the Goss lab, she has been closely involved in research on the
physiological and toxicological effects of pollutants on fishes.
Kimberly J. Ong is a graduate of the Biological Sciences Ph.D. program at the University of Alberta in Edmonton, Canada in
2013. She earned her B.Sc. in Marine and Freshwater Biology at the University of Guelph in 2006. Her most recent work
focussed on developing and validating appropriate biological testing techniques for nanotoxicity studies, and is also interested
in the effects of nanoparticles on fish development, behaviour and physiology.
Kevin J. Wilkinson is Professor at the Université de Montréal. His research is aimed at gaining a molecular level
understanding of contaminant bioavailability and mobility. Current projects are examining the nature of the physicochemical
processes influencing trace metal bioaccumulation by microorganisms and determinations of the fate (dissolution, aggregation) and bioavailability (bioaccumulation, genomic and proteomic effects) of engineered nanomaterials. Wilkinson is
currently an Associate Editor of Environmental Chemistry (2010–) and in the past has edited two volumes of Biophysicochemistry of Environmental Systems. He has over 100 publications, over 4000 citations to his work and an h-index of 40. He
has recently established a world-class analytical laboratory specialising in the characterisation of nanomaterials: Center for
the Analysis and Characterization of Engineered Nanomaterials (CACEN).
Jim McGeer is an Associate Professor in the Biology Department and Director of the Institute for Water Science at Wilfrid
Laurier University. He completed B.Sc. Agr. and M.Sc. degrees at the University of British Columbia and then a Ph.D. at the
University of Dundee in 1995. He joined Laurier in 2006 following postdoctoral studies at Ben-Gurion and McMaster
Universities and then a period as a federal government scientist and research manager at Mining and Mineral Sciences
division of Natural Resources Canada. The McGeer lab is focussed on solutions based research directed at integrating
fundamental understandings of how inorganic contaminants impinge on physiological processes in aquatic organisms and the
application of this understanding in prediction models that contribute to environmental protection by reducing uncertainty.
Geoffrey Sunahara is a Senior Research Officer at the National Resource Council of Canada (NRC). He received his B.Sc. in
1976 and M.Sc. in 1979, both from the University of Toronto, and his Ph.D. (Pharmaceutical Sciences) in 1984 from the
University of British Columbia. He completed a 2-year Fogarty post-doctoral fellowship at the National Institute of
Environmental Health Sciences (NIH). After working at the Nestec Research Centre (Switzerland) until 1994, he then joined
NRC to develop the Applied Ecotoxicology group that focuses on the ecotoxicology of emerging contaminants including
nanomaterials, biodiesel and other bioproducts. A major emphasis is made upon innovation, ecological relevance, risk
assessment and modes of toxicity.
Prof. Greg Goss is appointed in the Department of Biological Science, Faculty of Science at the University of Alberta with a
cross-appointment to the School of Public Health at the University of Alberta. He is a fellow of National Institute of
Nanotechnology, the Scientific Director of University of Alberta Water Initiative and Director of the Office of Environmental
Nanosafety at the University of Alberta. Dr Goss works jointly with industry, governments and academia to examine the
environmental toxicology of micropollutants including nanomaterials, pharmaceuticals and personal care products, hydraulic
fracturing fluid and hydrocarbon contaminated fluids.
208
Aquatic toxicity of manufactured nanomaterials
including nAg, nTiO2 and nZnO. In 2013, the Woodrow Wilson
database listed 1628 products containing NMs, 383 of which
contained Ag NMs, 179 contained TiO2 and 36 products contained ZnO NMs (Project on Emerging Nanotechnologies, see
http://www.nanotechproject.org/cpi/about/analysis/, accessed 25
October 2013). The growing interest and use of these metal-based
and metal oxide NMs for future applications will likely result in
greater release of these NMs and their resulting ionic counterparts
into aquatic environments. Indeed, a growing number of studies
have already confirmed the release of Ag NMs into wastewater
from washing machines,[14] clothing,[15–17] toothpaste, shampoo
and detergent,[15] whereas TiO2 NMs have been reported to be
released from textiles[18] and sunscreens.[19,20]
In response to these concerns, many national agencies
worldwide have established research consortia, including but
not limited to Centres for Environmental Implications of
Nanotechnology (CEINT, USA), NanoFATE (UK), NanoSafe
(Australia), Managing Risks of Nanomaterials (MARINA,
EU), NanoValid (EU) and Natural Sciences and Engineering
Research Council of Canada (NSERC), National Research
Council of Canada (NRC) and the Business Development Bank
of Canada (BDC) Nanotechnology Initiative (Environment
Canada) (NNBNI, Canada). Under the Organisation for Economic Co-operation and Development (OECD), a list of
priority NMs was defined and consortia were tasked with
providing environmental characterisation and toxicological
data of these materials for use in risk analysis.[21] This review
provides an update of the toxicological properties of key
OECD materials with a focus on the findings of the NNBNI
consortium (Canada). Other metal-containing and organic
NMs, including but not limited to, gold NMs,[22,23] carbon
nanotubes[22,24,25] and silicon-based NMs[22,26] also merit
toxicological analysis, however they are beyond the scope
of this review. Here we highlight some of the key advances
in our understanding of aquatic nanotoxicological testing
methods including NM transformations (dissolution, agglomeration and interaction with small molecules) in aqueous
test media and NM interference with common toxicological
assays. The characterisation requirements and controls that
will allow researchers to assess both NM transformations
and assay interference, and accurately interpret toxicological
results are also discussed. Finally, we discuss the advances
in our understanding of the toxicology of three key OECD
materials: TiO2, Ag and ZnO NMs and conclude with a section
on primary knowledge-gaps and key recommendations for
future toxicity studies of NM.
organisms therein. These NM transformations and interactions
are discussed in detail below, as well as recommendations for
advanced characterisation and proper controls to incorporate so
accurate interpretation of toxicity experiments can be made.
NM transformations in aqueous test media
Numerous abiotic factors influence the physicochemical state of
NMs, both spatially and over time. Any variation in exposure
conditions will create difficulties in the interpretation of toxicity
experiments. In contrast to soluble, well dispersed toxicants,
nanotoxicologists must understand both the novel properties of
the material and the colloidal nature of the NMs. Fig. 1 identifies
the key factors that should be considered when designing toxicity studies. For example, it is necessary to quantify factors
such as dissolution (Fig. 1a), agglomeration and sedimentation
(Fig. 1b) or the adsorption of ions and macromolecules (Fig. 1c)
in order to understand bioavailability and toxicity. Improved
knowledge of the catalytic activity of the NMs (generation of
reactive oxygen species (ROS), bandgaps for semiconductors
and quantum confinement (Fig. 1d)) will also allow the
researcher to better understand the potential effect this may have
on nanotoxicity, which will facilitate our ability to properly
assess the risk of the NMs.
Dissolution of NMs
It is well known that aquatic organisms are particularly
sensitive to free metals,[27] and exposure at sub-lethal concentrations can affect several processes, including development,[28]
movement,[29,30] hatch rate[31–33] and ion balance.[34] A major
concern associated with metal-based NMs is the dissolution and
release of ionic species in aquatic matrices. For example,
nominal concentration[35] and diameter will affect dissolution[36–38] whereby the relatively larger surface area to volume
ratio in smaller particles results in more exposed atoms and
higher dissolution rates.[36,37] The mechanism of dissolution of
NMs is different than that for bulk materials, and not always
fully predictable based on theoretical calculations.[36,37,39]
Stable surface coatings, binding of molecules from the test
media or exudates from the organism, or aggregation of NMs
can block exposed NM surfaces, potentially reducing dissolution rates.[40] In contrast, elevated temperatures[41] and low
pH[36,38,42] tend to increase the rate of dissolution of metals
and metal oxides. Media composition will also affect dissolution
whereby constituents may complex dissolved metals, compete
for binding sites on an organism[42] or form a surface coating on
the NMs themselves, thereby abrogating toxic effects.[43]
Although most studies are performed at circumneutral pH, the
potential exists for colloidal NMs that are accumulated by a cell
to encounter acidic compartments such as lysosomes, which
may increase dissolution inside the cell.[37,44,45] This so called
‘Trojan horse’ effect will be discussed later in this review
(see section Summary of progress on OECD priority nanomaterials: TiO2, ZnO and Ag).
