Accepted Manuscript Title: Environmental behavior of engineered nanomaterials in porous media: a review Author: Chang Min Park Kyoung Hoon Chu Jiyong Heo Namguk Her Min Jang Ahjeong Son Yeomin Yoon PII: DOI: Reference:

S0304-3894(16)30086-3 http://dx.doi.org/doi:10.1016/j.jhazmat.2016.02.006 HAZMAT 17439

To appear in:

Journal of Hazardous Materials

Received date: Revised date: Accepted date:

4-11-2015 25-1-2016 1-2-2016

Please cite this article as: Chang Min Park, Kyoung Hoon Chu, Jiyong Heo, Namguk Her, Min Jang, Ahjeong Son, Yeomin Yoon, Environmental behavior of engineered nanomaterials in porous media: a review, Journal of Hazardous Materials http://dx.doi.org/10.1016/j.jhazmat.2016.02.006 This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

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Environmental behavior of engineered nanomaterials in porous media: a review

Chang Min Parka, Kyoung Hoon Chua, Jiyong Heob, Namguk Herb, Min Jangc, Ahjeong Sond, YeominYoona*

a

Department of Civil and Environmental Engineering, University of South Carolina, Columbia, 300 Main Street, SC 29208, USA b

Department of Civil and Environmental Engineering, Korea Army Academy at Young-Cheon, 135-1, Changhari, Kokyungmeon, Young-cheon, Gyeongbuk 770-849, Republic of Korea

c

Department of Environmental Engineering, Kwangwoon University, 447-1 Wolgye-Dong Nowon-Gu, Seoul, Republic of Korea d

Department of Environmental Science and Engineering, Ewha Womans University, 52 Ewhayeodae-gil, Seodaemun-gu, Seoul, 120-750, South Korea

*Corresponding author: phone: +1-803-777-8952; fax: +1-803-777-0670; e-mail: [email protected] (Y. Yoon)

Highlights 

Nanoparticle behavior in the soil environment varies depending on various factors.



Numerical approach was discussed for nanoparticle breakthrough curves.



Literature review was conducted on nanoparticle transport in aqueous porous media.



The interactions between nanoparticle and heavy metal ions were outlined. 1

Abstract A pronounced increase in the use of nanotechnology has resulted in nanomaterials being released into the environment. Environmental exposure to the most common engineered nanomaterials (ENMs), such as carbonbased and metal-based nanomaterials, can occur directly via intentional injection for remediation purposes, release during the use of nanomaterial-containing consumer goods, or indirectly via different routes. Recent reviews have outlined potential risks assessments, toxicity, and life cycle analyses regarding ENM emission. In this review, inevitable release of ENMs and their environmental behaviors in aqueous porous media are discussed with an emphasis on influencing factors, including the physicochemical properties of ENMs, solution chemistry, soil hydraulic properties, and soil matrices. Major findings of laboratory column studies and numerical approaches for the transport of ENMs are addressed, and studies on the interaction between ENMs and heavy metal ions in aqueous soil environments are examined. Future research is also presented with specific research directions and outlooks.

Key words: Engineered nanomaterials; Transport; Porous media; Sorption modeling

1. Introduction Nanoscience and nanotechnology are fields that have developed to become significant scientific priorities in research and development. They exploit unique nanoscale properties and phenomena of nanomaterials, which are now used in engineering [1, 2]. Nanomaterials are materials having a domain size of ~100 nanometers or less, and may resemble complexes of protein molecules. However, they differ from protein molecules due to their distinct structural features (e.g., mechanical, electrical, optical, magnetic, and biological properties), chemical composition, shape, size, density, aggregation, and types of surface [3, 4]. Unlike metal salts in solution, the main components of nanoparticles are normally not ionized; their oxidation state is often zero. 2

However, it may be expected that the physicochemical properties of nanomaterials would differ substantially from those of bulk materials [1, 5]. Engineered nanomaterials (ENMs) can be homogeneous or heterogeneous; these include particles with core-shell structures in terms of structure and composition, or multi-functional nanoparticles such as ‘smart’ nanoparticles developed for medical purposes [4]. They have various structures, such as spheres, needles, tubes, and plates, with differing toxic effects in the human body and properties that can respond uniquely to biology or ecology. Due to their small sizes, nanometer-scale materials may pass into cells and translocate through or between cells. Electronic configurations and large surface-to-volume ratios, relative to bulk materials, result in increased particle surface energy with more biologically reactive sites as well as the structures per se [6-8]. Another important parameter of nanomaterials is their chemical composition, which can include metal/metal oxides, compounds, polymers, biomolecules, and various combinations of these components. Some nanoparticles can form aggregates stable enough to have various forms from dendritic structures to chains or spherical structures under ambient conditions [4]. The rapid proliferation of nanoparticle use has resulted in exposure to humans through environmental water, air, and soil systems [3]. Exposure to ENMs can occur directly by way of unintentional release (e.g., release during use and consumption of ENM-containing goods) and intentional injection of ENMs for siteremediation purposes using their advantageous properties, such as high surface area-to-volume ratio, or indirectly via sewage treatment plants, landfills, and incineration plants [9, 10]. The nanoparticle behavior depends strongly on particle-specific properties, solution chemistry, soil hydraulic properties, and soil matrix properties when ENMs are released into the soil environment. Such factors are of importance in controlling the transport process of ENMs and their interaction with heavy metal ions. Thus, it is important to consider use and potential discharge of ENMs into soil environments and to determine the effects of environmental conditions on the behavior of ENMs in aqueous porous media. Recent review articles have mostly covered risk assessment, toxicity, and life cycle analyses regarding the emission of ENMs. Some have also described occurrence,

3

biotransformation of, and interaction between nanoparticles and biological systems using computational modeling techniques such as molecular dynamic methods [11-16]. This review discusses applications and release of ENMs and assesses existing research on the transport of ENMs and the interaction between ENMs and heavy metal ions in soil porous media. Numerical approaches using computational modeling methods and simulation as well as laboratory column studies of ENMs are discussed to understand the transport processes of ENMs in response to complex environmental matrices under both favorable and unfavorable conditions. Finally, the adsorptive behavior of ENMs in aqueous soil environments is outlined, which incorporates solution properties including pH, nanoparticle concentration, and ionic strength.

