Risks of Engineered Nanoparticles for the Environment

Risks of Engineered Nanoparticles for the Environment Abstract: The objectives of this article are to: (1) investigate the current state of knowledge of the risks of engineered nanoparticles for the environment and human health, (2) estimate whether this knowledge is sufficient to facilitate their comprehensive and effective risk assessment and (3) provide recommendations on future research in the field of risk assessment of nanomaterials. In order to meet the objectives, the relevance of each of the four steps of the risk assessment methodology (i.e., hazard identification, dose-response assessment, exposure assessment and risk characterization) was evaluated in the context of the current state of knowledge of the risks of nanomaterials, limitations were identified and recommendations were given on how to overcome them. Keywords: engineered nanoparticles; risk assessment; hazard identification; dose-response assessment; exposure assessment; risk characterization; environmental sustainability; human health Introduction Background In contrast to the small size of the nanoparticles, the scale of their application is tremendous. Nanotechnology influences virtually all industrial and public sectors, including healthcare, agriculture, transport, energy, materials, information and communication technologies. Both the potential benefits and the risks, associated with the application of engineered nanoparticles (ENPs) have been largely debated in recent years. In contrast to the dominating optimistic projections that nanotechnology will bring significant technological development and well-being to society, it is considered that exposure to certain ENPs may cause environmental problems and/or do harm to human health. Since the early discussions about the risks of ENPs, the chemical risk assessment (CRA) has been put forward as the most relevant approach to understand, evaluate and quantify these risks. Currently, a variety of methodologies are being internationally discussed and evaluated with great vengeance with the idea that, in the near future, it will be possible to perform complete and scientifically sound risk assessment of ENPs. Objectives The objectives of this article are to: Investigate the current state of knowledge of the risks of ENPs for the environment and human health Estimate whether this knowledge is sufficient to facilitate comprehensive and effective risk assessment of ENPs Provide recommendations on future research in the field of risk assessment of ENPs Methodology This article is based on an extensive review of literature published in the period: January 1992- September 2009. The selected literature consisted mainly of scientific publications, but also books, information from conferences and patent data were used. Nanotechnology and Its Applications Nanotechnology and Nanoparticles Nanotechnology is a field of applied science and technology, dealing with the organization and control of matter on the nano-scale (i.e., between 1 and 100 nm) and the manufacturing of products and devices with dimensions, lying within this size range. A nanometer (nm), from the Greek nanos for dwarf, equals one billionth of a meter. Nanomaterials are all materials with sizes on the nano-scale in at least one of their dimensions [1], while nanoparticles are materials, nano-sized in at least two dimensions [2]. The nomenclature nanoparticles encompasses particles as well as fibrous materials and tubes, but it excludes materials, such as coatings, films and multilayers. Two types of nanoparticles (NPs) can be distinguished: (1) naturally occurring NPs (e.g., produced naturally in volcanoes, forest fires or as combustion by-products) and (2) engineered nanoparticles (ENPs), deliberately developed to be used in application (e.g., carbon black, fumed silica, titanium dioxide (TiO2), iron oxide (FOx), quantum dots (QDs), fullerenes, carbon nanotubes (CNTs), dendrimers). Naturally occurring NPs do NOT fall in the scope of this article. The paper encompasses only ENPs. The main reasons why materials, built of ENPs, have different optical, electrical, magnetic, chemical and mechanical properties from their bulk counterparts are that in this size-range quantum effects start to predominate and the surface-area-to-volume ratio (sa/vol) becomes very large [1]. The sa/vol of most materials increases gradually as their particles become smaller, which results in increased adsorption of the surrounding atoms and changes their properties and