Abiotic soil changes induced by engineered nanomaterials: A critical review.

CONHYD-03115; No of Pages 14 Journal of Contaminant Hydrology xxx (2015) xxx–xxx

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Abiotic soil changes induced by engineered nanomaterials: A critical review Ishai Dror ⁎, Bruno Yaron, Brian Berkowitz Department of Earth and Planetary Sciences, Weizmann Institute of Science, Rehovot 76100, Israel

a r t i c l e

i n f o

Article history: Received 13 January 2015 Received in revised form 4 April 2015 Accepted 9 April 2015 Available online xxxx Keywords: Engineered nanoparticles (ENMs) ENM retention Modified soil properties Change in soil matrix Environmental contamination ENM related soil chemistry

a b s t r a c t A large number of research papers on the fate of engineered nanomaterials (ENMs) in the soil– water system have appeared in recent years, focusing on ENM transport, persistence and toxicological impact. It is clear from these publications that soil is a major sink for ENMs, and that only a small portion degrades or is mobilized further into groundwater. However, to date, very few studies have examined the impact of ENMs on the natural soil–subsurface matrix and its properties. Moreover, it is now well accepted that chemical contaminants are capable of changing soil properties either by inducing direct chemical or physical changes, or through indirect changes by, e.g., influencing biological activity that in turn modifies soil properties. Here, we review studies on the deposition, retention, and accumulation of ENMs in soil, indicative of the extent to which soil acts as a major sink of ENMs. We then examine evidence of how these retained particles lead to modification of surface properties, which are manifested by changes in the sorption capacity of soil for other (organic and inorganic) solutes, and by surface charges and composition different than the natural surfaces. Finally, we demonstrate how this results in physical and hydrological changes to soil properties, including hydraulic conductivity, swelling capacity and wettability. The overall picture revealed in this critical review sheds light on a perspective that has received little attention thus far. These aspects of soil change, due to exposure and subsequent accumulation of ENMs, may ultimately prove to be one of the most important impacts of ENM releases to the environment. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Over the last years, it has been observed that irreversible changes in natural soils occur following contact with various chemical contaminants of anthropogenic origin (e.g., Ben-Moshe et al., 2013; Berkowitz et al., 2014; Chen et al., 2000; Pignatello, 2006, 2012; Yaron et al., 2008). The term “irreversible changes” refers to long-term, stable, and persistent transformations of soil physical and chemical properties – on a human time scale – which are also resistant to remediation procedures and to natural attenuation. Some of these changes occur following chemical contaminant adsorption–desorption processes, with irreversibility implied by both hysteresis and soil conditioning ⁎ Corresponding author. Tel.: +972 8 9344230. E-mail address: [email protected] (I. Dror).

phenomena. Due to their size and associated chemical properties, one can question if engineered nanomaterials (ENMs) have a similar impact on the soil solid phase as other groups of chemicals. ENMs form a new class of non-biodegradable pollutants that may reach the land surface as suspended particles in aqueous solution. ENMs are being incorporated into a wide variety of industrial and technological applications, and in consumer products, in increasing quantities. ENM residues are subsequently discharged in industrial and municipal wastewater, which may reach soils via diffuse releases, following effluent disposal and sewage sludge application, and via point releases. Comprehensive information on ENM fate, transport, persistence and toxicological impact as affected by soil solution chemistry and solid phase surface properties appears in many review articles (e.g., Bakshi et al., 2014; Bertsch et al., 2012; Bin et al.,

http://dx.doi.org/10.1016/j.jconhyd.2015.04.004 0169-7722/© 2015 Elsevier B.V. All rights reserved.

Please cite this article as: Dror, I., et al., Abiotic soil changes induced by engineered nanomaterials: A critical review, J. Contam. Hydrol. (2015), http://dx.doi.org/10.1016/j.jconhyd.2015.04.004

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I. Dror et al. / Journal of Contaminant Hydrology xxx (2015) xxx–xxx

2011; Chen et al., 2008, 2010, 2012; Garner and Keller, 2014; Hotze et al., 2010; Ju-Nam and Lead, 2008; Nowack and Bucheli, 2007; Pan and Xing, 2010, 2012; Peralta-Videa et al., 2011; Petosa et al., 2010; Qafoku, 2010; Riding et al., 2014; Wang, 2014; Yang et al., 2014). Existing studies have often been performed on inert porous media materials (e.g., glass beads, quartz, or sand) and, to date, only limited research has appeared that focuses on the impact of ENMs on physical and chemical properties of natural soils. Careful examination of the existing literature, however, indicates modifications of natural soil matrix and properties following exposure to ENMs. This review presents and discusses ENM interactions with natural soils that may alter soil matrix composition and surface properties. Here, the focus is on abiotic ENM-induced changes following direct physicochemical interactions. The effect of ENMs on microbial populations and activity in soil has been studied relatively intensively in recent years (e.g., Frenk et al., 2013; Mohanty et al., 2014; Shah et al., 2014a, 2014b; Yadav et al., 2014); these effects are beyond the scope of the current review. Because chemical contaminants may, under specific conditions, change the matrix and properties of the soil and soil constituents, it is logical to consider that irreversible retention of ENMs on the soil solid phase may also change the soil matrix and its properties. Studies by Elimelech et al. (1995), Brant et al. (2007), and Jiang et al. (2009) were among the first to suggest these possibilities. Here, we present selected evidence for the accumulation of ENMs in soil (Section 2). We then discuss and demonstrate in Sections 3 and 4 how the retention of ENMs in soil can change soil properties and structure. 2. ENM deposition and retention in soils 2.1. ENM–soil solution interactions In the soil liquid phase, the behavior of ENMs is controlled by the solution chemistry, characterized mainly by pH, ionic strength and ionic composition, as well as by the presence of natural organic and inorganic suspended colloids. These colloids can be discrete organic particles, clays, or other phyllosilicates (Citeau et al., 2006). The soil solution also contains dissolved humic substances which may form an amorphous domain coating the suspended minerals, and colloids with a fibril structure expressed in a transect profile by a dimension of 1 nm or less (Wilson et al., 2008). Solution pH and ionic strength can control precipitation of dissolved humic substances and the dissociation of functional groups of humic acids that can modify the negative charge of some ENMs, and consequently define the extent of electrostatic repulsive forces. In natural soil solutions, changes in the humic substance molecules are often due to contractions, which occur when the entire aggregate is compacted following addition of cations (Baalousha et al., 2006). Humic molecule configurations may also change as a result of chelating processes that are affected by the ionic composition and concentration in the soil solution (Jung et al., 2005; Pranzas et al., 2003). In a similar manner, dissolved organic matter in soil solutions may adsorb or even coat ENMs, changing their original surface properties and in turn their interactions with soil solid phase. Aggregation and deposition are the primary processes governing ENM redistribution from the land surface to