2. Engineered nanomaterials in the soil environment 2.1. Application of engineered nanomaterials ENM use has increased in electronic, biomedical, personal care, and automotive industries [17, 18]. Generally, engineered nanomaterials can be categorized as carbon-based nanomaterials (such as fullerenes (nC60 or bucky balls), single- or multi-walled carbon nanotubes (SWCNTs, MWCNTs), graphene, etc), metalbased nanomaterials (such as metal oxides (TiO2 and ZnO), gold, silver, etc), and polymeric nanomaterials [2, 4, 19]. Such nanostructured materials are applied to high-performance materials, energy storage and conversion, self-cleaning surface coatings, and stain-resistant textiles [4]. Different phases of ENMs can be mixed to obtain advantageous electrical, optical, or mechanical properties [20]. Other classifications are also used to refer different morphologies and specific groups of nanomaterials, but wide range of nanomaterials with many different properties will substantially impact on their environmental behavior [21]. Carbon-based nanomaterials are the most widely used ENMs and possess unique characteristics. Fullerenes, commonly denoted as ‘nC60’ are important molecules and appear promising in the nanotechnology field, being commonly used in optics [22], electronics [23, 24], and biomedicine [25]. Fullerenes are insoluble in water, forming large aggregates of tens to hundreds of nm (or larger) unless solvent exchange [26, 27] or 4

ultrasonication and stirring in water [28, 29] inhibit the formation of larger clusters, but are soluble in a small range of organic solvents. Additionally, the zeta potentials of fullerenes are negatively charged, although the mechanism by which a negative potential is acquired is still controversial [28]. SWCNTs are a single-layer graphene sheet rolled-up in cylindrical shapes, with a diameter of ~1 nm and a length of several micrometers, and possess industrially useful mechanical, thermal, photochemical, and electrical properties [30]. MWCNTs contain two or more concentric layers of SWCNTs in their structure of various lengths and diameters, and possibly have analogous physicochemical properties [31]. Materials containing CNTs are widely used in spacecraft, artificial muscles, land and sea vehicles, rechargeable batteries, fuel cell production, and electrical conductors, due to their excellent conductivity [32]. Both fullerenes and CNTs carry a strong negative charge (50 mV for fullerenes at or below ionic strengths of 0.05 M and -30 to -70 mV for hydroxylated CNTs), and are highly stable as a result of steric repulsion in the presence of natural organic matter (NOM) [28, 33, 34]. However, they rapidly aggregate into larger clusters at higher ionic strengths in the soil environment [35]. Graphene oxide nanoparticles (GONPs) are layered carbon-based nanomaterials that have been developed and used in various products and industries (e.g., drug delivery, biosensing, and nano- and microelectronics) over the past few decades due to their unique electronic, thermal, and mechanical properties [36-38]. Graphene oxides (GOs) have high solubility in water due to their hydrophilic oxygen-bearing functional groups [39]. GONPs are more resilient to changes in solution chemistry than other nanomaterials due to large numbers of reactive surface O-functional groups and negative surface charges [40]. Rapid growth in the production and use of GONPs could potentially release substantial quantities of GONPs into the natural environment [41]. Metal nanoparticles are potentially useful in sensing, catalysis, transport, and other applications in biological and medical sciences, thus potential diffuse source of nanoparticle contamination [21, 42-44]. Ag nanoparticles have been used in cosmetics and bacteriocides in fabrics due to significant antibacterial properties. However, they are undoubtedly a source of environmental nanoparticles because of the long-term release of silver ions (Ag+) by oxidation of Ag0 in the presence of water [45-48]. Zero-valent iron (ZVI) nanoparticles have received great attention in remediating contaminated groundwater systems attributed to their higher sorbing capacity for a wide range of environmental contaminants, including chlorinated organic solvents 5

and various inorganic compounds [49-57]. However, significant changes in pH and a reduction in the redox potential in groundwater can occur after remediation with ZVI nanoparticles [57]. TiO2 nanoparticles have been used in photocatalysis, photodegradation process, and complete mineralization of toxic organic pollutants in environmental technology [58-60]. ZnO nanoparticles have potential application in ultraviolet light emitting devices, chemical sensors, solar cells, sunscreens, and cosmetics [61-63]. CeO2 nanoparticles have been used as a free radical scavenger, oxygen sensor, and fuel additive [64-66]. Nobel metal nanomaterials (NMNs) exhibit interesting physical and chemical properties in fuel cell and analytical sensors [67]. Au nanoparticles are important for photothermal therapy and bioimaging in the near-infrared absorption [68, 69]; Pt nanomaterials have shown higher electrocatalytic activities toward small molecule oxidation and oxygen reduction reactions in the field of fuel cell than commercial catalysts [70-72]; nanoclusters (Au, Ag, and Pt in particular) provide enhanced applications in catalysis, electronics, photography, photonics, sensing, imaging, medicine, and information storage with controllable molecule-like properties [70, 73]. Polymeric nanoparticles are dominant nanocarriers developed to improve the bioavailability and pharmacokinetics of therapeutics [74, 75]. For example, a lipid-polymer nanoparticles can be a robust drug delivery platform with a high drug encapsulation yield [75]. The discharge of polymeric nanoparticles into the soil environment increasingly occurs as the commercial efforts in nanomedicine industry grow dramatically. However, much remains unknown about the chemistry of polymeric nanosized molecules and modification of nanoscale carriers to assess their fate and impact in the soil environment [74].