behavior. Once particles become small enough, they start to obey the quantum mechanical laws. Materials reduced to the nano-scale can suddenly show very different properties, compared to what they exhibit on the macro-scale, which enables unique applications. For example, opaque substances become transparent (copper); stable materials become combustible (aluminum); inert materials become catalysts (platinum); insulators become conductors (silicon); solids turn into liquids at room temperature (gold) [3]. Areas of Application Today, nanotechnology is available on the market for great variety of applications. Some examples are: cosmetics and sunscreens, water filtrations, glare filters, ink, stain-resistant clothing, more durable tennis balls, more lightweight tennis rackets, dressings for burns or injuries. [4]. Defining Hazard and Risk The term hazard has many definitions. This paper uses the definition of the United States Environmental Protection Agency (EPA) which defines hazard as the inherent toxicity of a compound [5]. According to this definition, if a chemical substance has the property of being toxic, it is therefore hazardous. Any exposure to a hazardous substance may lead to adverse health effects in individuals or even death. EPA defines risk with respect to the above definition of hazard as a measure of the probability that damage to life, health, property, and/or the environment will occur as a result of a given hazard [5]. According to this definition, if the probability of an exposure to a hazardous material is high and the consequences for the health or environment are significant, then the risk is considered to be high. It is important to consider both the frequency of the event and the degree of the hazard to estimate risk [2]. Usually two categories of risk are distinguished in literature: known risks and potential risks. When the relation between a cause and an effect is established, we talk of known risks. The responsibility for such risks can generally be attributed. When the causal relationship is established, prevention is possible. When the relationship between a cause and damage is not well known, we talk of potential risks. In case of potential risks, it is unclear whether there is a danger, how significant the damage can be or what is the probability of its occurrence [2, after 6]. This situation is characterized by a state of suspicion (not awareness) and it is generally admitted that a precautionary approach can be applied in order to prevent potential damage [2, after 6]. The risks of ENPs for the environment and human health fail in the second category: potential risks. It is very important to assess the risks of hazardous agents. The likelihood that a hazardous substance will cause harm (the risk) is the determinant of how cautious one should be and what preventative or precautionary measures should be taken. Risk Assessment of ENPs Since the early debates about the potential hazards of ENPs, the risk assessment of chemicals (CRA) has been put forward as the most relevant approach to understand and quantify the related risks [7]. CRA is a process, in which scientific and regulatory principles are applied in a systematic fashion in order to describe the hazard, associated with the environmental and/or human exposure to chemical substances. It is defined as a process, intended to calculate or estimate the risk to a given target organism, system or (sub)population, including the identification of attendant uncertainties, following exposure to a particular agent, taking into account the inherent characteristics of the agent of concern, as well as the characteristics of the specific target system [8]. The CRA is a four-step process, consisting of: (1) hazard identification, (2) dose- response assessment, (3) exposure assessment and (4) risk characterization. Its main outcome is a statement of the probability that whe n humans or other environmental receptors (e.g., plants, animals) are exposed to a chemical agent, they will be harmed and to what degree. The CRA methodology is internationally recognized and employed by major actors, such as the World Health Organization (WHO) and the Organization for Economic Co-operation and Development (OECD), as well as by several European and U.S. agencies [9]. It is considered a valuable tool, very important for the regulation of chemicals. CRA is also a fundamental ingredient of the new European Union (EU) chemical regulation policy, known as Registration, Evaluation and Authorization of Chemicals (REACH). In