Monitoring free metal release from a NM is essential to
distinguish metal ion effects from NM-specific effects. Carefully designed studies have shown that NMs can have effects
distinct from their ionic forms, resulting in damage by different
mechanisms than those reported for their dissolution (free ion)
constituents.[37,46–49] Therefore, great care must be taken to
determine not simply NM dissolution throughout an experiment
but also trace metal speciation where relevant. Because some
analytical techniques measure total dissolution (e.g. dialysis,
NM testing requires advanced characterisation
for proper interpretation
Accurate testing of NM toxicity to aquatic organisms is necessary to determine whether specific regulations should be mandated for NMs, however, conventional in vitro and in vivo tests
employed for traditional toxicants may not be appropriate for
many NMs. We discuss two major points where the absence of a
thorough characterisation will result in an incomplete or
incorrect interpretation of toxicological or environmental
results. First, physicochemical transformations of NMs readily
occur in aquatic and test media leading to uncertainty during
toxicity testing, which then affects our confidence in providing
accurate risk assessments. Second, the small size and unique
colloidal properties of NMs can lead to unpredictable interactions with the test media and its constituents, including the
209
A. G. Schultz et al.
(a)
(b)
NM
NM
NM M
NM
NM
NM
Me⫹
NM
NOM NM
NM
NM
temp
Me⫹
NM
pH
Me⫹
NM size
(c)
NM
NOM
Me⫹
(d)
n
n
NM
s
Protein
n
NM
n
NM
n
s
NM
s
O2
UV
O⫺
2
s
NM
NM
s
O2
NM
O2
O⫺
2
O⫺
2
Fig. 1. Potential nanomaterial (NM) transformations in aquatic test media. (a) Dissolution of NMs
releasing ionic metals into the test media, with dissolution rates increasing at higher temperatures, lower
pH’s, and as NM size decreases. (b) Agglomeration and sedimentation of NMs. The NMs may bind to
each other or other particles (e.g. natural organic matter, proteins) in test media. (c) NMs have the ability
to bind (adsorb) to small molecules, such as salts (s) and nutrients (n), and macromolecules, such as
proteins, in test media. (d) Catalytic activity of NMs in aquatic test media can generate reactive oxygen
species (e.g. O2 , H2O2 ).
agglomerate.[36,55,56] Some electrolytes also affect the stability
of NMs; in particular, high concentrations of divalent cations,
such as Ca2þ or Mg2þ can cause increased agglomeration as a
result of particle bridging, charge screening or surface
complexation.[54,56–58]
Sterically stabilised NMs, such as those with capping agents,
may not be as susceptible to changes in abiotic conditions as
electrostatically stabilised NMs.[30,54] In aquatic matrices, NMs
are known to adsorb molecules and form a ‘corona’,[59–62] a
dynamic outer layer that coats the NM and can change agglomeration kinetics among other properties. Spontaneous adherence, or replacement of the coating by natural organic matter
(NOM) in natural aquatic environments can produce charge and
steric stabilisation of NMs.[56–58] Similarly, within a whole
organism, intracellular and serum proteins can contribute to
this coating around the NM.[30,63] The extent of surface coverage
on the NM can change the agglomeration state. Whereas
complete coverage may result in steric or electrostatic stabilisation, only partial coverage may cause increased agglomeration
due to destabilisation by charge neutralisation, bridging or the
interaction of oppositely charged patches of the NM.[30,36,53,64]
The type and combination of proteins or organic molecules,[30]
their concentration[58,65] and the incubation period[30,57] are all
known to affect these interactions and must be taken into
consideration when performing nanotoxicology experiments.
Monitoring the size of NM agglomerates, their sedimentation
and the population of remaining monodispersed NMs throughout an experiment is essential when predicting their distribution
and availability to organisms in aquatic environments.[30] A few
studies have suggested that only NMs of a specific size will be
taken up by cells or organisms,[66,67] whereas aggregates above a
threshold size may be prevented from translocating across
protective membranes, such as the embryonic fish chorion.[68]
ultracentrifugation, ultrafiltration) whereas others measure
either free ions (e.g. ion selective electrode) or ionic species
(e.g. diffusion in thin film gradient technique, voltammetry),
data interpretation must carefully take into account methodological considerations. Furthermore, because dissolution rates
can change over time as NMs either aggregate, bind to other
molecules or surfaces or translocate across membranes, it is
highly recommended to follow dissolution under the precise
experimental conditions (e.g. medium, concentration, time,
light) being used. It is no longer acceptable to solely measure
dissolution of NMs in media or solutions other than those being
used in the experiment.
Agglomeration of NMs
Agglomeration of NMs is dependent on both biotic and
abiotic factors.[30] Agglomeration is known to reduce the
available active surface of the NM, decrease the stability in
the water column (i.e. increased sedimentation), and affect the
aggregates’ ability to translocate over membranes or surfaces.
Together, these processes act to reduce exposure and therefore
hazard in the suspended test organisms, but may increase
exposure and therefore hazard in benthic test organisms.[50]
According to classic Derjagiun, Landau, Verwey and Overbeek (DLVO) theory, agglomeration occurs when colloidal
particles’ attractive forces (e.g. van der Waals) overwhelm
repulsive forces (e.g. electrostatic).[51–53] Changes in a solution’s properties can affect the electrostatic double layer (EDL),
and particle electric charge. Compression of the EDL in solutions of high ionic strength can reduce the repulsive forces of a
NM (observed through a reduction in zeta potential), resulting in
increased agglomeration.[36,54–56] Furthermore, if the pH of a
solution is close to that of the point of zero charge of the NM,
repulsive electrostatic forces are weakened and NMs will likely
210
Aquatic toxicity of manufactured nanomaterials
Furthermore, large agglomerates that form will likely settle,
effectively concentrating the NM at the bottom of the experimental chamber. This may be particularly important in cell
culture studies[69,70] and experiments where the test organism is
located at the bottom of the test vessel (e.g. zebrafish embryos or
benthic organisms[71,72]).
(a) Optical interference
Intrinsic absorbance or fluorescence of NMs
Assays affected:
• MTT, LDH and DCF assays
Absorbance
1.0
Small molecule interactions
Upon addition to aquatic media, NMs adsorb to small
particles in the water column, which may be critical components
for an organism’s survival. Although nanotoxicologists tend to
ascribe the effects of NM exposure as a direct toxic effect, these
additional interactions may simply affect organismal survival by
sequestering nutrients and essential minerals, such as, hormones, growth factors, salts, proteins, nutrients and vitamins.[73]
Depletion of any of these factors by their adsorption to NMs may
reduce their effective concentration and indirectly limit growth
and survival.[74,75] For example, Guo et al.[76] found that amino
acids and vitamins are adsorbed to single-walled carbon nanotubes (SWCNTs), depleting essential nutrients that may result in
decreased cell health. Ions such as calcium or phosphate have
been shown to be sequestered by NMs,[75,77] potentially leading
to homeostatic imbalances or deficiencies. Similarly, the inhibition of enzyme activity necessary for fish hatch has been
shown to result in lethality in embryo tests.[47] It has been
speculated that binding of NMs to the gills and skin[78] of fishes
may lead to interactions with proteases and lysozymes that
protect fish from bacterial infection.[79]
Fluorescence
Au NM
Formazan
Absorbance
0.8
0.6
NM
0.4
0.2
NM
0
380 415 450 485 520 555 590 625 660 695
Wavelength (nm)
(b) Adsorption
NMs inhibit dye metabolisation or alter
enzymatic activity in assays
Assays affected:
• MTT, PI, LDH, alamar blue and neutral
red assays
Formazan
N
NH
N
N
N
S
l.
Metabo
Br⫺
N N
N⫹
N
S
Br⫺
MTT
⫹
NM
N
No
Metabol.
MTT
N N
N⫹
N
NM
N
(c) Catalytic properties
NMs produce ROS resulting in false positives
Assays affected:
• MTT assay and DCFH-DA dye
Nanomaterial interference with toxicological assays
High-throughput in vitro assays are commonly used to test the
toxicity of substances on aquatic organisms. In recent years,
several reports have emerged indicating that a variety of NMs
can interfere with standard toxicity assays, including cellular
viability tests, such as alamar blue[80–82] and neutral red[80,81,83]
lactate dehydrogenase (LDH) release,[82–86] and tetrazolium dye
based assays (water soluble tetrazolium, WST[80,87]; 2,3-bis-(2methoxy-4-nitro-5-sulfophenyl)-2H-tetrazolium-5-carboxanilide,
XTT[88]; 3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetrazolium
bromide, MTT[83,89–92]). Although nanoparticle-assay interferences are becoming more regularly reported, the majority of
published nanotoxicity papers still do not test for interference,[82] resulting in difficulties interpreting data. The interactions that occur between the NM and components of an
assay are still not well understood, making it difficult to predict
the consistency of each test. To date, the primary modes of
NM-assay interference identified are: (1) direct optical interference by the NM, (2) adsorption of the NM to important assay
components and (3) the generation of ROS by the NM (see
Fig. 2). Each of these modes is discussed below with subsequent
considerations for selecting assays of NM toxicity.
