ABSTRACT
The most important processes affecting the fate of nanoparticles in aquatic systems are agglomeration and aggregation, dissolution, redox reactions and transformation into new solid phases (Handy et al., 2008b; Klaine et al., 2008; Navarro et al., 2008a; Nowack and Bucheli, 2007). Agglomeration or aggregation of nanoparticles lead to larger particles, which may be removed from the water column and transported to the sediments. Nanoparticles are only weakly bound by agglomeration, whereas chemical bonds between particles are formed upon aggregation (Jiang et al., 2009). Bioavailability of NP and their dissolution behavior may also be different for larger agglomerated particles. Agglomeration is determined by the medium composition, mostly ionic strength, pH and the concentration of natural organic matter. To assess the effects of these various factors, the behavior of different types of NP, such as elemental MeNP (e.g., Ag(0), Au(0), Cu(0)), metal oxide NP (e.g., TiO2, Al2O3, CeO2, ZnO), quantum dots and carbon-based NP (carbon nanotubes and fullerenes) needs to be considered. Elemental MeNP are maintained in suspension by their surface coating, which may be negatively or positively charged, or may be stabilized based on steric effects. Examples are AgNP suspensions which are stabilized by carbonate, citrate, or charged polymers. Other polymeric coatings stabilize AgNP by steric effects (e.g., polyvinylpyrrolidone PVP). Effects of pH and ionic strength on the agglomeration of AgNP suspensions have been examined for various coatings. Agglomeration has been observed in particular at low pH, when the negative surface charge of the coating becomes neutralized, and at high ionic strength (Badawy et al., 2010; Elzey and Grassian, 2010; Gao et al., 2009; Huynh and Chen, 2011; Piccapietra et al., 2012). Humic acids are expected to influence the stability of the suspension by interactions with the surfaces and by steric effects. In natural waters, the water composition, i.e. pH, ionic strength, concentration of divalent cations and of fulvic and humic acids, will determine the colloidal stability of AgNP suspensions (Piccapietra et al., 2012). These factors have also to be considered in media used for biological experiments with AgNP. Metal oxide NP such as TiO2, Al2O3, and CeO2 behave as oxides with a strongly pH-dependent surface charge, due to the protonation and deprotonation of the surface OH-groups, if they are not coated
by organic compounds or functionalized. Agglomeration of these metal oxides is expected at neutral surface charge in the pH range of the zero point of charge, which is a distinct characteristic of each of the oxides. For example, CeO2 NP have a zero point of charge in the neutral pH-range 7-8 (De Faria and Trasatti, 1994; Nabavi et al., 1993). Agglomeration of CeO2 NP has been observed around pH 7, used in media for growth and toxicity testing of aquatic organisms (Rogers et al., 2009; Van Hoecke et al., 2009). TiO2 behaves in a similar way, with a zero point of charge in the neutral pH-range. Stabilization of TiO2 NP by humic acids has been demonstrated (Domingos et al., 2009). Dissolution reactions of NP are expected to play an important role in their toxicity, as toxic metal ions may be released from the NP, in particular Ag+ from AgNP, Cu2+ from elemental Cu or from CuO NP, Zn2+ from ZnO, Cd2+ from quantum dots (e.g., CdSe, CdTe). Toxicity of these ions to aquatic organisms is well known and strongly depends on speciation of these ions in solution (Campbell et al., 2002; HiriartBaer et al., 2006). Dissolution reactions of AgNP by oxidation of Ag(0) lead to the release of free Ag+ ions, which may be determining the toxic effects (Kittler et al., 2010; Navarro et al., 2008b). The rate and extent of dissolution is expected to be dependent on the solution conditions, such as pH, presence of oxygen and effects of ligands (Galloway et al., 2010). Increased dissolution at low pH, as well as the role of O2and H2O2 as oxidants was shown (Galloway et al., 2010). Increased dissolution of AgNP in the presence of algae has been postulated to influence their toxicity (Navarro et al., 2008b). Various coatings on AgNP determine the extent and kinetics of the dissolution reactions (Odzak et al., 2013). The release of Ag+-ions from AgNP in the presence of chloride or sulfide may lead to the formation of solid silver chloride or silver sulfide at the surface of the NP, as demonstrated upon aging of AgNP with chloride (Badawy et al., 2010), and for the formation of silver sulfide upon exposure of AgNP in a pilot wastewater treatment plant (Kaegi et al., 2011). Sulfide may form the very insoluble silver sulfide solid phase, but also dissolved silver complexes (e.g., AgHS0, Ag(HS)2-) and small AgxSy aggregates (Bell and Kramer, 1999; Luther and Rickard, 2005). The solubility of metal oxide NP varies over a wide range, depending on their composition and on the crystal structures. For example, Ce(IV)O2 has very low solubility at neutral pH, and soluble