order to achieve the objectives of this study, the current state of knowledge of the risks of ENPs for the environment and human health were summarized and evaluated in relation to each of the four elements of the CRA framework, as more important scientific findings were highlighted and limitations were identified and discussed. Hazard Identification Hazard identification (HI) is defined as the identification of the adverse effects, which a substance has an inherent capacity to cause [10, after 11]. Until recently, much of the discussion about the environmental and health risks of ENPs was considered to be rather speculative than realistic. In the last few years, however, a number of experimental studies found that exposure to certain ENPs can lead to adverse health effects in living organisms. In 2007, Hansen et al. identified 428 studies reporting on toxicity of ENPs [12]. In these studies, adverse health effects of 965 tested ENPs of various chemical compositions were observed [12]. Current State of Knowledge The following sections shortly describe some of the most important scientific findings, relevant for HI of ENPs. Their purpose is to summarize the current state of knowledge of the hazards of ENPs, based on experimental studies. For simplification, the studies are divided into two categories in vivo and in vitro studies. In Vivo Studies Carbon Nanotubes (CNTs) A study, performed by Lam et al. [13], demonstrated that single- walled carbon nanotubes (SWCNTs) are able to cause dose-dependent effects of interstitial inflammation and lesions in mice and rats (0- 0.5 mg kg-1 for 7 to 90 days). Warheit et al. [14] observed pulmonary grandulomas in rats after exposure to SWCNT soot (1 and 5 mg kg-1 for 24 hours to 3 months). In contrast to Lam et al. [13], however, the effects, observed by Warheit et al. [14] were not dependent on dose. Smith et al. [15] tested the ecotoxicity of SWCNTs, dissolved in sodium dodecyl sulphate (SDS) and sonication on juvenile rainbow trout (0.1, 0.25 and 0.5 mg l-1 for 24 hours to 10 days) and they observed a dose-dependent rise in ventilation rate, gill pathologies (oedema, altered mucocytes, hyperplasia), and mucus secretion with SWCNT precipitation on the gill mucus. They also observed a significant dose- dependent decrease in thiobarbituric acid reactive substances (TBARS), especially in the gill, brain and liver , which is an indication of oxidative stress. Multi- walled carbon nanotubes (MWCNTs) were shown by Carrero-Sanchez et al. [16], to exhibit acute toxicity in rats with LD90 of 5 mg kg-1. Long MWCNTs were shown by Poland et al. [17] to cause significant inflammation and tissue damage in mice, while shorter MWCNTs caused less inflammation, which suggests that CNT toxicity is influenced by the particle morphology. In addition, they concluded that water-soluble components of MWCNT do not produce strong inflammatory effects in mice. C60 Fullerenes Most studies on the toxicological effects of C60 fullerenes suggest that these materials tend to induce oxidative stress in living organisms [18-21]. Lai et al. [18] observed a significant increase in lipid peroxidation (LP) products (a sign of oxidative stress) after intravenous administration of 1 mg kg-1 C60 (OH)18 in male mongrel dogs. Oberdà ¶rster [19, 20] studied the effects of C60 fullerenes in the brain of juvenile largemouth bass and observed high LP levels (0.5 and 1 ppm for 48 h). Elevated LP was also observed by Zhu et al. [21] in the brain and gills of daphnia magna after exposure to hydroxylated C60 fullerenes (C60 (OH)24) and tetrahydrofuran (THF)- dissolved C60, as it was shown that THF did not contribute to the effect. Sayes et al. [22] detected an increase in the numbers of bronchoalveolar lavage (BAL)-recovered neutrophils (i.e., white blood cells) after intratracheal instillation of C60 and C60 (OH)24 in rats, 1 day after the exposure. They also observed a signi ficant increase in LP values 1 week after the exposure. Accute effects of functionalized C60 were also reported . Zhu et al. [21] estimated LC100 in fathead minnow after exposure to 0.5 ppm of THF-dissolved C60 for 6-18 hours. Chen et al. [23] observed a LD50 of 600 mg kg-1 polyalkylsulfonated C60 in female rats after intraperitoneal administration (0- 2500 mg kg-1 for up to 2 weeks). Oberdà ¶rster [24] tested uncoated, water soluble, colloidal C60 fullerenes and estimated a Daphnid 48-hour LC50 of 800 ppb. Metal and Metal Oxide ENPs Li et al. [25] found that metal ENPs induce more severe lung toxicity in mice