DCFH-DA
O
O
CI
DCF
O
CI
O
O
HO
CI
HO
O
O
O
ROS
CI
O
NM
OH
Fig. 2. Model illustrating the three main types of nanomaterial (NM)
interference on standard toxicological assays used to assess the effects of
NMs on aquatic organisms. (a) Optical properties of NMs including intrinsic
absorbance or fluorescence can interfere with assays. UV-vis spectral trace
displays the overlapping absorbance of gold (Au) NMs and formazan (the
dye in the 3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide
(MTT) assay used to measure cell proliferation). Assays affected by optical
interference include MTT, lactate dehydrogenase (LDH), and 20 ,70 -dichlorodihydrofluorescin (DCF) assays.[99] (b) NMs can adsorb to dye’s, such as
MTT,[89] and enzymes in toxicological assays preventing metabolisation
(Metabol.) of substrates and inactivation of enzymes. Assays affected
include, MTT,[89] PI (propidium iodide),[104] LDH,[108] alamar blue and
neutral red.[81,82] (c) Catalytic properties of NMs and their generation of
reactive oxygen species (ROS) in assay mixtures can generate false positives
by oxidising non-fluorescent probes to fluorescent probes, such as 20 ,70 dichlorodihydrofluorescin diacetate (DCFH-DA) to DCF.[100] MTT is
another assay that can be affected by ROS generation.[82]
Optical interference
Many plate-based assays are based on an end-point spectroscopic measurement of dyes or indicators. One main mode of
interference is that significant absorbance or fluorescence of the
NM occurs at the wavelength at which the assay is measured.
For example, Au NMs generally have a strong absorbance peak
between 520 and 580 nm.[93,94] They thus have the potential to
interfere with the MTT assay because the absorbance peak of the
Au NM directly overlaps with the absorbance of the formazan
salts in the assay (Fig. 2a). Moreover, the quantum properties
of semiconductor materials can result in their absorbance or
the production of interfering fluorescence spectra across UVvisible ranges.[95–99]
NMs may also interfere with the optical properties and output
of indicator molecules. When a NM binds to the indicator
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A. G. Schultz et al.
molecule, this may lead to either a reduction in resonance
frequency, and therefore intensity, or a shift in the absorbance
maxima for the dye. Such an effect will lead to an apparent
decrease in the absorbance or fluorescence measurement and
thus an underestimation or overestimation of biological effect,
depending on the assay characteristics.[87,99–103]
may be deemed acceptable; minimisation of nanoparticles in
the final product by centrifugation or filtration may limit effects.
Because many spectroscopic assays are affected by the presence
of NMs, the use of dye-free methods avoids the issues presented.
For example, flow cytometry, manual cell counting techniques,
electrical cell-based impedance protocols[109] or clonogenic
assays[110] can be used to assess viability.
A significant issue is our current inability to predict the
dominant mode of interference by each NM formulation on a
particular assay, therefore, each assay will be differentially
affected by the type and characteristics of the NM studied. To
be able to validate or create new standardised tests, highthroughput assay interference testing of NMs with similar cores
and different physico-chemical characteristics is needed to
enhance our understanding of, and ability to, predict interference. Adsorption capacity, dissolution, optical properties and
redox potential are particular NM characteristics that should be
explored. Given the propensity for NMs to influence assay
results, it is recommended that: (a) assays be conducted with
careful control over the final concentrations of NMs, (b) multiple assays be used to corroborate results[81,85] and (c) tests for
NM interference be conducted and reported for each assay and
NM used.
Adsorption to assay components
Owing to their small size and relatively high surface area to
volume ratio, NMs can adsorb to the components of the assay
and obstruct proper transformation of indicator molecules.[80,89]
For example, the MTT assay is contingent on the production and
subsequent dissolution of MTT-formazan crystals. However,
adsorption to SWCNTs renders the dye insoluble, thereby
affecting the final colourimetric measurement[89] (Fig. 2b).
Membrane impermeable dyes such as propidium iodide (PI)
are often used to identify non-viable cells with compromised
membranes (e.g. LIVE/DEAD Viability Assay Kit, Life Technologies). If NMs can bind to and traffic the dyes into viable
cells false positives may result,[104,105] and in other conditions,
NM binding to the dye may prevent the dye from translocating
and may make it appear that there is reduced toxicity.
It is well known that proteins and biomolecules can form a
‘corona’ (see reviews by Dawson et al.[61,106] and Monopoli
et al.[107]) around a NM and that these interactions can affect
the activity of enzymes.[108] Consequently, assays that measure
enzymatic activity are likely to be affected by NMs. For
example, the presence of active LDH in test media is often used
as an indicator of cell death following its release into the media.
As it has been shown that NMs can interact with LDH and
decrease enzyme activity,[108] likely through structural deformation of the enzyme, this could result in inaccurate results
regarding cell viability whereby toxicity would be masked.[84,85]
Summary of progress on OECD priority NMs:
TiO2, ZnO and Ag
There have been a substantial number of excellent, informative
and extensive literature reviews published in the past few
years on aquatic nanotoxicology.[8,111–114] Moreover, multiple
research consortia have contributed to the extensive and rapidly
growing body of knowledge in the field of nanotoxicology. In
order to reduce some of the complexity and overlap with these
recent reviews, we have focussed this review on the three major
NM groups that were extensively covered recently by the
NNBNI-Environment Canada, Canadian consortium. Relevant
published results contributing to our understanding of the toxicity of each NM type are highlighted.
For each of the three OECD priority materials, evidence for
both bulk effect, dissolution based (i.e. free metal) or a specific
toxicity attributable to the unique properties of the NM itself
have been identified. In the subsequent sections, we discuss each
OECD priority material and summarise recent research indicating distinct mechanisms of toxicity that might be attributable to
each material. Fig. 3 summarises each of the potential nanospecific, ionic or bulk mechanisms of toxicity relevant to metal
and metal oxide NMs. It is not the intention here to perform a
comprehensive review of the NM literature, but rather to draw
attention to studies of note that have illustrated well the properties of the NM that are important to toxicity and which should
inform future testing strategies.
NM generation of ROS
The physicochemical properties of some NMs enable them to
undergo oxidation and produce ROS that can potentially reduce
dyes to their fluorescent forms,[92,100] resulting in false positive
tests of toxicity. For example, incubation of carbon-based NMs
with 20 ,70 -dichlorodihydrofluorescin diacetate (DCFH-DA) (a
highly oxidisable dye that is a commonly used as a marker of
oxidative stress) results in the production of fluorescent 20 ,70 dichlorodihydrofluorescein (DCF) molecules even when cells
are not present (Fig. 2c).[100] Therefore, investigators should be
knowledgeable about the standard potential of dyes and oxidation state[100] of the NM. If the potential for NM-mediated
oxidation of the dye exists, alternative assays should be utilised
to prevent assay interference and the generation of false
positives.
Recommendations
Some assays appear to be more prone to error than others and
therefore, validating assays with each unique formulation of the
NM is essential. NM characteristics including method of synthesis, surface structure, core material, shape, etc. can modify
the degree of assay interference.[91,92,104,108] Varying the concentration of other biomolecules in the assay, such as bovine
serum albumin or foetal calf serum, may also influence NM
interference.[84,92,108] Preliminary tests assessing a range of NM
concentrations can indicate whether the assay is appropriate for
the NM used. In many cases, only high concentrations of NM
significantly affect an assay,[87,90] and lower concentrations
Titanium dioxide NMs
Titanium dioxide (TiO2) is a metal oxide that has been used
traditionally as a white pigment for paints, cosmetics, confectionaries, etc. Natural forms of TiO2 include the minerals
ilmenite (FeTiO2), rutile (TiO2) and sphene (CaSiTiO5), with
ilenite and rutile being the predominant forms (USEPA[115]).
Engineered TiO2 is generally produced in two-phase forms or as
a combination of these phases. The physico-chemical properties
of nanosized anatase and rutile (including hydrophobicity and
photoreactivity) are quite different than their micrometre-sized
counterparts.[115] These properties can be exploited in novel or
212
Aquatic toxicity of manufactured nanomaterials
(a)
(b)
4
(c)
2
TiO2
TiO2 TiO2TiO2
ZnO
TiO2
Apical
Zn
TiO2
14
Ag⫹
11
Ag
Embryo
hatch
Zn2⫹
C
TiO2
Ag
Ag⫹
Zn2⫹
Zn2⫹
ROS
⫹
Ag⫹
9
2⫹
3
Ag⫹
Zn2⫹
NP size
pH
6
TiO2
TiO2TiO2 TiO2
Ag
7
UV
Water
1
10
Cu(I)
Ag
ZHE1
5
UV
Zn2⫹
T
J
TiO2
Ag⫹
T
J
Ag
8
ZHE1
TiO2
Ag⫹
Zn2⫹
ROS
TiO2
Zn2⫹
Zn2⫹
Ag⫹
Zn2⫹
Ag⫹
Ag
Zn2⫹
ZHE1
TiO2
Nucleus
Nucleus
Nucleus
Hatch
Ag
13
12
Ag
Basolateral
Ag⫹
ATP
ECF
Ag
TiO2
Na⫹ K⫹
Fig. 3. Proposed model for (a) titanium dioxide nanomaterial (NM) (TiO2), (b) zinc oxide NM (ZnO) and (c) silver NM (Ag) uptake and possible
indirect and direct toxic effects within cells or the chorion of aquatic animals. (1) TiO2 NM aggregates binding directly to the cell membrane. (2) TiO2
may enter cells through endocytic processes and remain in the cells or pass through to the extracellular fluid (ECF; blood). (3) TiO2 NM binding and
disruption of the cell membrane. (4, 5) Irradiation of TiO2 NM in sunlight may produce reactive oxygen species (ROS) in experimental media or within
the cells. (6) The ‘Trojan horse effect’, where ZnO NM may be taken up by cells through endocytosis and release ionic Zn (Zn2þ) into the cell after
entering it. (7) Release of ionic Zn2þ from ZnO NM in experimental media, with the rate of dissolution affected by the size of the NM and pH of the
environment. Ionic Zn2þ released from ZnO NM may enter cells via ion channels (c). (8) After traversing the chorion, ionic Zn2þ released from the
ZnO NM may bind to zebrafish hatching enzyme (ZHE1) inactivating it and, therefore, inhibiting embryo hatch. (9) Silver NM (Ag) may enter cells
through endocytic processes and remain in the cells or pass through to the ECF. (10) Toxicity associated with the dissolution of ionic Agþ from Ag NM
in the test media. (11) The ‘Trojan horse effect’, where Ag NM may be taken up by cells through endocytosis and release ionic Agþ into the cell after
entering it, which results in toxicity. (12) Intact Ag NM that enter the cell may induce toxicity by binding to intracellular proteins, inactivating them,
such as Naþ, Kþ-ATPase. (13) Ionic Agþ released from the Ag NM can inhibit Naþ uptake by binding to and inactivating Naþ, Kþ-ATPase activity.