groundwater (Lowry and Casman, 2009). In the soil solution, both homo- and hetero-aggregation of ENMs may occur; aggregation between similar particles is known as homoaggregation, while aggregation between different particles is called hetero-aggregation. By summing van der Waals and electric double layer forces, expressed as a function of separation distance between ENMs, Hotze et al. (2010) demonstrated that ENMs can have a net attraction in a primary well (minimum), irreversibly forming aggregates, or in a secondary well, where ENMs are reversibly aggregated. In the soil–water environment, ENM aggregation can be prevented by adsorption of natural polymers and polyelectrolytes, which favors electrostatic repulsions. 2.2. ENM deposition from soil solution to soil solid phase The deposition of ENMs in a soil layer occurs mainly via two distinct pathways: (1) free flow through aggregate channels, and (2) diffuse movement into aggregate pore space. ENM interactions with the soil solid phase may occur within or on the surface of the soil aggregates. The deposition of ENMs in soil can be a result of surface interaction of natural minerals (e.g., clays) and organic matter (e.g., humic acids), or by direct sequestration within the soil pores. The ENM deposition process is often described in terms of the colloid attachment efficiency, a parameter related to the rate between favorable and unfavorable deposition (Petosa et al., 2010; Tufenkji and Elimelech, 2004). Unfavorable slow deposition may occur when the water solution chemistry favors dominating repulsive interactions, which leads to limited ENM deposition. These conditions require the overcoming of considerable energy barriers between ENMs and soil surfaces. In contrast, favorable fast ENM deposition on soil surfaces may occur when the chemistry of the soil solution favors non-repulsive forces between ENMs and soil surfaces. Under these conditions, ENM deposition rates should approach the mass transport limited rate. Suspended particles in the soil solution may be subject to repulsive interactions due to a negative charge of the soil matrix. An abrupt change in the chemistry of the soil solution may alter the deposition pattern. For example, the effect of pH increase on the decrease of hematite content deposited on a quartz surface, due to disappearance of the detachment energy barrier, has been reported (Ryan and Gschwend, 1994). ENM retention on the soil solid phase is not only a result of adhesive interaction, but may be affected by the pore structure and hydrodynamic forces (Bradford and Torkzaban, 2008; Torkzaban et al., 2007). Experimental results of Torkzaban et al. (2007) on deposition and detachment of carboxylate-modified polystyrene latex microspheres, in Ottawa sand and glass bead collectors, showed that significant hysteresis occurred. The extent of this phenomenon was a function of changes in solution chemistry and differences in solid surface morphology. This behavior was attributed to weak adhesive interactions, which depend on the double layer thickness, colloid mass transfer on the solid phase (controlled by the number of grain contacts), and surface roughness. Recently, Landkamer et al. (2013) defined particle population heterogeneity as an additional paradigm to be considered in colloid deposition and re-entrainment pathways. Some suspended ENMs with specific physico-chemical characteristics may be captured by a secondary minimum and retained irreversibly on the soil solid phase. The soil constituents and soil