2.2. Environmental behavior of engineered nanomaterials The discharge of ENMs into the soil environment is inevitable due to their highly anticipated production and applications. Global production of ENMs is estimated at 350 tons/yr for CNTs, 500 tons/yr for Ag nanoparticles (AgNPs), and 5,000 tons/yr for TiO2 nanoparticles. Soil concentrations have been estimated from those reported figures: 0.01 g/kg for CNTs, 0.02 g/kg for AgNPs, and 0.4 g/kg for nTiO2. Generally, direct nanoparticle release into the subsurface environment originates unintentionally at both point and nonpoint 6

sources. However, ENMs are also intentionally injected into the subsurface for remediation of soil contaminated with heavy metal ions. For example, CNTs have exhibited great potential to remove divalent metal cations such as Pb(II), Cu(II), and Cd(II) at large-scale remediation sites [76-79]. ENMs will transport, transfer, and be transformed within environmental media. Once they enter aqueous porous media, stabilization and transport occur within groundwater flow, which depends on the intrinsic and physicochemical properties of the ENMs, solution chemistry such as pH, ionic strength, and zeta potential, and sediment properties, such as grain size, flow condition, moisture content, porosity, and dissolved organic matter content. ENMs may be deposited, break through the porous media, or interact with heavy metal ions, which has been a major preoccupation for many years because of their toxicity in the environment [77, 80]. One of the key factors that controls the fate and transport of ENMs is stability in aqueous porous media. Well-dispersed ENMs can be transported readily and react with heavy metal ions in natural water systems, but largely aggregated ENMs will be deposited on the soil surface. Aggregation processes reduce the specific surface area as well as the interfacial free energy; thus they limit the reactivity of the ENMs [80]. This stabilization is attributed to the properties of ENMs (e.g., particle size, surface characteristics, and chemical composition) and solution chemistry (e.g., ionic strength, pH, and dissolved organic matter content). Thus, the various environmental factors that control the behavior of ENMs in aqueous porous media should be considered to better understand the transport of ENMs and their interaction with heavy metal ions in porous media.

3. Transport of engineered nanomaterials in porous media 3.1. Numerical approaches for nanoparticle transport Solute transport equations include provisions for kinetic attachment/detachment of the solute as well as linear equilibrium reactions, non-linear non-equilibrium reactions, zero-order production, and two-first-order degradation reactions. Fate and transport models of engineered nanomaterials employ the advection-diffusion equation for one-dimensional movement as follows [81-91]:

7









(1)

where = porosity, C = colloidal concentration in the liquid phase (Nc/L3, Nc = number of particles), S = sorbed 3

solute concentration (M/M), t = time (T), = soil bulk density (M/L ), x = distance traveled (L), M = mass unit, D = diffusion coefficient (L2/T), and q = the Darcy velocity (L/T); all the variables are summarized in Appendix A. The two terms on the right-hand side of Eq. (1) represent the diffusive and advective flux of the system. The interaction between dissolved substances and the soil media is described by various sorption concepts that differ with respect to the involved sorption isotherm, the assumptions made concerning time-dependency, and the sorption reversibility. The multiple sorption regions can be combined to describe the porous media heterogeneity and different accessibility of the potential sorption sites [92]. The isotherm-based models describe the linear/non-linear equilibrium distribution between solid and liquid phases that do not consider the kinetic behavior before steady state is reached, as follows.





(2)



where ks = the adsorption isotherm coefficient (L3/M), = the Langmuir adsorption isotherm coefficient (L3/M), and

i

= the Freundlich adsorption isotherm coefficient. A Langmuir or Freundlich isotherm model can be

constructed using the apparent steady-state concentrations [93]. When ks = 0, adsorption is not considered solute transport; when

= 1, the adsorption equation above becomes the Langmuir isotherm model; when

becomes the Freundlich isotherm model; and when both

= 0, it

= 1 and = 0, it gives a linear adsorption isotherm.

The attachment/detachment approach is based on multiple kinetic processes that were introduced to describe the transport of small particles or bacteria in porous media [94, 95]. In a two-site kinetic model, the sorbed solute concentration, S, is then given by:

S = Satt + Sstr

(3) 8

where Satt = colloidal concentration in the solid phase due to attachment (Nc/M) and Sstr = colloidal concentration in the solid phase (Nc/M) that Bradford et al. (2003) attributed to straining [94]. The attachment/detachment of particles can be described by the following non-mechanistic mathematical functions [81, 82, 85, 89, 91, 94, 96-103]:





(4)





(5)



α

(6)

where ψatt = dimensionless colloid attachment function and Smax = maximum solid phase concentration (Nc/M). The parameter ψatt accounts for time-dependent blocking/filling of retention sites using a Langmuirian approach [88, 104]. It is used to correct the curvature of a deposition curve and accounts for blocking, ripening, and a non-exponential spatial distribution. Smax is normally influenced by electrostatic interactions and hydrodynamic factors, such as flow velocity, particle size, collector size, and grain geometry. Retention decreases with time and depth and becomes more uniform with depth as S approaches Smax, having ψatt = 1 for clean bed conditions. The katt (1/T) and kdet (1/T) indicate the first-order attachment and detachment coefficients, respectively. The particle detachment rate coefficient (kdet) was found to be negligible because a sharp reduction in effluent nanoparticle concentrations was observed upon reintroduction of a nanoparticle-free solution, and no subsequent release upon retention by the solid phase [99]. The θw is the volumetric water content, dc is the mean diameter of the collector or porous medium, and vp is the pore-water velocity. The collision efficiency (α)

9

represents the fraction of particles remaining attached after collisions and η0 describes the frequency of particle collisions with grain surfaces in the porous medium using the following correlation [105]:

η0 = 2.4As1/3NR-0.081NPe-0.715Nvdw0.052 + 0.55AsNR1.675NA0.125 + 0.22NR-0.24NG1.11Nvdw0.053

(7)

where As = the Happel correction factor, NR = the interception number, NPe = the Peclet number, Nvdw = the London-van der Waals attractive forces number, NA = the attraction number, and NG is the gravitational number. The contributions of diffusion, interception, and sedimentation processes to particle collision are represented by the three terms in Eq. (7), while the first term accounts for 99.9% of the calculated η0 value, suggesting that diffusion is the dominant mechanism for particle collisions in the porous media [102]. Particle straining can be described by a first-order relationship with the suspended or liquid phase concentration [81, 82, 85, 88, 89, 91, 94, 99, 101]:





(8)





(9)

where kstr = the first order straining coefficient (1/T), ψstr = the dimensionless depth-dependent straining function, x0 = the starting position of straining (usually the location of the entrance of the column), and β = the empirical fitting parameter for controlling the shape of the curve. Eqs. (4) and (8) above indicate that particle attachment is reversible (i.e., kdet = 0), but particle straining is irreversible. When β = 0, an exponential distribution of retained ENMs with depth is similar to conventional filtration theory, but when β > 0, a hyperexponential shape that has a higher deposition rate close to the column inlet is seen for the retention profile (RP) [100]. It has been reported that a value of β = 0.765 (standard deviation = 0.1), which is used to adequately 10

describe the observed depth-dependency in RP shapes for different sized MWCNTs, was estimated from the average value from column experiments by simultaneously fitting β and katt to breakthrough curves (BTCs) and RPs from sand of three sizes [100]. The value was similar to β = 0.8, obtained by Wang et al. (2012) in MWCNT transport studies [101]. Choy et al. (2008) used β = 0.021 in a series of simulations to re-calculate kstr [99]; however, Bradford et al. (2003) obtained a higher β of 0.43 for a latex colloid-sand column system [94]. Liang et al. (2013a) chose β = 1.532 [106], instead of 0.432 for modeling the transport of AgNPs in sand [89]. As the straining function approaches 1, the influence of empirical straining is maximized. Conversely, the influence of straining decreases with deviation further from the column entrance. Different model formulations can be derived based on Eqs. (1) through (9). Kasel et al. (2013) reported four model formulations for BTCs and RPs of MWCNT: (i) a conventional attachment and detachment model is obtained by setting ψatt × ψstr = 1; (ii) an attachment, detachment, and Langmuirian blocking model is achieved when β = 0; (iii) a depth-dependent retention model is acquired by setting β = 0.765 and setting Smax to a larger value for Eq. (5); and (iv) a time- and depth-dependent retention model is given when β = 0.765 and ψatt is smaller than 1 [100]. Zhang et al. (2012) modeled the BTCs of nC60 using three different models: (i) a two-site transport model in which katt, kstr, and Smax were the fitting parameters; (ii) a modified colloid filtration theory (CFT) model in which katt and Smax were the fitting parameters; and (iii) a CFT model in which katt was the only fitting parameter [85]. Various BTCs for column studies with ENMs were simulated using the HYDRUS-1D code which includes a one-dimensional finite element model to provide a Microsoft Windows-based modeling environment [107]. This code is based on analytical solutions of equilibrium and kinetic transport models that numerically solve the Richards’ equation in saturated-unsaturated porous media. The model parameters of selected soil hydraulic and advective-dispersive solute transport obtained from measured transient or steady-state flow and/or transport data are inversely fitted to experimental BTCs and RPs using a Marquardt-Levenberg non-linear least-squares optimization routine [108]. The column porosity and dispersivity used to simulate solute BTCs for column tests are obtained from non-reactive tracer BTCs [102]. To our knowledge, transport models of the isotherm-based 11

distribution models and the attachment/detachment approach have not yet been systematically compared in column experiments. The model concepts of both approaches differ in whether the interaction occurs in the bulk porous media or its fractions [92]. However, the models are not readily comparable, because the model parameters in the attachment/detachment approach depend on porosity ( ) and soil bulk density ( ), while those used for the isotherm-based models are independent of environmental conditions.

3.2. Breakthrough column studies for engineered nanoparticle transport ENMs have great impact on the soil environment because they are most likely to contact aqueous porous media after their release. The transport process should be considered in determining the environmental behavior of ENMs. Additionally, cleaning up soil and groundwater requires knowledge of the transport behavior of ENMs (e.g., magnetic nanoparticles and CNTs) [78, 109]. Many studies have aimed to investigate the transport behavior of ENMs in saturated/unsaturated porous media. The transport of ENMs is subject to the properties of ENMs (e.g., size, morphology, chemical composition, and surface characteristics), solution chemistry (e.g., ionic strength and pH), soil hydraulic properties (e.g., moisture content of the soil media, flow velocity, and soil permeability), and characteristics of the soil matrix (e.g., grain size, porosity, and dissolved organic matter content). To date, laboratory column studies have been performed to examine the transport of ENMs in aqueous porous media using various materials, including carbon-based and metal-based nanoparticles. Column studies are the most commonly used laboratory experiments to draw solute BTCs for ENM transport. Information in the current literature on ENM transport is summarized in Table 1, which includes a wide range of nanomaterials, solution chemistry, and experimental conditions used in such studies.

3.2.1. Physicochemical properties of ENMs Physicochemical properties, such as chemical composition and surface characteristics, can substantially influence the transport of ENMs in porous media. ENMs with surface functionalization and surface physical modifications using anionic surfactants or polymers are believed to be more mobile in soil matrices because of 12

the stronger electrostatic repulsion between the ENMs themselves and decreased surface chemical heterogeneity. For example, greater retention of sodium dodecyl sulfate (SDS)-wrapped SWCNTs was observed at lower SDS concentrations (0.001-0.05% w/v), with a decreased critical aggregation concentration in porous medium [110]. Sonication-shortened sodium dodecylbenzenesulfonate (SDBS)-dispersed SWNTs (average diameter of 0.8 nm × length of 400 nm) were highly mobile in quartz sand under most experimental conditions, and surface deposition was low due to negatively charged interactive surfaces [111]. MWCNTs functionalized with 4-ethoxybenzoic acid (4-EBAc) were stable in suspension and readily transported through water-saturated quartz sand with an average grain size of 0.36 mm [101]. Anionic surfactants, such as SDBS, introduce negative charges on graphene surfaces to effectively disperse and stabilize graphene nanoparticles (average diameter of 0.8-3 μm × thickness of 0.8-1.2 nm), even at low concentrations (0.004% w/v) in aqueous solutions. Thus, the transport of SDBS-GR was much higher in saturated porous media at low surfactant concentrations [112]. Poly acrylic acid (PAA)-nanoscale zero-valent iron (NZVI) was highly mobile, like a tracer, without significant