(14) Ionic Agþ released from the Ag NM in the experimental media may enter algal cells via ion transporters or channels, such as CuI transporters and
Na channels, and cause toxicity. TJ, tight junction.
improved consumer products or industrial processes including
UVA–UVB sunscreens, semiconductors (anatase bandgap
energy ¼ 3.2 eV or 385 nm), air deodorisers or wastewater
treatment technology.[116,117] These new products and processes
may pose a burden to the aquatic environment if TiO2 NMs are
released unintentionally. The actual environmental concentrations of TiO2 NMs are not known, although model calculations
offer some estimates.[64,118,119] Based on probabilistic material
flow analysis of product life cycle, the Nowack group calculated
the predicted environmental concentrations (PEC) for nanoTiO2 in Switzerland to be 0.021 mg L 1 in surface water, and an
annual increase of 1.203 mg kg 1 year 1 for sediment and
11.2 ng kg 1 year 1 for soil.[120] In another paper, they estimated 0.015 and 0.002 mg L 1 in surface water, an annual
increase of 358 and 53 mg kg 1 year 1 in sediment and an
annual increase of 1.28 and 0.53 mg kg 1 year 1 in soil for
Europe and the US.[121]
It is not known whether these concentrations have an
ecotoxicological effect in the environment. ‘Effects’ studies of
TiO2 NMs (and other NMs) have most often been conducted in
the laboratory using methods that were developed to test the
toxicity of soluble chemicals. Whether these methods are
appropriate for non-soluble chemicals and particulates such as
TiO2 NMs is controversial. For example, the use of mass-based
concentrations (e.g. milligrams per litre) is problematic when
comparing anatase and rutile because they possess the same
molar mass (79.87 g mol 1), but they may differ in size
(i.e. nanometric v. micrometric scale) or aspect ratio (spherical,
needles, rods).
There is a body of knowledge describing the toxicity of
nanosized and micrometre-sized TiO2 on aquatic organisms,
including algae, invertebrates and fish (see reviews of Baun
et al.,[1] Klaine et al.,[3] the United States Environmental
Protection Agency (US EPA),[116] Farre et al.,[122] Kent,[123]
Handy et al.,[2] Menard et al.[124] and Navarro et al.[125]).
Although there have been exceptions reported in the literature,[126] there is broad agreement that TiO2 NMs can cause
effects in aquatic organisms under certain test conditions;
however, the levels of toxicity reported (most obvious where
determined effect concentration (effective concentration, ECx)
values have been calculated) show considerable variability
between studies. For example, the average 72-h median effective concentrations (EC50) for Evonik Degussa P25 (mixedphase: 80 % anatase and 20 % rutile) nanoparticle on algal
growth calculated from three separate studies[127–129] was
28 22 (s.e.) mg L 1. This value is 2-fold lower than the 96-h
EC50 value of 50.1 mg L 1 for Evonik Degussa P25 on algal
growth calculated by Metzler et al.[130] This variability in EC50
values for P25 is likely related to differences in test sample
preparation, algae species and exposure durations, use of
213
A. G. Schultz et al.
artificial sunlight,[127] a non-standard toxicity test medium[129]
or a UV pre-illumination period.[128] Inconsistencies like these,
present challenges to risk assessors as well as national and
international regulators of TiO2 NMs, who seek to minimise
risk, and need to be considered for future toxicity experiments.
Another source of variability can be attributed to the reporting of the research, such as the physico-chemical characterisation of TiO2 NMs before and during the ecotoxicity assays,[2]
preparation of TiO2 NM test samples and experimental procedures used for the toxicity assays and the toxicity response endpoints, making comparisons between studies
problematic.[131,132] Based on the principles described above
(see NM transformation in aquatic test media), TiO2 NMs
(primary particles) can form concentration- and time-dependent
agglomerates in solution and these processes are strongly
dependent on the pH and ionic strength of the test
media.[56,127–130,133–135] In turn, these dynamic physicochemical processes may alter the toxic responses of organisms to
TiO2 NMs. Research issues and strategies to develop a comprehensive environmental assessment for TiO2 NMs have
been described elsewhere.[116] Also, the OECD has sponsored
an international program (Working Party on Manufactured
Nanomaterials, WPMN) to characterise the fate and ecotoxicity
of TiO2 NMs for risk assessment.[136] The issue of interexperimental variability should be better resolved after the
results on the P25 from the OECD WPNM are published.
Preliminary results have been described recently by HundRinke and Hennecke.[137]
Despite considerable variability in reported toxicological
responses of organisms to TiO2 NMs, further analyses of the
literature can be performed. TiO2 NMs comprise a small group
of NMs for which particle size controls (i.e. micrometre or bulk
sized TiO2) have often been included in toxicity assessments. In
order to identify trends and draw a consensus as to the potential
for nanoscale specific effects of TiO2, the following discussion
has selected some physico-chemical and toxicological data from
studies with similar experimental conditions and endpoints.
Within comparable studies, an additional weighting has been
given to the most recent literature.
Table 1 summarises the results of eight studies that compare
the toxicities of nano-sized TiO2 (i.e. primary particle diameter
,100 nm) and ‘bulk’ TiO2 (i.e. primary particle diameter
.100 nm) on freshwater algae, using different toxicity test
methods. The studies used TiO2 samples that were obtained
from various manufacturers or suppliers, and included anatase
or anatase–rutile combinations, of varying purity. Test samples
were prepared by homogenisation, stirring or ultrasonication,
and the TiO2 materials were characterised in the test media by
using a variety of analytical techniques. Only a limited number
of physico-chemical characteristics of the primary and secondary particles were consistently reported in the studies.
Algal toxicity assays were based on international (International Organization for Standardization or OECD) or national
(China) toxicity test guidelines, using different algal species.
Effects of the TiO2 NMs on algal growth were measured using
cell counts or chlorophyll fluorescence. EC50 values are
expressed as milligrams per litre (nominal concentrations). In
some studies, biochemical measurements of photosynthetic
efficiency, algal cell damage and oxidative stress[129,135,138,139]
were also provided. Detailed discussion of the effects on
algal biomarkers (MTT, LDH, DCFDA, etc.) is beyond the
scope of the present review. Based on the present study dataset
(Table 1), there was a clear difference between the average
specific surface area (SSA) of nanosized (195 m2 g 1; n ¼ 7) and
micrometre-sized TiO2 (8 m2 g 1; n ¼ 6) particles. In contrast,
EC50 values (mg L 1) for the nanosized particles (52 28 (SE);
n ¼ 8) were similar to those of the micrometre-sized samples
(55 19; n ¼ 6). Nonetheless, closer examination of the individual studies of the dataset provides strong and convincing
evidence that a size (SSA)–effects relationship with toxicity
may exist. Seven of the nine TiO2 NMs were more toxic than
their micrometre-sized counterpart (used as a reference), as
indicated by the elevated nano-toxicity ratio (NTR) values,
where EC50 (micrometre)/EC50 (nano) is .1. Only two TiO2
NM samples yielded NTR values ,1. Of these, Warheit
et al.[140] used nano- and micrometre-TiO2 samples coated with
alumina, and Hartmann et al.[127] reported difficulty in obtaining
a reliable EC50 value (241 mg L 1, Hombikat UV100). Therefore, weight of this evidence suggests that size or SSA (nano- v.
micrometre-TiO2) is involved at least in part with the toxicity
of TiO2 on algal growth, and that the TiO2 surface coating can
alter this cause–effect relationship. This size-related inhibition
of algal growth is not caused by a ‘shading effect’, as shown by
others[127,133,141] but is likely mediated by sorption of secondary
TiO2 NMs to algal cell aggregates by electrostatic surface charge
interactions.[130] Examination using light microscopy (phase
contrast and fluorescence), scanning (SEM) and transmission
electron microscopy (TEM) shows that TiO2 NM aggregates can
bind directly to algae (see also Fig. 3a, step 1), as well as the algal
aggregates (flocculated by algal secretion of exo-polymeric
substances in an extracellular matrix).[129,130,139,142] In fact,
large TiO2 NM aggregates trapped more algal cells than the
bulk-TiO2 aggregates,[133] and can occur during incubation
under UV-light as well as in the dark.[127,139] Interestingly,
adsorption of TiO2 NMs to algae may also have novel detrimental food chain effects; Campos et al.[143] report binding of TiO2
NMs to algae and sedimentation potentially leading to food
depletion in Daphnia magna, a filter feeder, and resulting effects
on life history traits including reproduction.