than bulk particles from the same materials. Gordon et al. [26] tested the effects on humans of exposure to zinc (Zn) ENPs. After 2 hours of exposure to 5 mg m-3 of Zn ENPs, the exposed individuals started feeling sore throat, chest tightness, headache, fever and chills. Beckett et al. [27] repeated that test in three trials, 2 hours each, but at lower concentration (i.e., 500 ?g m-3), and found no indication of adverse effects. The latter two studies suggest that Zn ENPs toxicity is concentration- dependent and the most probable uptake path is through the respiratory system. A study of Sayes et al. [22] concluded that environmental exposure to Zn ENPs causes pulmonary (lung) inflammatory response in mice. Wang et al. [28] found that Zn ENPs can cause severe symptoms of lethargy, anorexia, vomiting, diarrhea, loss of body weight and even death in mice when gastrointestinally administered, whereas they obser ved limited effect for micro- scale Zn at equal concentrations. Yang and Watts [29] tested the effect of Aluminium (Al) ENPs on the relative root growth (RRG) in Zea mays (corn), Glycine max (soybean), Brassica oleracea (cabbage), and Daucus carota (carrot). The study found that the ENPs significantly inhibited the growth of the plants after administration of 2 mg ml-1 for 24 h. Oberdà ¶rster [30] and Oberdà ¶rster et al. [31] observed that smaller TiO2 ENPs tend to cause more severe pulmonary damage in mice than larger particles. In addition, Warheit et al., [32] found that smaller silicon dioxide (SiO2) particles cause stronger lung inflammation in rats than larger ones. Wang et al., [33] noticed that the smaller the TiO2 particle size is, the greater the concentration in the liver of mice is. Bourrinet et al. [34] reported hypoactivity, ataxia, emesis, exophthalmos, salivation, lacrimation, discolored and mucoid feces, injected sclera, and yellow eyes in dogs after single-dose intravenous bolus administration of 20 and 200 mg kg-1 FeO ENPs and a significant increase in fetal skeletal malformations in rats and rabbits. In Vitro Studies Carbon Nanotubes (CNTs) A number of cytotoxicity studies with SWCNTs were reported in the literature. Shvedova et al. [35] observed oxidative stress and cellular toxicity in human epidermal keratinocytes, after 2 to 18 hours exposure to unrefined (iron containing) SWCNTs in concentrations, ranging from 0.6 to 0.24 mg ml-1. Cui et al. [36] observed dose- and time- dependent inhibition of cell proliferation and a decrease in cell adhesive ability in human embryo kidney cells after exposure to SWCNTs in concentrations between 0.8 and 200 ?g ml-1. Sayes et al. [37] found that the surface functionalization of SWCNTs plays an important role in their cytotoxicity towards human dermal fibroblasts. Bottini et al. [38] noticed that MWCNTs were more cytotoxic when oxized towards Jurkat T leukemia cells, whereas Monteriro-Riviere et al. [39] observed a decrease of the viability of human osteoblastic lines and human epidermal keratinocytes after exposures to 0.1, 0.2, and 0.4 mg ml-1 of MWCNTs for 1 to 48 hours. Kang et al. [40] compared the cytotoxicity of commercially obtained MWCNTs in bacterial systems before and after physicochemical modification and they observed highest toxicity when the nanotubes were uncapped, debundled, short, and dispersed in solution. Kang et al. [40] concluded that there is need for careful documentation of the physical and chemical characteristics of CNTs, when reporting their toxicity. C60 Fullerenes Adelman et al. [41] observed a reduction of the viability of bovine alveolar macrophages after exposure to sonicated C60 and increased levels of cytokine mediators of inflammation (i.e., IL-6, IL-8 and TNF), while Porter et al. [42] found that C60 and raw soot were not toxic towards bovine- and human alveolar macrophages. The reason behind the discrepancy between the results of Adelman et al. and Porter et al. can be attributed to the fact that they used very different methods. Porter et al. used transmission electron microscopy (TEM) to image the distributions of the fullerenes within the macrophages, while Adelman et al. used a viability assay, based on metabolic activity as primary parameter. Studies on the effects of ENPs on alveolar macrophages are very important because the alveolar macrophages are the first line of cellular defense against respiratory pathogens [11, after 43]. Yamawaki Iwai [44] observed dose-dependent cytotoxicity of C60 (OH)24 (1- 100  µg ml-1 for 24 hours), resulting in decreased cell density