retardation, but NZVI itself cannot be transported in porous media under saturated or steady-state flow conditions for a two-dimensional physical model test [113]. Iron nanoparticles prepared with a green tea polyphenol-rich solution (GT-NZVI) were highly mobile so that most of the iron mass, 73% of the injected Fe in uncoated pure silica sand and 62% of injected Fe in aluminum-coated sand, was detected in the effluent [114]. Oxidation and surfactant coating on SWCNTs (diameters of 1.0-2.0 nm × lengths of 500-2000 nm) were effective at solubilizing and stabilizing in aqueous solutions, and the particles were highly mobile in packed sand columns [115]. Humic acid (HA)-coated SWCNTs and MWCNTs were readily transported through porous media and stable in aqueous solution [116]. Excess polymer injection made surface-modified NZVI more mobile in the subsurface media in that the surface charge of sand or clay became uniformly negative to increase the electrostatic repulsion between NZVI and the media. Transport of NZVI suspensions was enhanced significantly in the presence of mixtures of polyacrylate, a low-molecular-weight coordinating polyanion with poly(4-styrenesulfonate) (PSS), a high-molecular-weight non-coordinating polyanion, and bentonite clay, possibly by affecting aggregation of the particles [117]. Triblock copolymer-modified NZVI had the greatest 13

mobility, with the highest apparent zeta potential (-50±1.2 mV) in porous media due to electrostatic stabilization, while bare NZVI was immobile in water-saturated sand columns, with an apparent zeta potential of -30±3 mV [118]. Anionic, hydrophilic carbon (Fe/C), and poly acrylic acid (Fe/PAA)-supported iron nanoparticles demonstrated relatively high transport, inhibiting aggregation and exhibiting a lower sticking coefficient in Ottawa sand and in high-clay soil. In other soils tested (Hagerstown high silt loam and Pope high sandy loam), the transport of iron nanoparticles was limited, with a higher sticking coefficient and lower soil charge density [119]. SDBS-dispersed AgNPs were highly mobile, with less than 15% of the AgNPs retained in water-saturated quartz sand [120]. Conversely, several researchers have revealed that positively charged ENMs stabilized by cationic polymer compounds result in an electrostatic attraction between ENMs and the grain surface. Lu et al. (2014) reported that positively charged CNT suspensions stabilized by cetylpyridinium chloride (CPC) were fully retained in the soil columns due to electrostatic attraction to and/or precipitation on the grain surface with an straining effect, while negatively charged CNT suspensions, stabilized by SDBS and TX-100 (octyl-phenolethoxylate), exhibited mobility more-or-less through the BTCs [121]. Electrostatic interactions between the positively charged nanoparticles and negatively charged silica caused 4% of the injected Fe to be deposited onto pure silica particles [122]. Additionally, increased surfactant concentrations showed poor transport properties for ENMs in porous media because of their tendency to compress the electric double layer. Liu et al. (2015) found that an increase in surfactant concentration from 0.004% to 0.4% reduced the transport of SDBS-GR in clean quartz sand [112]. The surface chemistry of ENMs can have a strong influence on their mobility in porous media. Hydrophobic interactions enhanced the deposition of nanoparticles on soil surface in aquatic systems; thus, the transport of monodisperse suspensions of nC60 particles, which are in principle quite hydrophobic and nonreactive in water, was interrupted through the porous medium compared with fullerol (1.2 nm, monodisperse suspensions) and SWNT (0.7-1.1 nm × 80-200 nm, polydisperse suspensions) [123]. The hydrodynamic diameter and the electrophoretic mobility (EPM) of GONPs were sensitive to ionic strength in that the stability 14

changed from the lower range (10-3 and 10-2 M) to higher ionic strength (≥ 10-1.5 M KCl) [82]. The transport of stable suspensions of GT-NZVI was low in a mixture of calcareous soil and sand because of the progress of the neutralization reactions between the acidic GT-NZVI suspension and soil calcite [119].

3.2.2. Solution chemistry Solution chemistry, such as pH and ionic strength, is an important controlling factor of ENM transport in porous media. pH determines the zeta potential, and thus influences the stability of the ENMs. Minimum stability can occur when pH reaches the point of zero charge (pzc) or isoelectric point. ENM surfaces are negatively charged when the pH is above the pzc, and the zeta potential decreases with increasing pH. A lower zeta potential tends to aggregate ENMs in solution, whereas high zeta potential (negative or positive) imparts the suspensions [80]. Tian et al. (2012) reported that the electrostatic properties were controlled by increased system pH, which reduces the surface deposition of ENMs on natural sand porous media [124]. Many studies have investigated the impact of ionic strength on ENM transport. The aggregation and deposition between ENMs and porous media are strongly dependent on ionic strength. Long-range electrostatic interactions will be screened by dissolved counterions for electrostatically stabilized ENMs in solution. Thus the stability and transport of ENMs will decrease in porous media [118, 125, 126]. The DLVO theory has been used to explore dramatic impacts of ionic strength on the transport and retention of ENMs in more detail [102]. The interaction energy between ENMs is characterized by three regions: attractive primary minimum, repulsive potential energy barrier, and attractive secondary minimum [127]. GONPs are highly mobile at low ionic strengths, but retention increases as the ionic strength increases, mainly through secondary-minimum deposition, as described by the extended-DLVO interaction energy profiles in both saturated and unsaturated porous media (quartz sand with a size of 0.5-0.6 mm) [128]. The deposition of nC60 particles in porous media increased due to a net attractive force near the grain surface and suppression of the electrical double layer as ionic strength increased in the finer water-saturated Ottawa sand column (100-140 mesh, mean grain diameter of 0.125 mm) [129]. MWCNTs (average diameter of 40 nm × length of 400 nm) were relatively mobile in saturated columns packed with acid-cleaned glass beads and quartz sand of two different grain sizes at low ionic strengths of 1 15