Ce(IV) species occur only at pH < 4 (Hayes et al., 2002; Yu and O’Keefe, 2006). Free Ce4+ ions only occur in very acidic solutions. However, upon reduction to Ce(III), the Ce solubility is strongly increased (Hayes et al., 2002). In a similar way, the solubility of TiO2is very low at neutral pH (Knauss et al., 2001; Ziemniak et al., 1993). In contrast to these poorly soluble oxides, ZnO is readily soluble, and effects of ZnO have been shown to mostly depend on released Zn2+ (Franklin et al., 2007). Quantum dots, such as CdSe, are usually coated with a zinc sulfide shell and with organic polymers. In spite of this coating, substantial release of Cd2+ from CdSe quantum dots has been observed in toxicity experiments (King-Heiden et al., 2009; Klaine et al., 2008). 5.3 Fate in Model EcosystemsEcotoxicity studies of appropriate ecological relevance include those involving communities where various species occur within a complex network and interact with each other and the abiotic environment to provide ecosystem processes. Considering the entrance of NP in the aquatic environment, most urgent questions are (1) where do NP partition in aquatic systems and in which physico-chemical state; (2) are the NP bioavailable; (3) do NP affect the structure and function of biotic communities, and (4) do community changes entail changes in ecological processes? Among a dozen of published studies that can be considered as ecologically relevant, all have considered MeNP and have examined their distribution in diverse model systems enclosed in microcosm or mesocosms. Examination of the fate of gold nanorods in an estuarine ecosystem containing sediments, microbial biofilms, primary producers, filter feeders, grazers and omnivores identified the filter feeders as the most effective sink for NP, followed closely by biofilms (Ferry et al., 2009). While the question whether and how NP are taken up by aquatic organisms remains to be elucidated in detail (see Section 5.4), trophic transfer of NP from algae to daphnids (Bouldin et al., 2008) and to clams (Croteau et al., 2011), as well as from daphnids to fish (Zhu et al., 2010) was shown for metal-based quantum dots, ZnO and TiO2 NP, respectively, evidencing diet as one pathway through which daphnids, clams and fish accumulate the metals. In another study, quantum dots were transferred to higher trophic organisms (rotifers) through dietary uptake of ciliated
protozoans (Holbrook et al., 2008). However, these studies did not resolve whether intact particles or the dissolved metals have been transferred. Similar as for dissolved metals, biomagnification, a process through which concentrations of chemicals increase in organisms from a lower to a higher trophic level within the same food web, was not observed to occur in these studies. However, even in absence of biomagnification, aquatic organisms can accumulate large amounts of metals and become a significant dietary source of metal to their predators (Reinfelder et al., 1998). One study has reported on the trophic transfer of intact particles in a system including bacteria previously exposed to quantum dots and protozoa that fed on the bacteria (Werlin et al., 2011). In this system, quantum dots were shown to be biomagnified in the protozoa that also displayed inhibited digestion as shown by the presence of intact bacteria containing NP in the food vacuoles of the protozoa. Trophic transfer and biomagnification of NP was also demonstrated in a terrestrial system composed of plants dosed with AuNP that were fed to hornworm larvae (Judy et al., 2011). The observation that NP may be biomagnified not only confirms the importance of dietary exposure for the transfer of NP, but also the fact that NP might accumulate up the food chain to concentrations eventually resulting in toxicity for predators. Clearly, more systematic work is necessary to understand the apparent contradictory information on NP biomagnification.