physical properties are of prime importance in the process of ENM deposition and retention. A series of experiments was conducted to assess the association of water-dispersed, functionalized (acid-treated) multiwall carbon nanotubes (MWCNTs) and the soil minerals kaolinite and smectite, in soil solution, as a function of increasing Na concentration (Zhang et al., 2012). Because the minerals and MWCNTs had negative charges under the experimental conditions, repulsive electrostatic interactions were dominant between them. At the same Na concentration (4 nmol L-1) and mineral dosage (50 mg), smectite demonstrated a low tendency to associate with MWCNTs, while MWCNTs essentially disappeared from the aqueous solution in the presence of kaolinite. This may be explained by the surface charge and hydrophobicity of the two minerals. Based on these results, Zhang et al. (2012) concluded that the major interaction forces between clays and MWCNTs are acid–base interactions, electrostatic forces, and van der Waals forces. For the range of sodium concentrations studied, smectite stabilized MWCNTs in the aqueous phase while kaolinite induced its removal. The more negatively-charged smectite renders MWCNTs more stable than the less negativelycharged kaolinite. In a different study of ENM retention, the aggregation and deposition of ZnO ENMs in saturated sand columns, as a function of solution ionic strength and ionic composition, were reported by Jiang et al. (2012). A rise in ionic strength caused an increase in ZnO ENM deposition, in both NaCl and CaCl2 solutions. At a similar ionic strength, ZnO ENM retention was greater in the presence of divalent ions, as compared to monovalent ions. Jiang et al. (2012) suggested that concentrations of retained ZnO ENMs versus transport distance decreased faster than theory predicts, showing a log-linear decrease under the experimental conditions studied. Aqueous dissolution of ZnO ENMs retained as aggregates on a porous medium solid phase is another factor impacting soils. Bian et al. (2011) studied the dissolution/aggregation of small (4 nm) ZnO ENMs in an aqueous environment, as a function of pH, ionic strength and humic acid. These authors found that, in general, ZnO ENM dissolution increased with a decrease in ENM size. It is also noted that under acidic, oxidizing conditions, Zn2+ is one of the most soluble and mobile trace metal cations (McBride, 1994). Bian et al. (2011) further showed that Zn ions originating from ZnO ENMs are released to the aqueous phase at low and high pH, but the extent of dissolution decreases with increases in pH. Upon reaching the soil solution, Zn ions originating from ZnO ENM dissolution may interact with soil constituents such as clay minerals. Studies of Zn sorption on clay minerals (Nachtegaal and Sparks, 2004; Schlegel et al., 2001; Voegelin et al., 2002) noted its incorporation in newly-forming phases of the hydroxyl interlayered minerals (HIM). The speciation of Zn originating from eight acidic or neutral contaminated soils was determined by Jacquat et al. (2009), who found that a substantial fraction of total Zn (29–84%) was retained in the soil hydroxyl interlayered smectite, with high Zn loading. Analysis of EXAFS spectra of the contaminated soils indicated substantial sequestration into Zn-HIM. Based on these results, the authors suggested that under specific conditions, HIM may strongly affect the speciation of Zn in pristine soils.

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2.3. ENM deposition and retention with soil depth Column experiments that examined the deposition and retention of silver nanoparticles (ENMs), often referred to as AgNPs, in 11 natural Australian soils and 25 German soils with different mineralogical and physical properties were reported (Cornelis et al., 2013; Hoppe et al., 2014). Fig. 1 shows AgNP distributions with depth in soil columns over a large range of soil compositions and properties; for example, clay content, extractable Al, DOC, and pH of Kingaroy soil are reported to contain 56%, 2350 mg/kg, 147 mg/kg, and 5.93, respectively, while of Mt. Shank are reported to contain 22%, 17,400 mg/kg, 606 mg/kg and 5.85, respectively, and of Pinnarroo soil are reported to contain 19%, 670 mg/kg, 141 mg/kg and 7.58, respectively. Cornelis et al. (2013) show that variations over the same soil experiment (denoted as “soil name” 1 and same “soil name” 2 in Fig. 1) were larger than the differences between soil types. It was further found that with the exception of one soil, the AgNPs were retained in the soil column – mostly near the inlet – and did not leach. The authors attributed the variability between the replicates of the same soil to natural heterogeneities in soil porosity and exclusion of heteroaggregates. Cornelis et al. (2013) further suggested that the differential retention and transport of polyvinylpyrrolidone (PVP)-AgNP in these Australian soils (Fig. 1) were the result of favorable deposition of negatively-charged AgNPs to locally positively-charged aluminum on large soil aggregates, and/or straining of AgNPs enlarged by the formation of hetero-aggregates. Studying the retention of sterically stabilized AgNP and silver ions (Ag+), Hoppe et al. (2014) found that high retention of Ag+ ions and low retention of AgNPs vary among soils, which are controlled by their physical and mineralogical properties. The high retention observed in the clay-rich soils was favored by the dissolution of the AgNPs and their homo- and hetero-aggregation with clay minerals. Fig. 1 clearly shows that ENM deposition and retention depend on soil matrix and properties, underlining the role of the host solid phase on ENM deposition. As a consequence, the potential of ENMs to induce soil change will differ from soil to soil. Experiments that show retention of C60 fullerene aggregates in laboratory columns containing CaCl2 solution and glass beads or Ottawa sand were reported by Wang et al. (2014). While up to 49% of the introduced C60 mass was retained in the glass bead columns, retention in the sand columns reached 77%. In contrast, C60 fullerene transport and retention studied in column experiments packed with Appling or Webster natural soils demonstrated complete absence of breakthrough (i.e., retention of all C60 fullerenes) even after prolonged injection (Wang et al., 2010). This large retention capacity was explained by the authors as a consequence of the appreciable content of clays and organic matter in the soils. It was further attributed to possible physical straining and pore throat blockage of the C60 fullerenes. For both soils studied, the retention capacity approached a limiting value of 130 μg g- 1, leading to partial coverage of the solid phase. Zhang et al. (2012) showed that increasing the ionic strength and replacing the background solution from NaCl to CaCl2 enhanced the deposition of C60 in columns of both Ottawa sand and Lula soil, with the effect being more significant for the Lula soil.


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Fig. 1. Depth profiles of AgNP (silver nanoparticle) concentrations in selected columns of natural soils expressed as the ratio of the measured Ag concentration (M) to the total added Ag concentration (M0). Reprinted from Cornelis, G., Pang, L., Doolette, C., Kirby, J.K., McLaughlin, M.J., Transport of silver nanoparticles in saturated columns of natural soils. Sci. Total Environ. 463–464, 120–130, Copyright 2013, with permission from Elsevier.