mM. However, a slight increase in solution ionic strength resulted in strong deposition of MWCNTs in quartz sand (> 44%) and glass beads (> 39%) at the primary minimum at which attachment to the porous media was irreversible [130]. Fan et al. (2015) showed that retention of GONPs (hydrodynamic diameters of 1.2-1.6 µm at 1 mM and 1.6-3.9 µm at 10 mM) was strongly dependent on ionic strength in mixed Na-Ca electrolyte systems. Charge neutrality and metal ions (Ca2+) bridging lead to a higher deposition rate (Rd) of 98.5% at a higher ionic strength (10 mM) by compression of diffuse double layers, while the maximum Rd was 48.2% at 1 mM [131]. The transport of GONPs was inhibited significantly by 0.5 mM Ca2+ because of complexing with surface Ofunctionalities of both GONPs and soil components that can cause aggregation of GONPs and bridge GONPs with soil grains [82]. It was found that increased ionic strength or the addition of divalent cations, such as Ca2+ ions, caused higher deposition of carboxyl-functionalized SWCNTs in a well-defined porous medium composed of clean quartz sand with an average grain diameter of 263 µm, consistent with conventional colloid deposition theories [132]. The retention of GONPs was influenced strongly by ionic strength change in saturated quartz sand packed with an average diameter of 0.254 mm [133]. Jaisi and Elimelech (2009) reported that the deposition rate of carboxyl-functionalized SWCNTs increased as ionic strength increased, and divalent cations, 2+

such as Ca ions, were more effective in the retention of SWCNTs in columns packed with a natural soil, ranging in size from 0.42 to 1.0 mm [134]. Higher ionic strength in the presence of polysaccharide-type organic matter favors the retention of nC60 particles on porous media, such as a groundwater aquifer or a water treatment filter, and increases the potential for exposure, whereas lower ionic strength in the presence of humiclike substances tends to favor transport of nC60 particles [135]. The deposition of nC60 particles (55.7±1.8 nm) was largely irreversible, and the deposition rate of nC60 particles increased as the electrolyte concentrations increased [136]. The nC60 particles were readily transported through water-saturated quartz sand (40-50 mesh, mean grain diameter of 0.335 mm) at a low ionic strength of 3.05 mM after introducing less than 1.5 pore volumes of nC60 suspensions [102]. Increasing the ionic strength of NaHCO3 reduced effluent concentrations with higher carboxymethyl cellulose (CMC)-NZVI due to increased particle-collector attachment efficiency [137]. High Ca2+ concentrations (40 mM) significantly reduced the mobility of PAA-NZVI due to destruction of the PAA gelling network and weakening of the repulsion of the electric double layer [125]. 16

Electrolyte properties are another factor controlling the extent of ENM. Jaisi et al. (2008) suggested that SWCNTs would deposit in a secondary energy minimum in the presence of monovalent ions, but would deposit in a primary energy minimum in the presence of Ca2+. Thus, elimination of the secondary energy minimum would only influence the mobility of SWCNTs in low ionic strength solutions [132]. GONP deposition is sensitive to increases in ionic strength in the presence of Ca2+, while no obvious difference in the retention of GONP was found when ionic strength was increased in the presence of Na+ (Rd was 35.2% and 38.21% at 1 and 10 mM). A greater influence on GONP deposition was observed from the molar ratio of Ca2+/Na+ in solution at a 2+

higher ionic strength of 10 mM [131]. The electrophoretic mobility was more effectively reduced by Ca than Na+ ions in aqueous solution with nC60 particles. However, the deposition rate dropped sharply because of significant concurrent aggregation of nC60 which reduce the convective-diffusive transport toward the grain surface when the electrolyte concentrations approached or exceeded the critical coagulation concentration (CCC) [136]. Zhang et al. (2012) found that increased deposition of nC60 was observed in both sand and soil columns when the background solution was switched from 1.5 mM NaCl to 0.5 mM CaCl2 because of a significant decrease in the primary repulsive energy [85]. Wang et al. (2008) reported that approximately 95% of the introduced nC60 particles was retained in the finer water-saturated Ottawa sand column (100-140 mesh, mean grain diameter of 0.125 mm) at a higher ionic strength of 30.1 mM, regardless of the type of electrolyte (NaCl or CaCl2). The complexes of divalent cations (Ca2+) with negatively charged surface sites resulted in strong deposition of nC60 onto the surface due to repulsive energy reduction near the surface at low ionic strengths and a larger secondary minimum at higher ionic strengths. In the presence of NaCl, similar deposition on grain surfaces was found in that the secondary attractive region also became larger at higher ionic strengths (30.1 mM), while the primary repulsive energy was reduced to approximately125 kT [129]. A reduction in ionic strength or cation exchange of monovalent ions for divalent ions reduces the bridging interaction or expands the electrical double layer to further release ENMs from the soil. Several researchers have shown that a lower ionic strength solution can be used to release previously retained ENMs using chemical perturbation. For example, a reduction in ionic strength from 100 mM to 1 mM NaCl released up to > 99% of previously retained GONPs, while relatively higher retention of GONPs was found at 100 mM NaCl [128]. Jaisi 17

et al. (2008) reported that the deposited SWCNTs were released from the quartz sand upon the introduction of low ionic strength solution following deposition experiments with KCl [132]. The retention of AgNPs was much more pronounced in the presence of Ca2+ than K+ at the same ionic strength due to cation bridging, but decreasing the solution’s ionic strength enhanced the mobility of surfactant-stabilized AgNPs in undisturbed loamy sand soil under unsaturated conditions [106].