The manner in which TiO2 NMs interact with cell surfaces is
relevant to their toxicity in aquatic invertebrates, fishes and
algae. It is possible that cytotoxicity can be caused by cell entry
of small TiO2 NMs or agglomerates (presumably by endocytosis; Fig. 3a, step 2) associated with damage to the cell wall or
membrane.[135,139] Nonetheless, recent evidence suggests that
direct contact between TiO2 NMs and cell surfaces may be more
important to toxicity, especially becasue the bioavailability of
TiO2 NMs may be low. Lin et al.[135] studied the effects of nanoanatase on Chlorella sp. incubated in OECD media for 4 days.
They reported that the 96-h EC50 (growth inhibition) of pristine
anatase was 4.9 mg L 1, which increased to 47 mg L 1 in the
presence of humic acid (HA). Using SEM, they observed that
algal cells that were incubated with nano-anatase for 24 h
became flattened and damaged when compared to control cells
(no TiO2 NM added). This TiO2 NM effect was not observed
when TiO2 NM-treated cells were incubated in the presence of
HA, suggesting that cytotoxicity depended on the direct contact
between TiO2 NM and algal cells (Fig. 3a, step 3). In more
complex aquatic organisms, toxicity appeared most often to be
related to binding of the NM to sensitive external epithelia, in the
absence of significant tissue accumulation. Toxicity, including
effects on growth and reproduction, was associated with occlusion of the gut with TiO2 NM aggregates and decreased feeding
in Daphnia magna, a filter feeder.[144] Furthermore, effects
observed in the internal tissues of rainbow trout exposed to
1 mg L 1 TiO2 NM for 14 days were most likely explained by a
214
Product (Supplier)
18 particle
Toxicity test details
Nominal EC50 values
(mg TiO2 L 1)
NTR
Ref.
Purity
Size
(nm)
SSA (BET)
(m2 g 1)
Species
Test media and
method
142
38.5
Pseudo-kirchneriella
subcapitata
OECD 201
0.76
[140]
n.r.
5.8
72-h EC50 (cell count)
21 (95 % CI 16–21)
16 (95 % CI 12–22)
Nano-TiO2 (Sigma–Aldrich)
79 % rutile, 21 % anatase,
7 % alumina, 1 % amorph. silica
99 % TiO2 and 1 % alumina
coating
n.r.
25–70
n.r.
Pseudo-kirchneriella
subcapitata
[133]
n.r.
n.r.
n.r.
Nano-anatase (Sigma Aldrich)
99.7 % anatase
,25
200–220
OECD 201
2.09
[142]
Micron-TiO2 (Merck, India)
n.r.
Pigment white
n.r.
Scenedesmus sp.
(from a freshwater
lake, India)
72-h EC50 (fluorescence)
5.83 mg Ti L 1 (95 % CI 3.75–7.58)
(recalculated 9.71 mg TiO2 L 1)
35.9 mg Ti L 1 (95 % CI 31.4–41.7)
(recalculated 59.90 mg TiO2 L 1)
72-h EC50 (chlorophyll)
21.2
44.4
6.16
Bulk TiO2 (Riedel-de Haën)
OECD 201
(Enhanced
irradiation
400–500 nm)
Nano-anatase (Sigma–Aldrich)
99.7 % anatase
,25
200–220
OECD 201
[142]
n.r.
Pigment white
n.r.
72-h EC50 (chlorophyll)
16.1
35.5
2.20
Micron-TiO2 (Merck, India)
Chorella sp.
(from a freshwater
lake, India)
Hombikat UV100, (Sachtleben)
67.2 % anatase
32.8 % amorph.
72.6 % anatase
18.4 % rutile
9 % amorph
anatase
anatase
,10
288
Pseudo-kirchneriella
subcapitata
Modified
ISO 8692
0.60A
[127]
,30
47
72-h EC50 (fluorescence)
241 (95 % CI 95.6–609)A
71.1 (95 % CI 59.4–85.1)
,300
5–10
11.5
324
n.r.
n.r.
7
anatase
25
n.r.
Product 2. (Supplier: n.r.)
anatase
100
n.r.
Small particle. (Supplier: n.r.)
.90 % purity (phase not reported)
8
250
Large particle. (Supplier: n.r.)
.90 % purity (phase not reported)
150
8
Ultra-fine C (DuPont)
Fine TiO2 (DuPont)
215
P25 (Degussa)
Hombitan LW-S (Sachtleben)
HR3-anatase (Zhejian Hongsheng
Material Technology Co, China)
Bulk-TiO2 (Zhejian Hongsheng
Material Technology Co, China)
Product 1. (Supplier: n.r.)
A
Chlorella sp
Desmodesmus
subspicatus
Pseudo-kirchneriella
subcapitata
Author reported some difficulty to obtain a suitable concentration–response relationship to estimate the toxicity endpoint.
China State
Environ Protect
Bureau Guideline
201
Pre-irradiate with
simulated sunlight
(25–500 W) for
15–30 min, then
OECD 201
1 min stir, 3 min
sonication, then
OECD 201
145 (95 % CI 112–188)
144-h EC50 (cell count)
120
EC50 not reported because no
toxicity was seen at 1000 mg L 1 A
72-h EC50 (fluorescence)
44 (95 % CI 30–94)
EC50 (Particle 2) was not
reported. No clear concentrationeffect A
72-h EC20 (fluorescence)
16.5 (95 % CI 11.1–21.6)
29.8 (95 % CI 27.5–31.9)
2.04
.1.0
[214]
.1.0
[141]
1.81
[215]
Aquatic toxicity of manufactured nanomaterials
Table 1. Effects of titanium dioxide (TiO2) nanomaterial (NM) size on algal growth using different algal toxicity test methods
Nano toxicity ratio (NTR) compares median effective concentration (EC50) values between small (e.g. nanosized based on primary particle size) v. large (e.g. bulk for micrometre-sized) particles, and is defined as
EC50 (large)/EC50 (small). Toxicity ratios .1 indicate that the smaller particle is more toxic than the larger particle. Different EC50 time points are formatted in bold, with calculated values immediately beneath, including
confidence intervals (CI) in parentheses. n.r., not reported in cited article; SSA (BET), specific surface area determined by the Brunauer–Emmett–Teller method
A. G. Schultz et al.
developing systemic hypoxia arising from occlusion of the gill
by TiO2 NMs and resultant gill injury.[145] This hypothesis has
since been given further credence by the observation that
injection of TiO2 NMs into caudal vasculature, an approach
that avoids the issue of bioavailability, resulted in no significant
effects in internal tissues, including the kidney, the principle site
of accumulation.[146,147] In spite of their apparent low bioavailability, the interaction of TiO2 NMs at external epithelia may
have resulted in important ecological effects. For example,
Ramsden et al.[148] report decreased fecundity in zebrafish
exposed to TiO2 NMs, even though the accumulation of the
NM was not observed.
It is possible that environmentally released TiO2 NMs (those
residing at the top part of the water column) may be exposed to
natural sunlight and UV irradiation. Current data are scant,
however, elevated toxicity has generally been observed in
organisms where exposure to TiO2 NMs has been coupled with
UV irradiation. In zebrafish, the presence of sunlight has been
shown to increase the toxicity of TiO2 NMs by several orders of
magnitude.[149] Importantly, the increased sensitivity of zebrafish to photo-activated TiO2 NMs may be apparent only during
later developmental stages and this may be overlooked in typical
tests of 72 or 96-h duration post fertilisation. Elevated toxicity of
photo-activated TiO2 NMs appears to be linked to ROS generation because different solar UV spectra induced distinct intracellular ROS in Daphnia (Fig. 3a, step 4 and 5). Studies have
also shown that TiO2 NM toxicity is enhanced under simulated
solar radiation (UVA) as compared to ambient laboratory
lighting[150,151]; a linear correlation between ROS production
and D. magna immobilisation was also observed. Similar effects
have not been observed in algae; however, studies have generally pre-illuminated P25 samples to avoid the direct toxicity of
UV light to algae.[128] The 72-h EC50 data indicated that UV preirradiation did not alter the inhibitory effect of P25 on growth
(control: visible fluorescent lighting). Similar observations were
reported by Hund-Rinke and Simon[141] using anatase preirradiated with simulated sunlight (Table 1). These consistent
negative results suggest that photo-activation of TiO2 NMs is
not involved in algal growth inhibition; however, there was no
confirmation that P25 was photo-activated under these experimental conditions. It is also possible that concentrations of
bioavailable photo-activated TiO2 NMs or their activation byproducts (e.g. ROS) proximal to the algal cell wall were not
sufficient to yield a measurable effect.