2.4. Deposition and retention of ENM in soil: conceptual overview A summary of the various mechanisms affecting deposition and retention of ENMs in partially saturated soil is depicted in a conceptual diagram (Fig. 2), which includes processes governing ENM distribution and interactions within the soil environment. This diagram portrays the major phases of the soil environment (soil solution, air and solid phases), and some of the major processes reported to affect ENM deposition and interaction in the soil system.

3. ENM-induced structural changes to natural soil constituents 3.1. ENM-induced structural changes to clays Interactions between ENMs and soil minerals, in general, and clays, in particular, may provide basic information on potential ENM-induced changes to soils. Although some available information on ENM-clay interactions relates to studies on modified clays, this nevertheless provides a perspective on



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air phase H bonding Ligand exchange

Soil-air staining Cation bridging

Hydrophobic interaction NOM

Suspended ENM

Soil/solid phase Dissolved ions Chelation

Soil solution van derWaals interactions

Air water attachment

Fig. 2. A conceptual diagram of the processes governing ENM distribution and interactions with the soil matrix in a partially saturated porous medium (NOM = natural organic matter). (Color available online).

potential ENM impacts on pristine soil clays. Several examples of the association and incorporation of carbon-based ENMs into clay particles, which produced structural changes to the original state of the clay, are presented here. Even during early stages of nanotechnology research, the possibility of interactions between clay minerals and ENMs was examined. Mehrotra and Giannelis (1992) found intercalation of amino-functionalized C60 fullerenes in a mica type silicate. The amino-functionalized C60 intercalation into micatype fluorohectorite was observed by X-ray diffraction (XRD) and IR spectra. The authors showed that insertion of the guest amino-functionalized C60 fullerenes into the quasi-twodimensional host galleries of fluorohectorite, in an aqueous

suspension, is facilitated by the strong and weak interlayer binding forces in the host mineral. The XRD pattern of fluorohectorite in an amino-functionalized C60 aqueous suspension showed a relatively well-ordered multilayered structure corresponding to a primary repeating unit of 26.5 Å or a gallery height of 16.9 Å (Fig. 3a). In contrast, a gallery height of 3.6 Å was observed when only ethylenediamine was intercalated in fluorohectorite. Changes in the mica-type mineral matrix following intercalation of C60 fullerenes into fluorohectorite were also observed in infrared (IR) studies. IR spectra of the C60 fullerenes intercalated silicate, compared to the pure C60 fullerenes spectra, are shown in Fig. 3b. The presence of C60 fullerene clusters is

Fig. 3. Amino-functionalized C60 intercalated in fluorohectorite: (a) X-ray diffraction pattern ((001) and (002) are different planes), (b) infrared spectrum. Adapted with permission from Mehrotra, V., Giannelis, E.P., Intercalation of ethylenediamine functionalized Buckminsterfullerene in mica-type silicates. Chem. Mater. 4, 20–22. Copyright 1992 American Chemical Society.


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reflected in the bands 1740, 1603, 1443, and 1428 cm-1. Mehrotra and Giannelis (1992) determined also the thermal and oxidative stability of amino-functionalized C60 fullerenes using thermo-gravimetric analysis and differential scanning calorimetry. Through both measurement techniques, the authors observed significant enhancement in the thermal and oxidative stabilities of the intercalated C60 fullerenes, as a result of the confinement of C60 fullerenes in silicate galleries. Intercalation of Wyoming sodium montmorillonite by a C60 fullerene derivative (fulleropyrrolidine) was studied by Gournis et al. (2004). Organophilic derivatives were intercalated into organic modified clays, while water-soluble fulleropyrrolidine was introduced into clay galleries by ion exchange. Fig. 4 shows the Raman spectra of pure C60 fullerenes, pure fullerene derivative, and clay/fullerene derivative composite. The derivatization of the C60 fullerene shown in Fig. 4b is manifested by shifts and formation of a new broad band in the Raman spectrum. When the clay fullerene derivate composite was examined (Fig. 4c), the same major peaks (found at 1460, 1421, and 1570 cm-1) were found, proving that the C60 fullerene derivative was intercalated in the clay. Microscopic studies on interactions of clay minerals with water-stable C60 fullerene aggregates indicate a change in clay structure due to association between the fullerene and kaolinite

Fig. 4. Raman spectra of (a) pure C60 fullerenes, (b) pure fullerene derivative, and (c) clay/fullerene derivative composite. Reprinted with permission from Gournis, D., Geortgakillas, V., Karakassides, M., Bakas, T., Kordatos, K., Prato, M., Fanti, M., Zerbrtto, F., Incorporation of fullerene derivates into smectite clays: a new family of organic–inorganic nanocomposites. J. Am. Chem. Soc. 126, 8561– 8568. Copyright 2004 American Chemical Society.