and lactate dehydrogenase (LDH) release in human umbilical vein endothelial cells cavity (a sign of increase in non-viable cell numbers). Rouse et al. [45] observed a dose-dependent decrease in the viability of human epidermeal keratinocytes after exposure to C60- phenylalanine, as no contribution to the effect was attributed to the phenylalanine groups. Quantum Dots (QDs) The toxicity of QDs was found to be influenced by several factors: (1) composition, (2) size, (3) surface charge and (4) coating of the QDs [7, 46- 48]. Jaiswal et al. [46] found that CdSe/ZnS QDs (i.e., CdSe QDs in a zinc sulfide (ZnS) matrix), coated with dihydrolipoic acid (DHLA) had no effect on mammalian cells, while Hoshino et al. [47] reported adverse effects on mouse lymphocytes after exposure to CdSe/ZnS QDs, coated with albumin. In addition, Lovrà ­c et al. [48] observed that smaller (2.2  ± 0.1 nm), positively charged QDs exhibit stronger cytotoxicity than larger (5.2  ±0.1 nm), equally charged QDs under the same conditions. It was also found that the cytotoxicity of QDs is influenced by the exposure to light and by temperature [49, 50]. Green and Howman [49] observed 56% damaged DNA after exposure to CdSe/ZnS together with UV light versus only 29% after exposure to CdSe/Zn in the absence of UV light. Chang et al. [50] found that CdSe/CdS (i.e., CdSe QDs in a cadmium sulfide (CdS) matrix) were toxic to cancer cells at 37  ºC, but at 4  ºC they were not toxic at all. Metal and Metal Oxide ENPs Sayes et al. [51] found that anatase TiO2 ENPs are able to kill human dermal fibroblast (HDF) cells at LC50 of 3.6 ?g ml-1, while Wang et al. [52] observed decrease in the viability of human lymphoblastoid cells due to exposure to TiO2 ENPs (0-130 ?g ml-1 for 6-48 h). Chen Mikecz [53] found that SiO2 ENPs do significantly inhibit replication and transcription in human epithelial HEp-2 cells (25 ?g ml-1 for 24 h). Muller et al. [54] observed that Fe3O4 ENPs, coated with dextran, decrease the viability of human monocyte macrophages. Alt et al. [55] found that nano-particulate silver (Ag) is an effective bactericide against S. epidermidis, while Baker et al. [56] noticed that it effectively kills E. coli bacteria too. Sayes et al. [57] observed an increase in the production of LDH levels (an indicator of inflammation) in immortalized rat lung epithelial cells after 1 hour exposure to Zn ENPs at 520 ?g cm-2. Limitations to Hazard Identification of ENPs It is very important to note that the vast majority of the reviewed studies demonstrate some degree of hazardous effects on the tested organisms. Toxicity has been reported for many ENPs, as shown in the previous sections, but for most of them further investigation and confirmation are needed before hazard can be identified. A lot of studies, relevant for HI, have been carried out with different ENPs, but most of them were obviously not meant to facilitate risk assessment; they use non- standardized tests, differing greatly from each other in regard to endpoints, tested species, methods of administration, dose ranges and exposure periods [7]. The lack of standardized testing results in non-reproducible results and makes the univocal HI of ENPs impossible. Another significant drawback for the HI of ENPs is the serious lack of characterization data, which makes it difficult to identify which physical and/or chemical characteristics (or combinations of characteristics) determine the hazards, documented in the (eco)toxicological studies [12, 58, 59]. Dose-Response Assessment Dose- response assessment (DRA) is defined as an estimation of the relationship between dose, or level of exposure to a substance, and the incidence and severity of an effect [10, after 11]. It is the process of characterizing the relationship between the dose of an agent, administered to or received by an individual, and the consequent adverse health effects. The Concept of Dose In toxicological studies a dose is the quantity of anything that may be received by or administered to an organism. The dose is normally measured in mass units (i.e., ?g, mg, g), as higher doses of the same compounds are expected to cause more severe adverse effects. DRA studies with ENPs, however, suggest that the toxicity of some ENPs is not mass-dependent, but influenced by other physico-chemical characteristics (e.g., surface area, chemical composition, particle morphology) [7, after 60]. Oberdà ¶rster et al. [61] and Stoeger et al. [62, 63] found that the toxicity of low-soluble ENPs was better described by their