6. What are the potential environmental effects of nanomaterials?
- 6.1 What happens to nanomaterials in the environment?
- 6.2 What makes nanomaterials active after release?
- 6.3 Environmental effects
6.1 What happens to nanomaterials in the environment?
Inevitably, production, use and disposal will lead to releases to the environment. Wastewater treatment streams, landfill and combustion of products containing nanomaterials are means through which they may end up in the environment, although it is most likely that they do so as modified forms from their primary counterpart. In addition, some nanomaterials are used in environmental remediation applications and as such they are applied as primary nanomaterials to the environment.
The environmental fate and behaviour of nanomaterials has been recently reviewed by Klaine et al. 2008. Knowledge from colloid science can provide information on the likely fate and behaviour of nanomaterials (Lead and Wilkinson 2006). The behaviour of nanoparticles in the environment is expected to depend not only on the physical and chemical character of the nanomaterial, but also and perhaps predominantly on the characteristics of the receiving environment (Chen et al. 2008, Chen and Elimelech 2008, Saleh et al. 2008). It is generally known that small particles tend to aggregate or agglomerate to eventually become associated with other dissolved, colloidal and particulate matter present in the environment. Upon entry into the environment, nanoparticles may remain intact or undergo one or more of the following:
- speciation (i.e. association with other ionic or molecular dissolved chemical substances)
- biological or chemical transformation to other chemicals, and/or complete mineralization (to carbon dioxide and water)
So far, there are no peer reviewed publications providing information on concentrations or amounts of nanomaterials in environmental compartments such as surface waters and soils. Estimates of quantities of nanomaterials present in surface waters and other media derive from calculated exposure scenarios based on predicted nanomaterial use and not from actual measurements (Boxall et al. 2007, Mueller and Nowack 2008). Methods for measuring nanomaterials in specific environmental matrices are being developed for various materials (Christian et al. 2008, Hassellöv et al. 2008, Tiede et al. 2008). The appropriate metrics of the measurement of manufactured nanomaterials in relation to environmental risk assessment is still under discussion.
In order to assess the potential effects of nanomaterials in the environment, exposure concentrations or doses should be considered realistically. All forms, in which the nanomaterial occurs, not only the free nanoparticulate form, but all physical and chemical species, should be considered. It is important to realise that there may be some ‘hot spots’ where nanomaterials might concentrate due to their tendency to aggregate/agglomerate and potentially to adsorb to or associate with organic matter. In addition, it is likely that some of the nanomaterials going through the standard waste treatment stream will end up associated with the solid phase and then potentially be deposited in certain areas of the environment where they might reach higher loads.
In the environment nanomaterials are expected to occur mainly associated with sediments and soils (Baalousha et al. 2008, Klaine et al. 2008). The free dispersed form of nanomaterials is of particular importance and is addressed here specifically. It is recognised that for the estimation of the possible presence of free nanoparticles knowledge on release scenarios is very important. Unfortunately, to date, there is no suitable information available on this topic. Examples of exposure routes for nanomaterials via the environment are inhalation by humans, and other air-breathing species, and uptake by aquatic organisms from water or sediments.
Assessment of exposure concentrations of dispersed nanomaterials requires detailed insight into the processess that act on these materials in the environment. However, currently available knowledge of these processes is insufficient to allow quantitative predictions of the environmental fate of nanomaterials.
Information concerning the presence of nanomaterials in air is summarised in section 4.3.
Upon release to water, dispersed nanomaterials are expected to behave according to well-understood phenomena described and explained in colloid science (Jones 2002, Lyklema 2005). Colloidal suspensions of nanomaterials are generally expected to be unstable: i.e. upon collision, particles may approach each other close enough for attractive Van der Waals forces to become dominant over repulsive electrostatic forces and steric hindrance. As a consequence, particles adhere to each other and then settle due to gravity (Baalousha et al. 2008, Ju-Nam and Lead 2008, Saleh et al. 2008). Moreover, natural waters contain many other dissolved, colloidal and solid materials (including natural nanomaterials) to which nanomaterials can and usually will adhere. Suspensions of dispersed nanomaterials are stable only under narrow ranges of environmental conditions (Baalousha et al. 2008).