3.2.3. Soil hydraulic properties Soil hydraulic properties, such as moisture content and flow conditions, can influence the transport of ENMs. Moisture content is important in the retention of ENMs. Less recovery of GONPs in unsaturated sand columns than in saturated sand columns in sand columns at the same ionic strength was observed [128]. A reduction in moisture content promoted the retention of SWCNTs in the porous media for which grain surface attachment and thin-water film straining were responsible [115]. ENMs were more mobile at higher flow velocities in soil, but little effect on the transport of nC60 was observed with a change in flow velocity [85, 138]. For example, a favorable deposition of nC60 particles was obtained at the lowest flow velocity of 0.38 m/d on a soil column packed with Lula soil (0.27% organic carbon). The maximum percentages of nC60 breakthrough (C/C0) were 47% (very limited mobility), 60%, and 80% at flow velocities of 0.38 (typical groundwater flow), 3.8, and 11.4 m/d, respectively [139]. The BTCs of C60 clusters (nC60, 168 nm) did not exhibit velocity dependence of particle passage through the porous media with spherical silicate glass beads having a mean diameter of 355 μm when two different Darcy velocities (0.04 and 0.14 cm/s) were examined at an ionic strength of 0.01 M NaCl [138]. In saturated porous media, SWCNTs were parallel to the streamlines of the flow, so that retention by physical trapping was not observed, but in unsaturated porous media retention occurred near the column inlet at lower moisture content (< 0.10) and was sharply reduced with travel distance in the effluent of 40-50 mesh Ottawa sand [101].

18

3.2.4. Soil matrix characteristics Characteristics of the soil matrix will greatly influence the transport of ENMs. The physicochemical properties of the soil media, such as grain size, porosity, and dissolved organic matter content can determine the maximum solid phase concentration (Smax) on the grain surface, and thus the migration of ENMs. It has been observed that retention of MWCNTs in soil columns had a greater dependency on grain size due to the effect of pore structure in a greater number of retention locations [140]. Jaisi and Elimelech (2009) found that physical straining governs filtration and transport of SWCNTs due to their very large aspect ratio, highly aggregated state in aqueous solutions, and heterogeneity in soil particle size, porosity, and permeability [134]. Lu et al. (2013) reported that the transport of SDBS-, octyl-phenolethoxylate (TX-100)- and CPC-functionalized MWCNTs was unfavorable in smaller sized sands, which had greater specific surface areas and larger numbers of sites for sorption and retention of MWCNTs [141]. Zhang et al. (2012) showed that ENM transport was inhibited in Lula soil, with an average grain size of 120 μm, because of the smaller grain size, more irregular and rougher shape, and greater heterogeneity [85]. The effect of dissolved organic matter on the transport of ENMs is quite complex. ENMs are more mobile in the presence of NOM via steric repulsion. For example, the mobility of GONPs dispersed mostly as irregular with a thickness less than 2 nm, was enhanced in the presence of 10 mg/L Suwannee River HA (SRHA), and significantly inhibited the stacking of GO flakes at a high ionic strength of 35 mM NaCl in saturated quartz sand, with an average grain size of 0.26 mm [91]. The transport of GONPs was enhanced in the presence of 10 mg/L SRHA at 35 mM in quartz sand with much smaller grain sizes than heterogeneous Lula soil containing 45% sand, 36% silt, and 19% clay [82]. The presence of NOM increased the stability and transport of GONPs at 0.1-10 mg/L KCl and CaCl2 salts and the presence of a complex assortment of ions in subsurface water environments [142]. It was also reported that NOM facilitated the transport of AgNPs more deeply into the subsurface environment, thus increasing the potential risk of groundwater contamination [81, 89]. In contrast, soil organic matter such as humic acid (HA) was less likely to control the mobility of CNTs than soil texture in column studies packed with quartz sands based on Pearson correlation analyses [121]. Nanoparticle aggregation 19

with humic or fulvic acid and settling of aqueous suspensions of nC60 particles resulted in limited mobility in the spherical glass beads of porous media with ionic strengths greater than 0.001 M [143]. Cornelis et al. (2013) reported that the interaction between polyvinylpyrrolidone-coated AgNPs (40 nm diameter) and natural colloids in soils reduced their mobility due to favorable deposition and enhanced straining of AgNPs following heteroaggregation with soil colloids [81].

4. Interaction between engineered nanomaterials and heavy metal ions Heavy metal pollution has become a serious issue with the rapid development of some industries. Heavy metal ions in soil environments do not degrade biologically, but accumulate in organisms. Many researchers have investigated the removal of heavy metal ions from wastewater using adsorption behavior on carbon-based nanomaterials as superior adsorbents, or magnetic nanoparticles as a novel adsorbent [76, 77, 144]. The adsorption process has been developed as a simple and effective treatment method to remove heavy metals released into the environment from natural sources or anthropogenic activities. Among several adsorbents, such as biopolymers, activated carbon, metal oxides, clays, and microorganisms, engineered nanoparticles have been found to be the most effective adsorbent for the removal of heavy metals due to their extremely small particle sizes, large surface areas, and higher adsorption capacities [145-149]. However, limited information is available to describe the interaction between ENMs and heavy metal ions in aqueous porous media. The adsorptive behavior is strongly affected by solution properties including pH, initial metal concentration, nanoparticle dosage, and background competing ions for adsorption sites controlling the surface charge [52]. Adsorption of heavy metals on nanoparticles can be characterized using spectroscopic and microscopic techniques, but must be calibrated by macroscopic adsorption isotherms or surface complexation models (Table 2) [98, 145, 150154]. Equilibrium sorption data are modeled using notable adsorption isotherms, such as the Langmuir, Freundlich, Temkin, and Dubinin-Radushkevich isotherms, to describe the macroscopic relationship between the equilibrium concentration of the adsorbent and the amount of adsorbate bound to the surface at a constant 20