or smectite clays (Fortner et al., 2012). The association (and not intercalation) of smectite with negatively-charged (C60) was not expected to occur; however, some association of C60 with smectite was reported and explained by the presence of local positively-charged sites. Several additional examples of metal-based and silica ENM intercalation into clay minerals, and consequent modification of the clay structure and properties, have been reported in the literature. For example, Hilhorst et al. (2014) discussed the impact of silica nanoparticles on suspensions of well-characterized Wyoming montmorillonite clay. They used cryogenic transmission electron microscopy to probe changes in the structure of the montmorillonite suspensions induced by silica particle intercalation. The authors reported that addition of silica nanoparticles gave rise to a robust reduction in the storage modulus and yield stress of montmorillonite suspensions. Intercalation of silica into another clay mineral, laponite, was discussed by Cousin et al. (2008). In this study, addition of silica-coated maghemite nanoparticles shifted the rheological fluid–solid transition of the laponite. 3.2. ENM impacts on humic substances Humic substances may also be affected upon contact with ENMs. Due to the fact that natural organic material (NOM) affects ENM solubility, most existing studies have focused mainly on ENM—dissolved organic matter (DOM) interactions. Interactions between humic acid and nanosized inorganic oxides (TiO2, SiO2, Al2O3, and ZnO) were studied by Yang et al. (2009) using Fourier transform infrared spectroscopy (FTIR). This study identified several important interactions between functional groups of the organic matter and the ENMs. These interactions occupy some of the more reactive moieties of the organic matter, and hence have the potential to change its structure or reactivity towards other solutes in the soil water. For example, from FTIR spectra of humic acid and TiO2-bound humic acid (Fig. 5), it was observed that COOH groups are responsible for humic acid ligand exchange with TiO2 ENMs. The authors noted a strong interaction of humic acid phenolic OHs with TiO2 ENM surfaces, which may indicate a TiO2-induced modification of the natural humic acid structure. A significant weakening in the COOH peak at 1722 cm-1 was observed in the FTIR spectra of ZnO ENMs bound to humic acid (HA), due to the strong interactions of COOH with ZnO ENM surfaces (Fig. 5d). A fluorescence spectroscopic method was suggested as an additional method to analyze complex formations between HA and C60 fullerenes. Klavins and Ansone (2010) reported that fluorescence signal quenching is proportional to C60 fullerene concentration. They further suggested that this quenching can be used to estimate complex formation. The interactions between C60 fullerene and HA are related to hydrophobic attractions. Indications of chemical change in the soil organic matter composition when CuO ENMs and strong oxidation conditions were applied are presented in Fig. 6. Each panel in Fig. 6 is a comparison between the excitation–emission of initial nonexposed (Bet Dagan, Israel) dissolved soil organic matter and the same compounds after addition of CuO ENMs and H2O2 (which is known to catalyze degradation of persistent organic compounds) (Ben-Moshe et al., 2009; Fink et al., 2012; Yecheskel et al., 2013). In both cases, changes in the humiclike substances were detected, although the change in the



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Fig. 5. FTIR spectra of (a) bulk humic acid (HA) and the following nano-oxide-bound HA, (b) α-Al2O3, (c) γ-Al2O3, (d) ZnO, and (e) TiO2. The differential spectra of nanooxide-bound HA were obtained by subtraction of the IR spectra of pure nano-oxide from that of the HA-coated nano-oxides after freeze drying. Reprinted with permission from Yang, K., Lin, D., Xing, B., Interactions of humic acid with nanosized inorganic oxides. Langmuir 25, 3571–3576. Copyright 2009 American Chemical Society.

amount of ENMs added resulted in different excitationemission matrices, indicating different transformations. A review by Cornelis et al. (2014) indicates that there are many different modes of interactions between dissolved organic carbon (DOM), which is ubiquitous in soil pore water and ENMs. Such interactions include, for example, ligand exchange with surface hydroxyl groups of DOM (Yang et al., 2009) or π–π interactions with aromatic groups of carbon-based ENMs (Sedlmair et al., 2012; Yang and Xing, 2009). Cornelis et al. (2014) noted that adsorbed DOM often coats or replaces the original coating of ENMs, thus changing their behavior. These DOM–ENM interactions have a mutual effect, and can lead also to changes in availability of DOM for other interactions in the soil solution. Some of the results of DOM–ENM interactions can be: (1) capture of many of the DOM functional groups by ENMs that reduce DOM availability for other interactions, (2) change in DOM net charge, (3) steric limitations due to already existing DOM–ENM interactions, and (4) hetero-aggregation. All of

these factors may cause formation of large complexes and/or changes in the behavior of the DOM. 4. ENM-induced changes to soil properties 4.1. ENM-induced changes to soil morphology ENM retention on clays, soils and sediments may lead to changes in their natural surface properties, which impact interactions with constituents/colloids in the soil solution. One example is the change in clay rich Ultisol soil as a result of addition of polyvinylpyrrolidone (PVP)-coated silver nanoparticles (PVP-AgNPs). Wang et al. (2014) observed deposition of PVP-AgNPs on the clay soil surface during transport experiment in water-saturated Ultisol soil columns (Fig. 7a–e). Changes to montmorillonite and goethite specific to Ultisol as a result of PVP-AgNP retention were explained and confirmed by transmission electron microscopy (TEM), as illustrated in