The dominant factors in colloid stability under natural conditions are known to be pH, ionic strength and presence of natural organic matter (Lead and Wilkinson 2006). In sea water with high pH and ionic strength, electric double layers of colloid particles are much smaller than in freshwater, allowing for closer interparticle approach, which usually leads to more aggregation. In addition, the intrinsic properties and characteristics of the materials, including their specific chemistry, will influence their fate and behaviour. The surface properties of the nanomaterials are very important for their aggregation behaviour, and thus for their mobility in aquatic and terrestrial systems, and as such for their interaction with and general bioavailability to organisms.
The humic and fulvid acids of "brown waters" will cover the nanoparticles with a coating that keeps them probably longer and more dispersed (Hyung et al. 2007). For example, the presence of natural organic matter (NOM), as well as iron oxide, have a stabilising effect on aquatic suspensions of fullerene and carbon nanotubes, at least in fresh water systems (Baalousha et al. 2008, Chen and Elimech 2008, Christian et al. 2008, Saleh et al. 2008). A similar effect has been shown recently for nanoparticles of CeO2 (Quik et al. 2008). On the other hand, Baalousha et al. (2008) have shown that the effects of humic acids and varying pH can have combined effects on the fate of iron oxide nanoparticles with increasing pH resulting in a higher level of aggregation.
Likewise, surface modification of nanomaterials can influence the environmental fate and behaviour. As carbon nanotubes are considered to be highly hydrophobic and with a tendency to aggregrate, they would be expected to settle in the natural environment. However, Kennedy et al. (2008) have indicated that surface modifications, which are widespread (e.g. functional groupings and coatings) lead to increased dispersability, increased water column stability and lower settling rate, especially in combination with natural organic matter.
To understand the fate of nanomaterial dispersions in the environment it will be necessary to characterise the nanomaterial colloidal properties and the aqueous phase physical-chemical properties to a greater degree than is necessary for gas phase or dissolved substances (e.g. present as environmental contaminants). For example, moderate changes in the ionic strength will have little effect on the solubility or many organic substances (e.g. PAHs, most pesticides) but can have major effects on the suspension stability of nanoparticles. In saline environments nanomaterials have a tendency to aggregate (Nielsen et al. 2008, Stolpe and Hassellöv, 2007) and thus would most likely tend to settle. In sedimentary systems it would be important to determine how these nanomaterials might interact with organic matter and potentially be adsorbed and sequestered. This would have influence on their bioavailability and determine their biological uptake.
The estimation of concentrations in water is essential to environmental risk assessment. In sharp contrast with the situation for conventional chemical substances, there is neither theoretical nor empirical evidence that can be used to predict residual concentrations of nanomaterials in suspension under conditions of limited colloid stability. When substances enter the environment, they distribute themselves between the various phases of the system (partitioning). The environmental distribution of substances is often predicted by the octanol-water partition coefficient (Kow). However, there is no reason to assume that the Kow of the substance of which the nanoparticles are made of is predictive of the extent to which nanoparticles associate themselves with other particles. The Kow is probably not applicable to non-soluble nanomaterials for risk assessment purposes. Due to the interactions of nanomaterials with various components of the environmental system, generally near-zero concentrations of the nanomaterial in its original form would be expected. It is of great importance to gain understanding of the environmental conditions under which stable colloidal suspensions of dispersed nanoparticles can be formed.
For nanomaterials which may be solubilised, Kow could be applicable. Recently, Jafvert and Kulkarni (2008) have studied the octanol-water partition coefficient (log Kow) of buckminsterfullerene (C60) and its aqueous solubility. The authors obtained a value for log Kow of 6.7, and a value for the solubility of C60 in water-saturated octanol of 8 ng/L. Based upon this high Kow, it is expected that C60 has high affinity for lipids and organic matter, indicating that in the natural environment, C60 will tend to sorb to solid phases.
Some recent predictive modelling work has been published for TiO2 and silver nanoparticles and carbon nanotubes (Boxall et al. 2007, Mueller and Nowack 2008). However current knowledge of the behaviour of nanoparticles in natural waters provides insufficient basis for the full assessment of environmental exposure concentrations of dispersed nanomaterials. There is an urgent need to improve knowledge in this area.
It should be noted that in wastewater treatment plants partitioning of nanomaterials into the solid biomass is likely to be an important fate for hydrophobic materials which end up in the sewage stream.
Soil and sediments
As described above, depending on receiving environment and material, nanomaterials, if not degraded or dissolved, will tend to aggregate and eventually settle onto the substrate. Within soil and sedimentary systems it is expected that these materials will adhere to solids.