surface area than by their total mass [7, after 61, 62, 63]. Wittmaack [64, 65] suggested the number of particles as the most appropriate dose metrics, while Warheit et al. [66, 67] found that toxicity of some ENPs was associated with the number of their surface functional groups. Despite these findings, however, it is still largely unknown which properties influence the toxicity of most ENPs and this gap in knowledge is partly attributable to the fact that the tested ENPs are seldom well characterized. Characterization of ENPs Developing understanding about the physical and chemical properties of substances and materials is fundamental for their risk assessment [59]. Studying the standard properties (e.g., composition, structure, molecular weight, melting point, boiling point, vapor pressure, octanol-water partitioning coefficient, water solubility, activity, stability) is sufficient for the characterization of most chemical compounds. For ENPs, however, more profound investigation is needed and other properties, such as particle size distribution, sa/vol ratio, shape, electronic properties, surface characteristics, state of dispersion/agglomeration and conductivity need to be studied [5]. The high complexity and great diversity of ENPs, however, make their characterization very difficult [59]. As it can be inferred from the table above, most of the current research on the properties of ENPs is focused on the identification of metrics and associated methods for the measurement of ENPs and their properties. This type of research is fundamental in the sense that without reliable measurement methodology it would be impossible to develop good understanding of the physical and chemical properties of the ENPs. Only few comprehensive studies on the development of standard, well-characterised reference nanomaterials were published so far. To facilitate the appropriate interpretation of testing results, it is essential to select representative sets of ENPs, characterize them and share them among laboratories worldwide. Exposure Assessment Exposure assessment (EA) is defined as an estimation of the concentrations/doses to which human populations (i.e., workers, consumers and man exposed indirectly via the environment) or environmental compartments (aquatic environment, terrestrial environment and air) are or may be exposed. [10, after 11]. EA is a very important element in risk assessment of ENPs, since if no exposure to ENPs occur, it would be impossible that they cause any harm and there would be no risk at all. EA can be divided into three sub-areas: (1) occupational exposure assessment (OEA), (2) environmental exposure assessment (EEA) (including indirect human exposure from the environment) and (3) consumer exposure assessment (CEA). Environmental Exposure Assessment The environment may be exposed to ENPs during all stages of their life-cycles: raw material production, transport and storage, industrial use (incl. processing and/or trade), consumer use, waste disposal (incl. waste treatment, landfill and recovery) [11] (Figure 2.). A very important element of the EEA of ENPs is the study of their environmental fate. The fate of ENPs, released in the environment is determined by their mobility in the different media (i.e., soil, water, air), as well as by their potential to biodegrade or undergo chemical transformation. Environmental Fate of ENPs In order to determine the extent of environmental exposure to ENPs, it is necessary to understand their behavior in the environment. Until now, only a limited number of environmental fate studies with ENPs have been reported and the fundamental mechanisms behind their distribution are still not clearly understood (table 3). Fate of ENPs in Air The fate of ENPs in the air is determined by three main factors: (1) the duration of time particles remain airborne, (2) their interaction with other particles or molecules in the atmosphere and (3) the distance they are able to travel in the air [68]. The processes important to understand the dynamics of ENPs in the atmosphere are diffusion, agglomeration, wet and dry deposition and gravitational settling [68]. These processes are relatively well understood from studying the air-suspended ultrafine particles and that knowledge can be applied to ENPs as well [69]. In some cases, however, there can be considerable differences in behavior between ENPs and ultrafine particles, especially when the latter cannot agglomerate because they are coated [5]. With respect to the duration of time ENPs stay in