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sandy clay loam Mediterranean soil and to Rendzina soil from an arid zone forest led to ENM coating of much of the soil grain surfaces and changes in their morphology. Guo and Barnard (2013) demonstrated natural distribution of iron oxide nanoparticles on iceberg-hosted sediments. In this case, too, the presence of nanoparticles changed the surface morphology and therefore the interface properties. ENM deposition on soil surfaces may lead to alteration of the aggregate surface charge. Arab and Pourafshary (2013) showed that deposition of metal oxide ENMs on glass beads causes substantial changes in surface charge. For example, the zeta potential of beads upon exposure to nano γ-Al2O3, MgO, and ZnO changed from −44.00 mV for the unexposed beads to +0.82, +0.11, and +1.57 mV, respectively, for distilled water; for the same setup in 0.03 M NaCl solution, the zeta potential changed from −27.60 mV for the unexposed beads to +33.20, −5.7, and −14.20 mV for nano γ-Al2O3, MgO, and ZnO, respectively. Arab and Pourafshary (2013) suggested that the combined effects of the electric double layer repulsion, London– van der Waals attraction, Born repulsion, and hydrodynamic forces control the fate of these ENMs. Charge modification was found to lead to more positive surface potential, which in turn reduces the electric double layer repulsive forces between suspended ENMs and the surfaces. These modifications may trigger a feedback mechanism that increases adsorption of suspended ENMs on the surfaces, and reduce transport of suspended natural and anthropogenic colloids in the soil, thus reducing pore clogging and hydraulic permeability (as discussed further in Section 4.3). Similar phenomena were reported by Cornelis et al. (2014), who relate mainly hetero-aggregation of ENMs and deposition of negatively-charged coated ENMs to changes in the isoelectric point of soil surfaces. 4.2. ENM-induced soil changes leading to interactions with associated contaminants Fig. 6. Changes in 3D excitation–emission matrices of Bet Dagan soil extract after treatment with H2O2 and addition of (a) 1% CuO ENMs, (b) 5% CuO ENMs. Reprinted from Ben-Moshe, T., Frenk, S., Dror, I., Minz, D., Berkowitz, B., Effects of metal oxide nanoparticles on soil properties. Chemosphere 90, 640–646. Copyright 2013, with permission from Elsevier. (Color available online).

Fig. 7. The negatively-charged PVP-AgNPs preferentially bind to the edges of the positively-charged montmorillonite (Fig. 7a). In particular the PVP-AgNPs form larger aggregates (~159 nm in diameter) on the edge of the montmorillonite grain with a relative high charge density (Fig. 7b), as confirmed by EDX analysis (Fig. 7c). The PVP-AgNP interactions with positivelycharged goethite clusters in the soil are shown in Fig. 7d–f. These observations are one example of many which show deposition of ENMs on soil surfaces; they are related to the results discussed in Section 2, but are also important here because deposited ENMs coat part of the surface and yield morphologies and properties different than the natural soil clays. The partial coatings of ENMs have different composition, surface roughness, and charge as discussed below. Other studies show similar deposition of different ENMs on soil. For example, Ben-Moshe et al. (2013) examined CuO and iron oxide ENMs deposition on two soils; the authors found that addition of high concentrations of ENMs added to a red

Exposure to ENMs may lead to changes in the retention capacity of pristine soils for associated micro-contaminants. Studies of interactions between selected environmental microcontaminants and ENM-affected soils reveal such impacts of ENMs. For example, the retention of naphthalene in soil as affected by the presence of C60 ENMs was reported by Cheng et al. (2005). In this study, the C60 was first shown to remain fixed in Lula sandy surface soil. At a later stage, the same soil, in the presence of the deposited C60 at a concentration of 0.18%, was tested for retention of naphthalene in column transport experiments. An increase in the retardation factor from 8.5 to 13.1 for naphthalene was observed by comparing the unamended soil to soil containing C60 ENMs. The increase in retardation was related to sorption of the naphthalene on the fixed C60 ENMs in the soil. Reduced mineralization of 2,4-dichlorophenol in silt loam paddy soil, amended with various concentrations of single and multiwall carbon nanotubes (SWCNTs and MWCNTs), was reported by Zhou et al. (2013). Both SWCNTs and MWCNTs at concentrations of 2 mg CNT/g soil (0.2%) were found to reduce mineralization and increase persistence of 2,4-dichlorophenol. The effect of CNTs was shown regardless of the order of application, i.e., organic contamination followed by CNT deposition or vice versa. The SWCNTs were shown to have a



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Fig. 7. Representative TEM images of PVP-AgNPs associated with soil particles (e.g., montmorillonite and goethite). The negatively-charged PVP-AgNPs were attached onto montmorillonite edges that were positively charged. (a and b) PVP-AgNPs forming large aggregates on the edges of montmorillonite mineral grains (d and e). Goethite provides favorable retention sites for the PVP-AgNPs. The blue boxes marked in (a) and (e) were selected for EDX analysis as shown in (c) and (f), respectively. Reprinted from Wang, D., Ge, L., He, J., Thang, W., Jaisi, D.P., Zho, D., Hyperexponential and nonmonotonic retention of polyvinylpyrrolidone-coated silver nanoparticles in an Ultisol. J. Contam. Hydrol. 164, 35–48. Copyright 2014, with permission from Elsevier. (Color available online).

stronger effect on the degradation and residue distribution of 2,4-dichlorophenol in the soil. Similar results were observed by Towell et al. (2011) regarding the impact of SWCNTs and MWCNTs on the behavior of phenanthrene and benzo-pyrene in a clay loam soil from Preston, UK. An inverse relation between extractability of the phenanthrene and benzo-pyrene from the soil and CNT concentration was reported. Inhibition in phenanthrene mineralization was reported at 0.05% CNT concentration. The authors further reported that the SWCNT had stronger effect compared to MWCNT, as noted in the study of Zhou et al. (2013). Furthermore, the mobility of polyaromatic hydrocarbons (PAHs) in sandy loam soil was found to be affected by the

presence of CNT (Li et al., 2013); significant retention enhancement of PAHs in CNT-amended soil was noted at a concentration of 5 mg g-1. Retention of arsenic in soils following application of stabilized iron-based ENMs – zero-valent iron (nZVI), iron sulfide, and magnetite (Fe3O4) – was reported by Zhang et al. (2010). Two As-pesticide contaminated soils (sandy soil from Washington and red clay soil from Alabama, USA) were amended with the ENMs at Fe/As molar ratios ranging between 5:1 and 100:1, and tested for arsenic content 3 days after ENM application. As the Fe/As increased, higher arsenic retention was observed, with a similar sorption pattern for both soils studied. The Fe3O4