It is likely that the OECD test methods for a number of physico-chemical methods for environmental distribution are applicable, although this needs to be further assessed, taking into account the administration of the sample to the test system.
Methods for assessing the environmental distribution of nanomaterials have been described (Christian et al. 2008, Hassellöv et al. 2008, Klaine et al. 2008, Tiede et al. 2008;). They are progressively being developed so that the complex issues of fate in different media may be addressed. Nevertheless, much information is still needed in this area.
For reasons explained above, it is doubtful whether standard tests of vapor pressure, water solubility, octanol-water partition coefficient and ready biodegradability are adequate and sufficient to describe and predict the distribution of nanomaterials in the aquatic environment.
Vapor pressure and solubility
Vapor pressure and water solubility of conventional chemicals are used to predict air/water partitioning. If measurable at all for nanomaterials, these properties may not be very useful in predicting the extent to which nanoparticles may partition from water into air. However, solubility of nanomaterials in water, or rather: the rate of dissolution of nanoparticles in water is important for an entirely different reason. Toxic effects of the presence of nanomaterials may well result, at least in part, from the presence of dissolved species that originate from dissolution of the nanomaterials. To date, it is unclear to what extent the effects observed can be attributed to the dissolved form or to the nanoformulation the effect being a combination of fraction and size (Franklin et al. 2007, Navarro et al. 2008b). Nevertheless, Lin and Xing (2008) have suggested that phytoxicity observed on exposures to ZnO nanoparticles may not be attributed solely to dissolved zinc. Griffit et al. (2008) drew a similar conclusion with respect to nanosilver.
The OECD assay on water solubility (OECD 1995) may not be very useful in this context (OECD, 1995). Rather, standard measurement of rate and extent of dissolution under natural water conditions would be helpful. Many nano materials are highly insoluble in water, so that specialised methods are likely to be needed to measure or estimate their water solubility anyhow. For example, the solubility of fullerene is usually estimated by measuring solubility in alcohols and extrapolating to a zero carbon alcohol, i.e. water (Jafvert and Kulkarni 2008).
Environmental persistence of nanomaterials (i.e. resistance to transformation and degradation) depends on the chemical composition of both core and surface material. Although it is possible that most nanomaterials will be persistent in their original particulate form, this cannot be assumed in general. It seems likely that the organic coatings of nanomaterials are readily transformed or degraded, but there is lack of data in this area. As mentioned above, dissolution may occur for at least some metal nanomaterials (Franklin et al. 2007, Luoma 2008). Whether or not followed by degradation of the dissolved material, the process of dissolution makes nanoparticles disappear and become less persistent. In standard tests for ready biodegradability of chemical substances, either disappearance of dissolved organic carbon or the generation of CO2 is measured. Therefore, it is necessary to examine first whether the nanomaterial can be utilised as an energy or nutrient source for microorganisms. Secondly, the nanomaterial must be available to microorganisms in order to be degradable. If the material is unlikely to reside in the water column or if it is not soluble in water, biodegradation is unlikely and testing in surface water may be unnecessary.
For C-containing nanomaterials, the biodegradation screening methods (e.g. for ready biodegradability) measuring dissolved carbon are not applicable. In principle, the methods measuring carbon dioxide production or oxygen uptake are applicable, but they require large amounts of test material. It is also important to consider whether carbon based nanomaterials such as fullerenes and nanotubes can be degraded at all under any conditions. However, some data indicated that fullerenes could be taken up by wood decay fungi, suggesting that the carbon from fullerenes could be metabolised (Filley et al. 2005).
Simulation tests for biological degradation in various environmental compartments are applicable in principle, but again the detection of the nanomaterials is the challenge. The possible degradation to carbon dioxide, integration into biomass or other partitioning can be followed using labelled test material. However, it should be noted that the use of labels needs specific attention in terms of association of the label with the nanomaterial.
Likewise, for hydrolysis testing, the chemical structure of the material and whether it contains groups which could be subject to hydrolysis dictate whether this test is necessary or appropriate. In view of the sometimes very long lifetime of ecosystem processes, other non biological degradation mechanisms have to be investigated (e.g. UV induced, slow dissolution).