Although an effort was made to include results of the recent
literature in order to interpret prior ecotoxicity data, interpretation of endpoint data (such as EC50 analysis) should not be done
in isolation of the morphological changes of TiO2 NMs
(agglomeration). Appropriate experimental design will allow
for the control of these dynamic processes and reduce the interexperimental variability that is presently an issue for risk
assessors and regulators of this emerging NM.
Zinc oxide is a widely used metal-oxide NM with applications in many commercial products including industrial coatings
and pigments, electronics, bioremediation and personal care
items. The most recent estimate from 2010 reported global
production of ZnO NMs in excess of 30 000 metric tonnes,
and is predicted to have since increased further.[152] The broad
utility of ZnO NMs is attributable to their optical, piezoelectrical and antimicrobial properties, which are unique to, or
enhanced on, the nanoscale (see Xu and Wang[153] for an
excellent review of the properties of ZnO NMs). Zinc oxide
NMs are wide-bandgap semiconductors with a bandgap energy
of 3.3 eV at room temperature,[154] which offers potential for
their use in: photovoltaic cells for solar energy harnessing,[155]
the photocatalytic degradation of organic pollutants[156] and the
attenuation of light in the UVA–UVB range (peak absorbance
,370 nm), which is integral to the effectiveness of ZnO NMs in
sunscreens.[157] Uncoated ZnO NMs are also soluble in aqueous
media (unlike TiO2 NMs, and more so than Ag NMs) releasing
Zn2þ from the NM surface.[158,159] Although this offers further
commercial applications of ZnO NMs, including better delivery
of Zn2þ as a nutritional supplement,[160] it can also be detrimental in aquatic environments because Zn2þ is a well-characterised
toxicant at higher concentrations in aquatic organisms, e.g. fish
(see recent review by Hogstrand[161]). For this reason, Zn2þ
levels in the environment are tightly regulated in most jurisdictions (e.g. in Canada, by the Canadian Council of Resource and
Environment Ministers).[162] Accordingly, a major challenge in
studies of ZnO NMs has been to discriminate between effects
attributable to nanoscale specific properties of ZnO NMs,
including the Trojan horse hypothesis (Fig. 3b, step 6) and
effects caused through the release of Zn2þ (Fig. 3b, step 7), in
order to establish if ZnO NMs present a novel risk to aquatic
organisms.
Overall, ZnO NMs have been shown to be toxic to a diverse
range of both fresh and marine aquatic organisms including
algae, molluscs, crustaceans and fishes.[163] A key issue emerging in the literature for ZnO, as well as other metal NMs, is the
importance of the physico-chemistry of the test media to the
behaviour of ZnO NMs, and its influence on toxicity (see NM
transformation in aquatic test media). In suspensions, uncoated
ZnO NMs aggregate, sediment and solubilise, and equilibrium
Zn2þ concentrations are affected by pH, particle size (especially
where ZnO NM , 6 nm) and ionic strength of the media.[158,159]
Toxicity in marine organisms has often been closely correlated
with the concentration of released Zn2þ. For example, cell
growth inhibition (measured as decreased cell numbers and
chlorophyll concentrations) of the marine phytoplankton
Thalassiosira pseudonana exposed to ZnO NMs was closely
correlated with the free Zn2þ concentration in suspension.[164]
This observation may also extend to freshwater organisms[165];
the toxicity of ZnO NMs to freshwater crustaceans was closely
correlated with Zn2þ release.[166] Dissolution may also underpin
observations of increased toxicity on the nanoscale compared to
bulk ZnO[165,167] because the rate of dissolution is closely
correlated with particle surface area.[159]
An important emerging effect of ZnO NMs (and other metal
oxide NMs) is in early life stage fishes. In particular, researchers
from several groups have reported decreased hatching of zebrafish embryos exposed to ZnO NMs.[47,168–170] For example, Ong
et al.[47] reported a decrease in hatching of zebrafish exposed to
10 and 100 mg L 1 of ZnO NMs that eventually culminated in
the mortality of unhatched embryos. Interestingly, where ZnO
NM exposed embryos were mechanically freed from their
Zinc oxide NMs
The toxicity of ZnO NMs to aquatic organisms has recently been
the subject of several excellent reviews[8,112,113] that offer
greater scope to the subject than is afforded here and to which we
refer readers seeking a more comprehensive overview of the
literature. Rather, the intention here is to highlight studies that
have identified: (i) properties of the ZnO NMs that are important
to toxicity in aquatic organisms; (ii) likely environmental scenarios of importance and (iii) recommended approaches for
more robust toxicity testing and future avenues for investigation.
216
Aquatic toxicity of manufactured nanomaterials
chorions, larvae exhibited no treatment dependent developmental deformities or significant effects on larval behaviours
(including spontaneous activity) leading the authors to conclude
that it was the process itself rather than embryonic development
that was sensitive to the ZnO NMs.[47] The hatching process in
zebrafish is initiated by the release of a protease, zebrafish
hatching enzyme (ZHE1), into the embryonic fluid, and it
culminates with zebrafish movements physically disrupting
the ZHE1 weakened chorion.[171] Although Ong et al.[47]
observed no effect on the movement of hatched larvae, ZnO
NMs inhibited protease activity in isolates of chorionic fluid
extracted from zebrafish embryos in vitro. This specific inhibition of ZHE1 by ZnO NMs has also been further elucidated
since by Lin et al.[172] who reported the in vitro inhibition of
the proteolytic activity of recombinant ZHE1 expressed in
Escherichia coli.
The mechanistic basis for hatch impairment in embryos
exposed to ZnO NMs has been debated in the literature, but a
weight of evidence implicates Zn2þdissolution (Fig. 3b, step 8).
Although NMs have been demonstrated to inhibit enzymes in
vitro through a particle specific mechanism,[108] physical separation of ZnO NMs from embryos with a Zn2þ porous membrane
did not ameliorate a zebrafish embryo hatch.[172] Furthermore,
the use of a metal ion chelator and the doping of ZnO NMs with
Fe, which decreased Zn NM dissolution, both improved hatching rates.[170] In contrast, Ong et al.[47] reported no hatch
impairment in embryos exposed to Zn2þ at dissolution control
concentrations. Nonetheless, dissolution analysis was performed in ultrapure water whereas dissolution will likely be
greater at the higher ionic strength of the tap water used in the
experiments and also likely in the high ionic strength environment of chorionic fluid where nanoscale materials have been
observed.[173] Moreover, the study of Ong et al.[47] is one of very
few to directly compare the toxicity of uncoated and coated NMs
where they found minor effects on hatch in embryos exposed to
polyacrylic acid stabilised ZnO NMs, further suggesting that
Zn2þ dissolution dominates as the source of hatch impairment
and therefore it is difficult to define the role of colloidal NM
interaction with the hatching enzyme in Zn2þ-containing
NMs.[47]
Although the role of released Zn2þ towards toxicity has not
been fully resolved, Zn2þ appears to have greater bioavailability
than nanoscale ZnO. Whereas binding of NMs to cell membranes has been postulated to contribute to toxicity,[2] empirical
evidence of ZnO NMs crossing cell membranes has been scant
to date. Where the uptake of ZnO NMs has been measured
directly, accumulation of Zn by aquatic organisms appears to be
strongly correlated with Zn2þ, including in benthic organisms
where sedimentation could lead to elevated localised concentrations of ZnO NMs. Using a mass balance approach to the
partitioning of 68Zn-labelled ZnO NMs in a simplified marine
system with the benthic invertebrate Corophium volutator,
Larner et al.[174] reported that 97 % of 68Zn was associated with
the sediment with relatively little in water (2.5 %). Less than
0.5 % was accumulated in C. volutator after 10 days. Nevertheless, the partitioning of ZnO NMs to the sediment was strongly
correlated with dissolution: almost identical mass balances of
Zn were observed in systems where 68Zn2þ was added (as
ZnCl2). Moreover, Zn accumulation in C. volutator was strongly
correlated with the dissolved phase; Zn bioconcentration factors
were identical for both ionic and nanoscale Zn and strongly
correlated with the concentrations of aqueous phase 68Zn2þ,
again strongly indicating that dissolved Zn2þ was the
bioavailable Zn fraction. Other studies with benthic organisms
have also reported similar observations. Biodynamic modelling
of Zn accumulation in the estuarine snail Peringia ulvae
revealed comparable uptake and efflux rates of Zn2þ and ZnO
NMs suggesting that Zn2þ was the bioavailable fraction.[175]
Together these studies strongly indicated that the behaviour of
ZnO NMs in the environment, even during relatively short
periods (10 days in the study of Larner et al.[174]) may be little
different from Zn2þ.
Although these studies have undoubtedly demonstrated the
important role of Zn2þ release in the toxicity of ZnO NMs, the
high NM dissolution rates may potentially obscure other important facets of toxicity, including nanoscale specific properties.