the air, it is considered that they may follow the laws of gaseous diffusion [70]. The rate of diffusion is inversely proportional to the particle diameter and the rate of gravitational settling is proportional to it [70]. It is generally considered that particles in the nanoscale (d> 100nm) have shorter residence time in the air, compared to medium-sized particles (100nm> d> 2000nm), because they rapidly agglomerate into much larger particles and settle on the ground [71]. Here again ENPs with anti-agglomerate coatings make an exception and their residence time cannot be predicted [71]. It is considered that deposited ENPs are usually not likely to be re-suspended or re-aerosolized in the atmosphere [72, 70]. Many nano- sized particles are photoactive [72], but it is still unknown whether they are susceptible to photodegradation in the atmosphere. ENPs also show high absorption coefficients [69], and many of them can act as catalysts. However, no information is currently available on the interactions between ENPs and the chemicals they absorb, and how this interaction might influence atmospheric chemistry. Fate of ENPs in Water The fate of ENPs in water is determined by several factors: (1) aqueous solubility, (2) reactivity of the ENPs with the chemical environment and (3) their interaction with certain biological processes [5]. Because of their lower mass, ENPs generally settle more slowly to the bottom than larger particles of the same material [5]. However, due to their high surface-area-to-mass ratios, ENPs readily sorb to soil and sediment particles and consequently are more liable to removal from the water column [73]. Some ENPs might be subject to biotic and abiotic degradation, which can remove them from the water column as well. Abiotic degradation processes that may occur include hydrolysis and photocatalysis [72]. Near to the surface ENPs are exposed to sunlight. It is likely that light-induced photoreactions can account for the removal of certain ENPs and for changing the chemical properties of others [72]. In contrast to the removal processes mentioned above, some insoluble ENPs can be stabilized in aquatic environments. Hoon et al. [74] investigated the aqueous stability of MWCNTs in the presence of natural organic matter (NOM). MWCNTs were readily dispersed as an aqueous suspension and remained stable for over 1 month. Hoon et al. [74] found that NOM is more effective in stabilizing the MWCNTs in water than a solution of 1% sodium dodecyl sulfate (SDS), a commonly used surfactant to stabilize CNTs in the aqueous phase [74]. The C60 fullerenes were found to spontaneously form insoluble, dense aqueous colloids of nanocrystalline aggregates and remain in the aqueous phase for long periods [5]. Another known interaction, which can delay nanoparticle removal from the water column, is the absorption of humic acid. Sea surface microlayers, consisting of lipid-, carbohydrate- and protein-rich components along with naturally occurring colloids, made up of humic acid, may attach ENPs to their surfaces and transport them over long distances [75]. Fate of ENPs in Soil The behavior of ENPs in soil media can greatly vary, depending on the physical and chemical characteristics of the material. Some ENPs can strongly sorb to the soil particles and become completely inert and immobile [5]. On the other hand, if ENPs do not sorb to the soil matrix, they might show even greater mobility than larger particles, because their small size might allow them to travel easily through the pore spaces between the soil particles. The possibility to sorb to soil and the respective sorption strength of ENPs is influenced by their size, chemical composition and surface characteristics [5]. Studies by Zhang [76], Lecoanet and Wiesner [77] and Lecoanet at al. [78] showed considerable differences in mobility of some insoluble ENPs in porous media. The properties of the soil, such as porosity and grain size, further influence the mobility of the particles. Just like the mineral colloids, the mobility of ENPs, agglomerated in colloid-like structures might be strongly affected by electrical charge differences in soils and sediments [76]. Surface photoreactions might induce photochemical transformations on the soil surface [72]. Biodegradation and Chemical Transformation of ENPs In some cases, the biological processes in the environment can lead to the complete degradation of ENPs and sometimes they can only change their physical and/or chemical properties [5]. The mechanisms, which account for