Current work assessing uptake has focussed on exposures in media with different nanomaterial loads over a specific time interval, followed by total body burden assessment, especially if species are small, such as Daphnia species, copepods or Lumbriculus (Fernandes et al. 2007, Petersen et al. 2008, Roberts et al. 2007). If organisms are larger, specific studies have focussed on detection, following exposures, of loads within specific organs, such as liver, kidney, muscle, gills (Handy et al. 2008a). In terms of detection, it may not always be possible to identify the form of such material. This may be particularly important for materials that may tend to get into solution such as silver (Luoma 2008, Navarro et al. 2008b).
From studies of biological exposures of nanomaterials it is clear that adsorption and aggregation of the material onto surface of the organism is commonly observed (Fernandes et al. 2007, Handy and Eddy 2004, Nielsen et al. 2008, Rosenkranz et al. 2009). This has also been shown by the aggregation of single wall carbon nanotubes on the gill mucus of rainbow trout (Smith et al. 2007) and of carbon black and titanium dioxide nanoparticles on the carapaces of Daphnia, (Fernandes et al. 2007), as well as the entanglement of macroalgae gametes by clusters of carbon black (Nielsen et al. 2008).
Given the tendency of nanomaterials to aggregate, and thus their likelihood to end up associated with sediment (Klaine et al. 2008) bioaccumulation studies on sediment organisms would be especially important. OECD has recently adopted a new standard test for the assessment of bioaccumulation into sediment worms using Lumbriculus variegatus. This method could be relevant to be used in a test battery for risk assessment as OECD has also published recently a toxicity test (OECD TG 225) based on the same species which could provide effects data (OECD 2008b).
For organic substances, there is an established relationship between octanol/water partition coefficient (Kow) and bioaccumulation or bioconcentration factor (BCF). However, this relationship may not hold true for nanomaterials.
The main challenge in testing the bioaccumulation of nanoparticles is their detection and characterisation in tissues and body fluids. Radiolabelling could make detection and quantification easy but it has also limitations; for example, the labelled material can behave differently from the non-labelled particles. Petersen et al. (2008a) used radio- labelled CNTs to assess uptake and depuration by Lumbriculus variegatus. Another possibility could be the radio-activation of metal and metal oxide nanoparticles (Oughton et al. 2008). It enables both localisation and quantification within tissues or organisms. This technique is still at experimental stage and a key aspect is how ionisation of manufactured nanomaterials may interfere with the exposure assay and any results.
Standard BCF testing protocols such as OECD 305 (OECD 1996) may have limitations in testing of bioaccumulation of nanoparticles. It has been observed for substances dissolved in water that a large molecular size effectively (MW > 600, or effectively a diameter size > 0.5 nm) limits direct uptake. It is likely that in most cases the relatively large size (1-100 nm) of nanoparticles compared to dissolved molecules limits their direct uptake by fish gills. Fish dietary bioaccumulation factor (BAF) testing (Fisk et al. 1998; Stapleton et al. 2004) is not a standard OECD testing protocol yet. This spiked food method is suitable for testing of poorly soluble, large molecules and might be suitable for testing several classes of nanoparticles, either by itself or in combination with the OECD 305 testing. However, more data using a harmonised OECD dietary protocol, especially for testing nanomaterials, are needed. The testing results of human health endpoints should also be taken into consideration if available when generating environmental testing plans for specific nanomaterials. Uptake studies from mammalian studies may give valuable basic information on uptake characteristics, rates and mechanisms of nanoparticles also in non-mammalian species.
6.2 What makes nanomaterials active after release?
Uptake by biota is likely to be via the respiratory or digestive tracts in animals, or via the root system in plants. Uptake across epithelial surfaces is also possible. Plants and fungi have cell walls that act as an initial barrier to the entrance of nanomaterials.
In terrestrial (and water) systems some nanomaterials may preferentially bind to NOM (see section 18.104.22.168) and thus become less bioavailable. Sediment feeders may be able to uptake these nanomaterials. In fact they may preferentially ingest them if they are associated with NOM (Roberts et al. 2007) and strip/de-associate them within the gastro- intestinal tract. Li et al (2008) have reported the reduced bioavailability, and resulting reduced antibacterial activity, associated with the increased sorption of C60 to soil organic matter .
In aquatic systems stabilisation by NOM may maintain nanomaterials within the water column which may result in increased bioavailability to aquatic organisms (Kennedy et al. 2008). Although such association with NOM also may reduce and even eliminate antibacterial activity (Li et al. 2008).