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nanoparticles appeared to be particularly effective for As retention, which increased by 58% in the sandy soil and by 67% in the red clay soil. These results indicate that, in general, stabilized iron-based ENMs may increase arsenic retention, and thus represent an example of ENM-induced impact on the retention capacity of natural soils. Increases in the natural retention capacity of soils as a result of addition of nZVI ENMs, for other inorganic microcontaminants, were discussed by Gii-Diaz et al. (2014). Acidic and calcareous soils from Spain were initially spiked with solutions of Pb(NO3)2 and Zn(NO3)2, incubated for 30 days, and analyzed for concentrations of retained contaminant. A suspension of nZVI was added to the Pb- and Zncontaminated soils, and the retention capacity for Pb and Zn was determined. Examining the sequential chemical distribution of Pb- and Zn-contaminated soils before and after addition of nZVI showed that ENM exposure led to a change in the soil retention capacity of the associated micro-contaminants. A significant decrease in Pb/Zn bound to exchangeable and carbonate soil fractions was observed, while increased sorption of the same elements in the residual fraction was found after nZVI addition to the soils (Fig. 8). 4.3. ENM impacts on soil physical and hydraulic properties Recently, Ng and Coo (2014) studied the effect of adding Al2O3 and CuO ENMs on the hydraulic conductivity of kaolin clays. They found a reduction of 30% and 45% in the hydraulic conductivity, compared to the untreated clay, with 2% addition Al2O3 and CuO ENMs, respectively (Fig. 9a). Further increases in ENM concentrations, up to 6%, led to further reduction in hydraulic conductivity (almost 40% for Al2O3 and almost 70% for CuO; Fig. 9a). These reductions in hydraulic conductivity can be attributed to pore clogging. Measurements of pore size distribution showed a 20% reduction in pore size for the ENMtreated clay. Ng and Coo (2014) further suggested that the stronger effect of CuO ENMs results from the larger diameter of the CuO (40 nm), as compared to the Al2O3 (20 nm). In support of this finding, a similar trend in reduction in hydraulic conductivity was reported by Kananizadeh et al. (2011), who added a natural nanosized material (montmorillonite) to finegrained soil to obtain an order of magnitude reduction in hydraulic conductivity. Clogging of soil pores or pore throats has been observed in additional studies. For example, Tellam et al. (2011) attribute the reduction in breakthrough concentration of metal oxide ENMs in sandstone to their accumulation in soil pores, while Dunphy-Guzman et al. (2006) found that a pH near the point of zero charge led to strong aggregation of TiO2 ENMs. Jeong and Kim (2009) studied the behavior of CuO ENMs in etched glass flow cells, showing that ENMs tended to aggregate at all flow rates, even when surfactants were added to the CuO ENM solution. These aggregates were often deposited in pore throats, which in turn can change fluid flow paths. Fig. 9b shows a microscopic image of CuO ENM accumulation, forming large aggregates (even when treated with surfactant) that block pore throats. Jeong and Kim (2009) further showed that lower flow rates induced higher deposition of CuO ENMs. These results provided a general indication of the fraction of CuO ENMs entering soil that will be deposited and accumulate, thus influencing soil properties.

Fig. 8. Sequential chemical distribution of (a) Pb and (b) Zn in acidic soil (AC), calcareous soil (CA), exchangeable fraction (EX), carbonate-bound fraction (CB), Fe/Mn oxide-bound fraction (OX), organic matter-bound fraction (OM), and residual fraction (RS). Lowercase a denotes significant differences in a soil fraction between treated and untreated acidic soils. Uppercase A denotes significant differences in a soil fraction between treated and untreated calcareous soils. Reprinted from Gii-Diaz, M.M., Perez-Sanz, A., Vicente, A.M., Lobo, M.C., Immobilization of Pb and Zn in soils using stabilized zero-valent iron nanoparticles: effects on soil properties. CLEAN—Soil, Air, Water 43, 1–9. Copyright 2014, with permission from John Wiley and Sons.

The effects of CuO ENMs on the expansion–shrinkage of four soil types were studied by Taha and Taha (2012). The authors found that the strain of both expansion and shrinkage was reduced with addition of optimum concentrations of CuO ENMs to all examined soil types. It was also found that addition of CuO ENMs decreased the development of desiccation cracks on the surface of compacted soil samples, without decreasing the hydraulic conductivity. This behavior was attributed to the high density of the ENMs: when mixed with soil, they increased the overall dry density of the (soil–CuO ENM) mixture, which in turn decreased shrinkage and expansive strains (Fig. 9c). It was further suggested by Taha and Taha (2012) that an increase in agglomerated ENM content led to larger water content through formation of larger amounts of voids in the soil mixture, but without change in the hydraulic conductivity. Similarly, reduced swelling of montmorillonite in the presence of polyethyleneglycol (PEG)-coated silica ENMs was reported (Pham and Nguyen, 2014). A reduction in the magnitude of montmorillonite swelling was found when NaCl solutions were added at concentrations of 0–6% (Fig. 9d). As the ENM concentration increased (up to 1%), the swelling index decreased; further addition of ENMs (up to 3%) did not affect the reduction in the swelling index. A reaction mechanism for such behavior was proposed by Liu et al. (2004), who related enhanced clay