Nonetheless, the pre-irradiation of ZnO NMs with visible, UVA
or UVB light had no observed effect on their toxicity to
freshwater algae.[128] However, it should be noted that the
experimental design did not include a bulk control for the
comparison of phototoxicity and dissolution of ZnO NMs
provided acutely toxic Zn2þ concentrations within 24 h.
Moreover, the minimal effect of shape (spheres, rods, needles) and size of the ZnO NMs on their toxicity to marine algae
was attributed to the rapid dissolution (and to a lesser extent
agglomeration) observed for all ZnO NMs.[176] Furthermore, at
elevated concentrations of ZnO NMs (10 and 80 mg L 1) no
corresponding increase in toxicity was observed over a 72-h
exposure.[176] Zinc oxide NMs rapidly attain equilibrium with
free Zn2þ in seawater (reportedly within 4 h in Peng et al.[176])
and thus increasing the concentration of ZnO NMs above
10 mg L 1 may not have altered the Zn2þ concentrations in
solution.[176] Although this is further evidence for the principle
role of Zn2þ in the toxicity of ZnO NMs, full characterisation of
the dissolution behaviour is essential in order to ensure that the
toxicity of ZnO NMs is not underestimated. Similar characterisation of NM behaviour is also relevant where the concentrations of the ZnO NMs are below solubility limits; comparable
toxicity of Zn2þ, ZnO NMs and bulk ZnO to sea urchin embryos
is perhaps unsurprising because complete solubilisation of both
nano and bulk ZnO will occur within 12 h.[177] To fully assess
nanoscale specific effects of ZnO NMs where high rates of
solubilisation are expected, investigators should seek to identify
biomarkers for ZnO NMs that are responsive during the very
short-term (hours) experiments.
Silver NMs
Silver is one of the most common elements manufactured into
NMs.[111,178] Most of the applications for Ag NMs are focussed
on the antibacterial and antifungal properties of the manufactured NM that result from the release of free silver ions
(Agþ). In aquatic systems, ionic forms of Ag can induce toxicity
in sensitive organisms at very low (mmol L 1) concentrations.[179,180] As a result, the threshold concentration of Ag in
water quality criteria and guidelines is low relative to other
metals.[181] The combination of high production volumes for Ag
NMs, uncertainty about the bioavailability, uptake and stability
of NM forms of metals and a low toxicity threshold for Agþ has
resulted in an elevated level of concern over the potential
environmental effects (see the reviews of Fabrega et al.[111] and
Reidy et al.[178]). Although there has been considerable effort
towards characterising Ag NM toxicity, a comprehensive
understanding is still lacking, due in part, to the complexities of
the particle behaviour in aquatic media. As for other NMs,
factors that have been identified as being important for Ag NM
toxicity include its size, shape and surface area, its behaviour in
217
A. G. Schultz et al.
environmental media (including agglomeration, oxidation and
dissolution), as well as the influence of coating or capping
agents and manufacturing impurities.[178] A mechanistic
understanding of the Ag NMs is further complicated by the
observation that its behaviour and toxicity depends on
the geochemistry of the exposure medium and the age of the
material.[113,182,183]
Research efforts towards resolving the uncertainties about
the mechanisms of Ag NM toxicity have focussed on the
adsorption and uptake of intact NMs (Fig. 3c, step 9), the
transformation and dissolution (oxidation) of Ag particles in
solution (Fig. 3c, step 10) and possible biological interactions
(see the reviews of Shaw and Handy[112] and Levard et al.[184]).
Some studies assign the toxic effects to Agþ dissolution in the
test media[185–188] (Fig. 3c, step 10) or within the cells (known as
the ‘Trojan horse effect’,[188–191] Fig. 3c, step 11), whereas
others conclude that intact NMs induce responses[189–192]
(Fig. 3c, step 12). Many of these studies take a comparative
approach and include treatments with both Ag NMs and Agþ at
similar concentrations in order to allow a direct comparison of
concentrations that induce responses. Both acute and chronic
sub-lethal toxicity have been examined. A few of the studies
exploit the fact that the mechanism of Agþ toxicity occurs by
blockade of Naþ uptake as a result of Naþ/Kþ-ATPase inhibition on respiratory surfaces in freshwater fish[179,193] (Fig. 3c,
step 13) or by Agþ biouptake over a CuI transporter in
algae[188,194,195] (Fig. 3c, step 14). These works apply physiological study design elements focussed on mechanistic uptake
measurements to investigate mechanisms of Ag NM uptake and
toxicity.
Studies that compare Ag NM responses to those of Agþ can
be loosely grouped into two categories. The first includes studies
that show responses at elevated Ag NM exposure concentrations
relative to the Agþ response concentrations. The second group
includes those studies where responses occur at similar Agþ
concentrations or at relatively low concentrations, for example
near water quality guidelines and criteria values. In works where
Ag NM toxicity threshold concentrations are elevated and are
above those for Agþ in the same organism, authors often
implicate dissolution, and release of Agþ as a significant causal
factor. For example, Hoheisel et al.,[186] working with Daphnia
magna and fathead minnows exposed to uncapped Ag NMs
found that EC50 and EC20 concentrations (ranging from 46 to
89 mg Ag L 1) were ,5-fold higher for NMs than for Agþ, and
therefore concluded that particle dissolution was the likely cause
of Ag NM toxicity. Jo et al.[196] compared a variety of solution
preparation methods for uncapped Ag NMs and reported 24-h
EC50 values as low as 4.2 mg Ag L 1 but that spanned three
orders of magnitude. Nonetheless, dissolved Ag concentrations
in solution (as determined by ultrafiltration) explained toxicity
and thus they concluded that particle dissolution was responsible for the observed effects. Studies with Daphnia magna
exposed to Ag NMs coated with polyvinylpyrrolidone (PVP)
noted reduced Naþ influx and disruption of Naþ regulation at
500 mg Ag L 1, however this effect was also attributed to dissolution of Agþ given that is was not seen when cysteine was
added to the medium (cysteine–Agþ complexes are not bioavailable[197]). In some of these studies, there is an indication
that both NMs and Agþ induce toxicity, and it is generally not
possible to distinguish one potential cause from another. For
example, although effective concentrations were relatively high
(ranging from 1.3 to 10.6 mg Ag L 1), Laban et al.[198] concluded
that both NMs and dissolved forms of Ag were responsible for
acute toxicity (96-h LC50) in fathead minnow embryos. Similarly,
Massarsky et al.[199] determined an LC50 for zebrafish embryos
exposed to a polyacrylate capped Ag NM to be 1180 mg Ag L 1,
whereas for Agþ, the LC50 was 70 mg Ag L 1. Nonetheless, even
though the dissolution of particles was very low (0.5 %), it was
not possible to exclude the presence of Agþ as a possible
contribution to the observed toxic effects. Finally, although
Leclerc and Wilkinson[188] showed that when exposed to nAg,
Ag biouptake by Chlamydomonas reinhardtii exceeded what
was predicted based upon measured Agþ concentrations, their
measurements of darkfield–hyperspectral microscopy and
expression levels of the CuI transport protein (CTR2) indicated
that nAg increased Ag biouptake by locally increasing surface
concentrations of Agþ. In other words, Ag biouptake was
increased by oxidation of the nAg at the cell surface, but only
Agþ could be considered to enter the cells.
In the second group of studies Ag NMs are shown to induce
acute or chronic sub-lethal effects at similar concentrations to
Agþ. Often sensitive organisms are used and response thresholds are very low, at or near water quality guidelines and criteria
values. These studies generally reach the conclusion that Ag
NMs induce toxicity independently or in combination with Agþ.
For example, Das et al.[200] characterised acute toxicity of
carboxy-functionaliszed capped Ag NMs in Daphnia magna
(48-h EC50 of 2.75 mg Ag L 1) and concluded that responses
were NM specific as insufficient Agþ was present to induce
toxicity. Citrate capped Ag NMs, which had a very low
dissolution rate, were shown to independently reduce Naþ
uptake and inhibit Naþ/Kþ ATPase in juvenile rainbow trout.[46]
Allen et al.[185] found that the acute toxicity (48-h LC50) of
citrate coated Ag NMs and AgNO3 were similar (1.1 mg Ag L 1)
in Daphnia magna and also documented accumulated NMs
within Daphnia magna. Although it would seem reasonable to
suggest that NMs and perhaps dissolved Agþ were responsible
for Ag NM toxicity, the authors did not reach this conclusion.
Exposure of the freshwater mussel, Elliptio complanata, to Ag
NMs at 0.8, 4 and 20 mg L 1 resulted in a suite of sub-lethal
cellular responses, some of which were similar to Agþ but
others, for example the presence of ubiquitin in the digestive
gland, were unique to the NM.[201] Responses unique to Ag NM
exposure that do not occur in matched Agþ exposures are an
indirect way to distinguish between the potential toxicity of
particles and that of dissolved Agþ in solution.
Genotoxicity and cytotoxicity studies have contributed to our
understanding of particle v. Agþ responses. In these types of
studies, the effects of Ag NMs vary considerably depending on
NM characteristics, species and endpoints, but generally
responses are reduced when compared to those provoked by
Agþ (reviewed by de Lima et al.[202] and Kim and Ryu[203]).