If the nanomaterial readily dissolves in water current protocols and guidelines developed to measure bioavailability of conventional chemicals are applicable, although in such cases the test would address the dissolved form rather than the nanoparticulate form. However, given that depending on the nanomaterial and the receiving environment the rate of dissolution could be quite variable, there is a possibility that a combination of nanoparticles (i.e. size) and substances (dissolved nanomaterial) elicits the detected toxic effects (Luoma 2008). Limbach et al. (2007) have shown that for partially soluble nanomaterials such as cobalt oxide and manganese oxide the nanoparticles may be taken up into cells preferentially to their respective ionic forms. It is then possible that once inside the cells these nanoparticles may dissolve, resulting in enhanced toxic effects. Regardless of the key causes, it should be considered that given the increased use, and thus release, of nanomaterials with different levels of solubility this will lead into potential increased levels of soluble substances which may result in undesirable environmental effects. These effects may be enhanced (depending on material, receiving environment and species) by a combination of forms, i.e. particulate and soluble.
For nanomaterials that do not dissolve readily, it should be determined whether they will form stable dispersions in air or stable suspensions in aqueous media (in both fresh and sea waters).
Exposure to nanomaterials in experimental studies
One of the major problems in aquatic ecotoxicological fate and effects testing is the absence of consistent and broadly-applicable information on how nanomaterials are suspended in various exposure media used in ecotoxicological testing. There are essentially three approaches to achieve as uniform as possible stock solutions for testing: dispersion with strong solvents and detergents, dispersion by sonication or dispersion by prolonged stirring (Crane et al. 2008, Klaine et al. 2008). Suspension methods used to date include the use of strong solvents e.g. tetrahydrofuran (THF) (Oberdörster 2004), dispersion agents e.g. sodium dodecyl sulphate (SDS) (Smith et al. 2007), bath or ultrasonication with filtration to remove aggregates (Lyon et al. 2006), stirring (Hund- Rinke and Simon 2006, Oberdörster et al. 2006), and combinations of these methods. Natural organic matter (NOM) can be a suitable dispersant and can keep the nanoparticles suspended longer than a 1% solution of SDS (Hyung et al. 2007). In addition, the time used during mixing (sonication or stirring) is also very variable (Klaine et al. 2008). Henry et al. (2007) has referred to some of the key issues regarding the use of solvents when dispersing nanomaterials in aqueous media.
In most cases, the verification of exposure and the characterisation of the nanomaterials in the resulting suspensions is limited to working on stock solutions rather than on the actual concentrations, either after a dilution series is generated or periodically over the duration of an exposure or media-renewal period. As such, it is likely that the methods reported might not produce similar results for different forms of a nanomaterial. In addition, similar to the testing for health effects also the possible temporal evolution of nanomaterials during the perfomance of the assays should be considered.
Food chain effects and secondary poisoning
Not much work has been published on potential food chain effects of nanomaterials. A recent study (Holbrook et al. 2008) on the possible transfer of quantum dots in a simplified aquatic food chain has found that these materials can be transferred to rotifers through dietary uptake of ciliated protozoans. Although there was transfer across these levels, bioconcentration (accumulation from surrounding environment) in the ciliates was limited and no biomagnification (enrichment across levels) in the rotifers detected. This study indicates potential for transfer across food chain levels but this would depend on material type and food chain, as is mostly the case for other studies of standard materials. Fortner et al. (2005) have also observed that fullerene nanoparticles accumulate in microbial cells, in worms eating those microbes and possibly in animals higher up the food chain. Furthermore, Bouldin et al (2008) have reported the transfer of quantum dots from dosed algae (Pseudokirchneriella subcapitata) to Ceriodaphnia dubia. Petersen et al. (2008b) have also indicated that CNTs were not readily bioaccumulated by the earthworm Eisenia foetida with results indicating bioaccumulation factors 2 orders of magnitude smaller than those measured for pyrene used for comparison.