Hydraulic conductivity 10 ( -10 m/s)

b

a

12

11

10 Nano CuO 8

6

4 2 0

2 4 Nanomaterialpercent

6

c d

Fig. 9. (a) Hydraulic conductivity of clay as a function of ENM concentration (from Ng, C.W.W., Coo, J.L., Hydraulic conductivity of clay mixed with nanomaterials. Can. Geotech. J. 2014 © Canadian Science Publishing); (b) Microscope image of CuO ENM aggregates in the presence of surfactant (0.01%); 2D porous medium made of etched glass (Reproduced from Jeong, S.-W., Kim, S.-D., Aggregation and transport of copper oxide nanoparticles in porous media. J. Environ. Monitor. 11, 1595–1600, 2009, with permission of The Royal Society of Chemistry); (c) Effect of CuO ENM concentrations on the total strain of soil (Reprinted from Taha, M.R., Taha, O.M.E., Influence of nano-material on the expansive and shrinkage soil behavior. J. Nanopart. Res. 14, 1–13, 2012, With kind permission from Springer Science and Business Media); (d) Swelling index of montmorillonite in the presence of varying ENM concentrations as a function of NaCl concentration (Reprinted from Pham, H., Nguyen, Q.P., Effect of silica nanoparticles on clay swelling and aqueous stability of nanoparticle dispersions. J. Nanopart. Res. 16, 2014). (Color available online).

aggregation to interaction of polyglycols (found on the PEG coating), which in turn inhibits clay swelling. The application of ENMs to modify the wettability of surfaces and enhance fluid flow in porous media, especially for the oil industry, has received relatively broad attention in recent years. For example, Hendraningrat and Torsæter (2014a) showed that suspensions of metal oxide ENMs injected into Berea sandstone adsorb to grain surfaces and alter the quartz surfaces to become more water-wet, thus reducing the interfacial tension. A reduction of the contact angle from 54o to 21o was found, indicating that metal oxide-based suspensions can render the quartz surface more strongly water-wet. A similar phenomenon was reported by Karimi et al. (2012) for carbonate rock. In this case, nanoZrO2 was found to promote wettability alteration of the rock from strongly oil-wet to strongly water-wet. Ehtesabi et al. (2014) also reported change in rock wettability from oilwet to water-wet after treatment with TiO2 ENMs. The exact mechanism causing this change in surface properties is not clear. It was suggested, however, by Giraldo et al. (2013) that ENMs create ordered structures near the contact line (wetting wedge) of a drop on a solid surface, which promotes spreading

of ENM suspensions along the surface as monolayer particles. This layer has high surface energy which in turn changes the wettability of the surface. The most common ENM used for wettability modification of natural surfaces is nano-SiO2 (e.g., Hendraningrat and Torsæter, 2014b; Ju and Fan, 2013; Safari, 2014). As explained above for the metal oxides, nanosilica is adsorbed onto pore walls, altering surface properties which in turn affect multiphase (oil/water or gas/water) flow patterns through porous media. 5. Conclusions and perspectives Over the last decade, significant research efforts have been devoted to examining the fate and environmental risks related to the growing use of ENMs. Gottschalk et al. (2009, 2013a, 2013b) calculated the emission and distribution of ENMs in various environmental compartments in both the USA and the European Union, for the major types of ENMs in use. They found that soil, sludge and sediments are the major sinks of ENMs, and that while ENM concentrations in all compartments are currently very low, they are constantly increasing


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over time. Moreover, it is clear from many studies that ENMs are retained in soil, and that only a small portion degrades or is mobilized further into groundwater. It is also now well accepted that chemical contaminants are capable of changing soil properties either by inducing direct chemical or physical changes (Berkowitz et al., 2014; Yaalon and Yaron, 1966; Yaron et al., 2008, 2012a, 2012b) or through indirect changes, for example, by influencing biological activity that in turn modifies soil properties. While natural soil formation is considered over geological time scales, accumulated anthropogenic changes to soils resulting from the growing release and deposition of ENMs are relatively fast. This may have substantial implications that are difficult, if not impossible, to reverse in a relatively short time (even if the changes are not immediate). Current literature provides a large body of information regarding soil water chemistry impacts on ENM aggregation and deposition, in relation to their transport to and through soils and groundwater. This information has been obtained mainly by studying carbon- and metal-based ENM interactions with nonreactive earth materials, such as sand, quartz or silica. To date, studies focusing on ENM interactions with soil are limited, and only a few publications have examined the potential for ENMinduced physical and chemical changes to the natural soil system. Based on the existing information, it is suggested here that despite size differences, ENMs, like other chemical contaminants, may cause irreversible changes in physical and chemical properties of pristine soils. It should be noted further (although this was not considered in the current review) that numerous publications have appeared recently on the effect of ENMs on soil biological activity, which in turn can have a strong impact on soil properties. It is clear that additional research on soil–ENM interactions is required to further examine the occurrence of ENM-induced irreversible changes to soil. Studies of ENM retention in clays and humic substances, at the molecular level, should be coupled to soil column experiments under various environmentally relevant conditions (such as different saturation, salinity, and pH levels) where ENM retention and release, and impacts on soil physical status, will provide information on the reversibility (or not) of these processes. Based on evidence provided in this review, and the fact that ENMs are produced and released in growing amounts, accumulating mainly in the soil and sediment compartments, we suggest that much more attention should be devoted to study the fate and behavior of ENMs in the natural environment, with implications to soil sustainability. Acknowledgments We acknowledge the support and constructive comments of three anonymous reviewers. This material is based upon work supported by a research grant from the P. & A. GuggenheimAscarelli Foundation. B.B. holds the Sam Zuckerberg Professorial Chair in Hydrology. References Arab, D., Pourafshary, P., 2013. Nanoparticles-assisted surface charge modification of the porous medium to treat colloidal particles migration induced by

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