Nonetheless, the pattern of responses can be used to demonstrate
particle specific effects that are distinct from those attributed to
Agþ. For example, gene expression profiles in zebrafish gills
have been shown to differ for Ag NMs and AgNO3, suggesting
that the biological interaction of the NMs is unique.[49] The
observation of unique responses between Ag NMs and Agþ was
also found by Powers et al.[204] for developmental neurotoxicity
endpoints in developing zebrafish, and by Poynton et al.[205]
for gene expression profiles in Daphnia magna. In the latter
study, the 48-h EC50 values for Ag NMs (citrate coated) was
1.8 mg Ag L 1 as compared to 0.04 mg Ag L 1 for AgNO3.[205]
In the study by Pham et al.,[206] expression levels of the genes
coding for metallothionien and glutathione were induced
by exposures at 1 mg L 1 of Ag NM but not to exposures of
218
Aquatic toxicity of manufactured nanomaterials
1.6 mg L 1 of AgNO3, indicating that NMs directly increase
metal detoxification and antioxidant defence responses.
Hepatic expression of stress related genes in medaka also
differed between exposures to either Ag NMs or AgNO3 (both
at 1 mg Ag L 1) after the first day of exposure but were similar on
subsequent days.[207]
A key research question is whether intact Ag NMs induce
toxic responses independently or whether they arise from Agþ
from NM dissolution.[112,178] From the data currently available
it would appear that both mechanisms can occur. If particles
undergo transformation and dissolution to release Agþ to a
sufficient degree then toxicity can be induced. The relatively
high inherent toxicity of Agþ means that only a small degree of
dissolution need occur to produce toxicity at intermediate Ag
NM concentrations (e.g. tens to hundreds of milligrams of Ag
per litre). When Ag NMs are stable (Agþ release is low) it may
be possible for NM induced toxic effects to occur, perhaps at low
concentrations in sensitive organisms.[46,200,208] A third scenario is that stable Ag NMs are not bioavailable and in this scenario
very high concentrations may be required to induce toxicity.
Greatly compounding the interpretation of the studies seeking to
distinguish the effects of NMs and free ion is the perhaps
counterintuitive observation that particle dissolution increases
greatly at lower particle concentrations.[35,41] Such an observation again reinforces the contention that Ag NMs must be
characterised under the precise experimental conditions of the
exposure.
Ultimately, one of the key questions is whether environmental protection criteria and guideline values, which are measured
either as total (e.g. Environment Canada) or total dissolved Ag
(e.g. USA Environmental Protection Agency) are sufficiently
protective.[186] In this regard, most studies with Ag NMs
show effects thresholds above water quality guidelines and
criteria, however there are a few that show effects at levels at
or below regulatory values. Clearly, understanding these Ag
NMs should be a priority. For example, the gastropod snail
Physa acuta appears to be particularly sensitive to Ag NMs. In
chronic exposures, reproductive impairment (egg production)
occurred at and above 0.01 mg Ag L 1.[209] Stress responses
and thyroid hormone signalling dynamics in the frog (Rana
catesbeiana) were also disrupted by Ag NMs at very low levels
(0.6 mg Ag L 1).[210] The significance of these studies and
the determination of whether water quality guidelines are
sufficiently protective should be of primary consideration.
Experiments performed under simulated sunlight as appropriate for the properties
Investigations of toxicity over longer-term exposures
Investigations of toxicity of NMs to marine organisms,
especially phytoplankton
Key gaps in ZnO NM research
Experiments with numerous controls (e.g. ionic metal, bulk
material and NM sizes)
Determinations of the relative roles of Zn2þ and ZnO NMs on
biouptake
Determination of ZnO solubility in complex toxicological
media by carefully distinguishing among Zn2þ and Zn
complexes
Determination of the role of surface coatings on fate, bioaccumulation and bioavailability
Determination of the trophic transfer of ZnO NMs in aquatic
ecosystems
Investigations of ZnO toxicity under simulated sunlight as
appropriate for the properties
Investigations of the toxicity of ZnO NMs to marine organisms, especially phytoplankton
Discrimination of the bioavailability and fate of natural from
anthropogenic ZnO
Key gaps in Ag NM research
Experiments with numerous controls (ionic metal, metal
complexes)
Further comparisons in toxicity among Ag NMs, ionic Agþ
and other Ag complexes, such as Ag nanoclusters; mechanisms of Ag uptake
Bioaccumulation and trophic transfer of Ag NMs in aquatic
ecosystems
Investigations of Ag NM transformations in aqueous
systems and discover how Ag NM transformations affect
bioavailability
Identification of the role of Ag NM size on its transformations
Investigations of the toxicity of NMs to marine organisms,
especially phytoplankton
Early nanotoxicology studies included minimal NM characterisation limited to their pristine state, within stock solutions. It is
now clear that NMs must be characterised as they are found in
the aquatic test media, such as cell culture media. Interpretation
of data from these early studies is, therefore, challenging and
difficult to compare with many (but not all!) modern studies.
Whereas there is ongoing debate regarding which NM properties require routine characterisation in nanotoxicological studies, it is recommended that NM size, agglomeration, crystal
structure, composition, surface coating and solubility be considered.[211] Furthermore, the degree and type of characterisation required should be NM specific. Researchers should also
attempt to characterise the NMs throughout the duration of the
toxicity experiments (start, middle and end). The time course
NM characterisation will provide researchers with a clear indication of any alterations in the NM properties that may have
occurred during the investigation. In addition, there should be a
focus on reporting important chemical components such as pH,
ionic strength and the presence of nutrients (as found in cell
culture media) or proteins that may alter the physiochemical
properties of NM. Greater characterisation of both NMs, as well
Data gaps and recommendations
In this section, data gaps and recommendations have not been
placed in any particular order of importance and are intended to
serve as a guide to improve future experimental design, execution and data reporting. For each of the three NMs above, key
recommendations, identified by the NSERC-NRC-BDC consortium are provided below in their order of importance. Each of
these points and others discussed below are critical to the
advancement of the field of aquatic nanotoxicology and are vital
to allow future comparisons among NMs and among experimental systems. This information is crucial for regulators in
order to allow the development of quantitative structure–property relationships in predictive toxicology.
Key gaps in TiO2 NM research
Thorough characterisation of TiO2 NM agglomeration over
the duration of the toxicological study
219
A. G. Schultz et al.
as the aquatic test media used in toxicity studies will allow more
direct comparisons between studies and will enhance our
understanding of the toxicity of the NMs.
Suitable controls are essential in any NM toxicity study but
dependent on the specific type of NM being investigated. The
importance of ionic metal controls has previously been discussed (see sections Zinc oxide NM and Silver NM); however, a
bulk material (micrometre sized) control should also be considered for metal and metal oxide NM. To date, particle size
controls have been limited primarily to studies assessing the
aquatic toxicity of TiO2 NMs (see section Titanium dioxide
NMs), with few other studies (e.g. Zn and Ag NMs) incorporating such controls. The inclusion of NM size controls will enable
nano-specific effects to be more comprehensively investigated
and permit more informative comparisons to be made between
studies. This information will improve our understanding of
NM toxicity to aquatic organisms and consequently enable risk
assessors to predict more accurately the types and specific
properties of NMs that incite toxicity.
In the presence of UV, some NMs possess the unique
ability to undergo quantum confinement (bandgap semiconductors[91–94,212]), and catalytic activity (generation of
ROS[92,100,213]), because of their small size and unique properties. These UV-mediated NM transformations have the potential
to enhance the toxicity of the NMs to aquatic organisms,
therefore, future studies should consider the effects of UV
exposure on toxicity levels. Finally, NM toxicity may be
enhanced or diminished over time, therefore, chronic toxicity
involving bioaccumulation and trophic transfer of NMs should
be studied to gain a greater understanding of the effects NMs
may have on entire aquatic ecosystems over time.
In summary, the field of nanotoxicology has recently made
great gains in our understanding of the toxicological properties
for these specific OECD relevant materials. It was a great early
challenge to adapt toxicological principles and techniques
designed for well dispersed toxicants to the nascent field of
colloid (nano) toxicology. Specifically, alteration in experimental design to identify the different behaviour of these materials as
colloids, developing an ability and appreciation necessary to
understand the chemical and physical properties of these materials in solution and developing and understanding how these
materials may interact with organisms were essential for proper
and accurate toxicological assessments. This review summarised
the fundamental gains in our understanding on how to do and
interpret toxicology for a few high volume NMs in commercial
applications today and we have outlined some key gaps in
knowledge and recommendations for study design that are
essential to perform. Moreover, we will continue to be challenged by new materials and nanoformulations entering the
marketplace. Training toxicologists with a firm multidisciplinary understanding of the physics, chemistry and biology of these
materials is essential for accurate toxicological assessment.
[4]
[5]
[6]
[7]
[8]
[9]
[10]
[11]
[12]
[13]
[14]
[15]
[16]
[17]
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