6.3 Environmental effects
Ecotoxicological testing in soil and sedimentary systems has been the focus of relatively few studies. As a result, methodology and practical approaches have not been as widely discussed in the literature. Suspensions of nanomaterials in aquatic media, followed by either mixing or spraying on sediments/soils would lead to similar issues to the ones raised above. Other methods include mixing or applying nanomaterials directly to soils and sediments. It is clear that similar issues of standardisation also apply to these systems. There is a need, therefore, to address methodological variability associated with the current studies assessing the hazard of nanomaterials in environmental models. Although it is accepted that methods need to be appropriate to the materials being tested, as well as the test organisms and end points studied, it is important that standardised methology is developed and implemented so that variability can be kept to a miminum and results widely accepted and replicated.
In general, it is necessary to check that the suspensions used for aquatic testing are suitable for the test organisms. Salt concentration, pH, solvent and amount of solvent have to be within ranges tolerated by the test organisms. For organic and inorganic nanomaterials different procedures are usually applied. Inorganic nanomaterials (e.g. metals and metal oxides) are weighed into an aqueous solution; they are homogenously dispersed by ultrasound or by stirring and followed by filtration. Organic, water-insoluble nanomaterials are dissolved in solvents such as tetrahydrofuran (THF), toluene or benzene, or dispersed by detergents. By adding water and removing the solvent, a stable aqueous suspension is obtained. However, there are indications that traces of THF used to solubilise fullerenes may remain in the suspension resulting in toxicity due to the solvent (Zhu et al. 2006).
Environmental test systems
Uptake of nanomaterials by microbial organisms might be via diffusion, specific or non- specific uptake, or via membrane damage (Klaine et al. 2008). The bioavailability and antibacterial activity of C60 fullerene in soil and water were found to be affected by the concentration of humic acids (Li et al. 2008). In general, the higher the carbon content of the soil, the stronger the likely adherence to soil/sediment. A list of some microbial effects observed is presented in Table 1.
[insert table – not accessible from PDF] Table 1 Effects of nanomaterials on microbial species (from Klaine et al. 2008, Wiesner 2006, and references therein)
Generalised microbial effects reported have been: disruption of membrane/membrane potential, production of ROS, oxidation/damage to proteins, interference with electron transport/respiration, potential DNA damage, with in general more serious effects observed on Gram-positive species (Klaine et al. 2008 and references therein).
Few studies on the effects of nanomaterials on soil organisms have been published to date. One of the first studied the effects of aluminium oxide nanoparticles on the emergence and growth of plants (Yang and Watts 2005). The authors observed clear effects but it was later debated whether these effects were due to the nanoparticle form of aluminium or to a soluble fraction of aluminium ions (Murashov 2006). Regardless of the key cause, it is clear that a negative effect was observed that resulted from the exposure to aluminium in a nanoparticulate form.
Recently, Cañas et al. (2008) studied the effects of single walled carbon nanotubes on the root elongation of crop species. A few minor effects were detected on some species after rather short exposure times of 24 and 48 hours. However no uptake of single walled carbon nanotubes was observed. Lee et al. (2008), using a plant agar test system, indicated a toxic effect of copper nanoparticles as demonstrated by a reduced growth of seedlings of mung bean (Phaseolus radiatus) and wheat (Triticum aestivum). Using transmission electron microscopy and energy dispersive spectroscopy, the authors observed the accumulation of copper particles in the cells.
Reduced enzyme activities for catalase (CAT) and glutathione-S-transferase (GST) were observed after ingestion of titanium dioxide nanoparticles (anatase) by terrestrial isopods (Porcellio scaber) (Jemec et al. 2008). However the overall endpoints like survival and growth were not affected. Scott-Fordsmand et al. (2008) detected effects on the reproduction of earthworms (Eisenia veneta) when the worms were exposed to double- walled carbon nanotubes in food.
Studies focussing on the hazard of nanomaterials to a range of aquatic species were reviewed by SCENIHR (2006). Further developments took place through 2008 with new data and knowledge generated. Most of the recent studies have indicated that key aspects of aggregation may result in exposures not reflecting the highest doses. Issues considering solvents and the role of NOM, as a stabiliser of nanomaterials, have also been highlighted (see above). Nevertheless, it is important to note that increased dispersion has not led to increased bioavailability and hazard in all studies. Results depend on nanomaterial composition (Franklin et al. 2007, Navarro et al. 2008a). Exposure to raw carbon nanotubes resulted in the viability of Daphnia magna being reduced, but this was not found in stable dispersions when compared with functionalised and stabilised forms of carbon nanotubes (Kennedy et al. 2008). In contrast, Kang et al. (2008) indicated that some methods used to stabilise the dispersion of CNTs (e.g. functionalisation) resulted in increased toxicity for the bacterial systems they were studying.
Recent results have indicated the physical interference (e.g. movement hindrance, clogging) of nanomaterials with biota (e.g. Nielsen et al. 2008). Sublethal effects observed included lipid peroxidation, altered haematology, changes in behaviour with some effects being observed at different stages in the life cycle (Zhu et al. 2007).
Interactions of nanomaterials with other pollutants
It has been suggested that nanomaterials may interact with contaminants which may result in toxic effects on biota (Cheng et al. 2007, Xia et al. 2004). Baun et al. (2008) indicated the potential of nanomaterials to enhance the toxic effects of organic contaminants.
Methods of assessment in vitro
Methods of assessment in vitro have mirrored work in the area of mammalian toxicology. For example, several approaches used to study oxidative assessment have been used, and more slowly work is taking place in the area of genomics and proteomics. Other methods, borrowed from mammalian work, such as assessment of effects of nanomaterials on fish hepatocyte function evaluated by enzyme lactate dehydrogenase (LDH) as a marker for cell toxicity (Bopp and Lettieri 2008) or assessment of lysosome stability in mollusc hemocyte cells, may also be used (Castro et al. 2004).
Linking effects at different levels of organisation is important in the assessment of potential long term effects, as well as to improve knowledge on how some toxic effects may occur. In vivo exposures, followed by assessment of specific endpoints at organ and organelle levels, as well as biochemical endpoints, provide a comprehensive approach to assessment of effects. In vitro studies, such as the ones described above, allow a focused assessment of mechanistic effects at a specific level of organisation.
Methods of assessment in vivo
There is still much debate in the literature (Fernandes et al. 2007, Handy et al. 2008b, Klaine et al. 2008) regarding what may be considered optimal approaches for exposures. Exposure media, mixing or suspension of materials within the media and consideration of realistic exposures, have all been a particular focus of attention. In this context, an important point of consideration has been the characterisation of the nanomaterials in the exposure studies. This has been particularly debated in studies of sedimentary systems. Mixing of nanomaterials with sediments/soils, as well as characterisation over time, are still areas at a very early stage of development. In this context, consideration of detection, in a background of natural abundance of specific types of materials such as carbon-based products, zinc and silicon, is an area that is still currently advancing in technical terms.
A wide range of methods have been used to assess the hazard of nanomaterials on environmental species. The approaches were chosen according to the species studied. Laboratory studies have focussed on the effects of a range of nanomaterials on standard species used in ecotoxicology. Most have focussed on aquatic species including: primary producers (mainly Pseudokirchneriella subcapitata (Franklin et al. 2007, Van Hoecke et al. 2008) and Desmodesmus subspicatus (Hund-Rinke and Simon 2006); invertebrates, mainly Daphnia species but also other crustacean (Fernandes et al. 2007, Hund-Rinke and Simon 2006, Lovern and Klaper 2006, Rosenkranz et al. 2009), and fish (such as rainbow trout Oncorhynchus mykiss, zebra fish Danio rerio, largemouth bass Micropterus salmoides, fathead minnow Pimephales promelas and Japanese medaka Oryzias latipes;( Federici et al. 2007, Griffit et al. 2007, Lee et al. 2007, Oberdörster 2004, Smith et al, 2007, Warheit et al. 2007, Zhang et al. 2007). The toxicity to various microbial organisms has also been studied (e.g. Lyon et al. 2006, Lyon and Alvarez 2008, Sondi and Salopek-Sondi 2004). In some of these studies, toxic effects were observed.
The integration of endpoints such as mortality, growth, feeding and reproduction, are widely used in ecotoxicology. In addition, specific biomarkers, such as means of assessing oxidative stress (in a specific organ, or whole body; e.g. lipid peroxidation), genetic damage, CYT P450 levels, gene expression, damage to specific cell organelles (e.g. mitochondria or nucleus ) are all widely used in the assessment of effects of nanomaterials. Cytological responses such as celluar apoptosis and necrosis have also been used. Although the methods employed in these studies tend to be standard methods used routinely in ecotoxicology studies, modifications have been implemented to address the specific particle issue, or to address the effects of interference of the materials with reading of results.
Less research has taken place using soil or sedimentary species but work is now progressing at a steady pace in these systems. The procedures that were adopted have in general followed OECD guidelines, particularly when using standard species, but also when other species were used. Depending on the aims of the study, different durations of exposure have been used.
The number of scientific studies assessing the environmental effects of nanomaterials has increased dramatically in the period 2007-2008. The main focus is still on micro- organisms and invertebrates, followed closely by studies on fish species. Still very much lacking are studies on soil systems and terrestrial species in general, including primary producers. There is also a general paucity of studies on marine species. This is not surprising given the complexity associated with dispersing and suspending nanomaterials in exposure media. Nevertheless, published results to date indicate clearly the potential for hazardous effects, at lethal and sublethal levels, including behaviour, reproduction, growth and development, ROS production, induction of inflammatory responses and cytotoxic effects. In addition, a small number of studies have indicated the potential for transfer to embryos, accumulation and potential food chain transfer.
Nevertheless, the exposure levels organisms may endure in their natural environments and how the results in the laboratory can be extrapolated to assess hazard in the field is less clear. Information on environmental loads is at present lacking. One important aspect in this context is the understanding of any interactions of nanomaterials with micro-organisms in sewage treatment plants, and the consequent effects on the treatment process.
A few key issues need to be brought out when assessing critically the results obtained to date on the environmental hazard of nanomaterials in order to focus on what is important and optimise the approach and design of future studies.
A first issue, which has already been widely discussed, concerns the protocols used in laboratory exposures and a related link to the current lack of standardised protocols. The use of mechanical or chemical means to suspend nanomaterials may lead to changes in the physical-chemical properties of the test material. It is unclear what the extent of these may be and how they may impact any effects observed. Arguably, dispersants/surfactants/solvents may need to be used in certain situations; however, it is important that they must not add to the toxicity of studied materials. It is suggested that results of studies where THF was used should be treated with caution, as at least in one study, the observed toxicity was due to traces of THF (Zhu et al. 2006). The same caution may apply to other dispersants for which there is lack of knowledge regarding their interaction with the test material (e.g. SDS). Further work with humic and fulvic acids, as well as widely used detergents (which are likely to be encountered in the environment) should be undertaken.
Related to this topic is the use in hazard assessment of ready-made (off-the-shelf) suspensions of nanomaterials. It is unclear what the interactions might be of the used preparation dispersants on the properties (and thus behaviour) of the test material (as described above). Thus any reported effects might not be comparable with effects observed on exposures of the same species to the same component material but which is in a different form (i.e. solid and suspended nanomaterials in the laboratory vs nanomaterials obtained as a suspension).
Therefore, standardisation of protocols, as possible, is desirable for the comparability of studies as well as reliability of results, and the derivation of information to risk assessment and risk management.
Regarding experimental design and approach, characterisation of exposures, via appropriate method(s) should be carried out and chemical analyses undertaken, as possible. The assessment of the solubility of the nanomaterials being studied is very important in this context so that any observed effects can be attributed to the different fractions. This is particularly important in the case of certain metal nanomaterials, as well as in the case of CNTs and quantum dots.
Some studies have highlighted the importance of assessing contamination of the nanomaterials being studied. This should be undertaken for similar reasons. Another important point is the comparison of effects between nano and equivalent, larger, material. This has not been consistently incorporated in the published studies and would also allow the correct attribution of effects.
There is a lack of information regarding the fate and form of the test nanomaterials within biological systems following in-vivo exposures. It is unclear what particular form (e.g. soluble or particulate) is preferentially taken up into tissues and cells. It is likely that this would depend on the material composition; nevertheless these studies are not routinely carried out.
Studies should be conducted on a range of guilds and endpoints, with fate within the body and tissues assessed and depuration quantified, as possible. Micro/mesocosms studies should be undertaken. Furthermore, dietary studies, the role of nanomaterials’ coatings in uptake and translocation within the body, should be conducted, as well as the assessment of the role, if any, of their interaction with other environmental contaminants.
In this context it is crucial to ascertain the fate of nanomaterials in the environment so that their availability for environmental exposure can be assessed. Environmental fate and load assessment of nanomaterials must, therefore, be undertaken. The use of the current approach to the derivation of Kow in the assessment of environmental fate is unlikely to be beneficial to risk assessment. Nevertheless, the derivation of alternative approaches may be useful and may allow the development of appropriate predictive modelling. Finally, further information on the degradability (bio and abiotic) of nanomaterials should be derived.