temperature [155, 156]. The Langmuir isotherm model describes the formation of a monolayer adsorbate on the finite identical sites of the outer surface of the adsorbent quantitatively to represent the equilibrium distribution of metal ions between the solid and liquid phases [155]. Uniform energies of adsorption onto the surface are assumed for the Langmuir isotherm with no transmigration of adsorbates in the plane of the surface. This yields a plateau in the Langmuir isotherm at high concentrations of adsorbate, corresponding to site saturation. The Freundlich adsorption isotherm is commonly used to describe reversible heterogeneous sorption characteristics that are not restricted to monolayer adsorption capacity [157]. In many cases, the Freundlich isotherm model provides a better prediction of metal ion adsorption because of the empirical incorporation of adsorbent and site heterogeneity [158]. The Temkin isotherm is characterized by a uniform distribution of binding energies that ignores extremely low and large concentrations, while assuming the heat of adsorption of all molecules in a layer [159, 160]. The Dubinin-Radushkevich isotherm does not assume a homogeneous surface or a constant adsorption potential, allowing it to provide more information on the chemical and physical properties of the adsorption process [161, 162]. It is used to express high solute activities and an intermediate range of concentrations with a Gaussian energy distribution on a heterogeneous surface. However, none of these models are capable of explaining the effects of pH or ionic strength on metal ion adsorption on the particle surface without changing the capacity terms. Surface complexation models have been developed to incorporate the impact of pH and ionic strength, and the relationship between surface charge and potential, see Table 2 [163]. In many ways, they share the assumptions of the isotherm models, as an extension of the Langmuir model in the adsorption process between adsorbate, vacant surface sites, and occupied surface sites, except that the reaction involves the release or consumption of a proton. The surface sites are typically represented by ‘>MeOH,’ and those sites are often treated as diprotic acids, described in the classic 2-pK approach, as follows:

>MeOH + Hs+ → >MeOH2+

K10

(10)

21

>MeOH → >MeO- + Hs+

K20

(11)

The development of surface charges apparent in the reactions described in Eqs. (10) and (11) is responsible for the formation of an electrical double layer. Intrinsic equilibrium constants are determined by equilibrium adsorption computer codes such as ECOSAT or FITEQL that combines a non-linear least squares fitting routine with a chemical model for aqueous speciation and adsorption, yielding the optimum fit to experimental adsorption data [164, 165]. In the equilibrium adsorption computer codes, a series of mass-action constraints are included, and a set of species is defined as the product of thermodynamic reactions with a set of components. An iterative strategy is used to solve the chemical equilibrium problems. These constants are regarded as nonadjustable parameters to describe metal ions at the particle surface over a range of pH, ionic strengths, and competing metal ions and ligands. The sorptions of metal ions are defined by complexation reactions, located at various distances from the surface, based on the metal ions’ affinity for the surface and electrostatic interactions with the functional groups. Different representations have been used for different surface complexation models in terms of the relationship between surface charge and potential and the location of the sorbing species. Early versions of surface complexation models include the Diffuse Double Layer Model (DLM), developed by Stumm’s group (around 1970), and the Constant Capacitance Model (CCM), developed by Schindler’s research group. Davis and Leckie [163, 166] first proposed a generalized mechanistic triple layer model (TLM) that provides a reasonable description of the electrical double layer, incorporating an inner Helmholtz surface plane (0-plane), an outer Helmholtz plane (β-plane), and the classical diffuse plane (Fig. 1). The TLM was revised by Hayes and Leckie [167] to account for the effects of ionic strength to allow adsorbates to form inner or outer sphere surface complexes at the interface; this depends on whether a water molecule is intercalated between the surface and its adsorbed ions. The surface metal ion is termed an outer sphere complex if at least one molecule of water resides between the sorbing metal ions and the surface [168, 169]. Regardless of the type of surface

22

complexation model, the outermost layer is regarded as the diffuse layer, described using the classic GouyChapman theory [170].

5. Conclusions and areas of future study The release of ENMs into the soil environment is inevitable due to the rapid development and application of nanotechnology. Humans and the environment can be directly exposed to ENMs via intentional injection for remediation purposes and unintentional release or indirectly via landfills, waste incineration, and sewage treatments [9, 10]. The behavior of ENMs is strongly dependent on their particle-specific properties and background environmental conditions when they are discharged. Recent reviews have described modeling and analytical studies to provide ENM concentrations in the environment, life cycle assessments of ENMs to investigate the sensitivity of algae, daphnia, and fish to ecotoxicity, interactions between nanoparticles and biological systems with computational modeling, simulations of nanoparticles in biological systems, and transformations of nanomaterials [12, 14-16]. However, the literature reporting the potential discharge of ENMs and their behavior in aqueous-phase porous media, based on physicochemical properties of ENMs, solution chemistry, soil hydraulic property, and soil matrix is limited. Many studies have investigated the transport of ENMs in saturated/unsaturated porous media with numerical simulations of BTCs for column studies, although laboratory systems might not represent the complex nature of aqueous porous media. ENMs are stabilized and transported readily under ‘unfavorable’ conditions such as surface functionalization, surface physical modifications using anionic surfactants or polymers, increased system pH, low ionic strength, and high flow velocity. Under ‘favorable’ conditions such as destabilization by cationic polymer compounds, increased surfactant concentrations, increase in ionic strengths, low moisture content, and small grain size, which enhance the retention of nanoparticles on porous media, limited mobility has been observed with less extensive travel distances in porous media. The mechanisms for ENM transport of associated with dissolved organic matter remain unclear. In addition, 23

information on the effects of other complex factors, such as sunlight, temperature, redox potential, and biomacromolecules (e.g., proteins and polysaccharides) is quite limited, which calls for further research. The intentional injection of ENMs for remediation purposes can be one of the sources of nanoparticle contamination in soil environment. Recently, CNTs and GONPs have been used as effective adsorbents, and use of magnetic nanoparticles has attracted interest for the uptake of heavy metals using polymer modifications, functionalization, and nanoparticle composites, because the adsorption processes are considered effective, constituting one of the most popular processes due to their simplicity, convenience, and efficiency [109, 171187]. Few studies have examined the interactions between ENMs and heavy metal ions in aqueous porous media. Future efforts should be directed toward defining the relationship between metal-based nanoparticles and metal ions bound to their surfaces. In terms of modeling approaches, most of the adsorption studies for heavy metals on ENMs have used macroscopic adsorption isotherms. Further research is needed to address the impact of pH and ionic strength on heavy metal ions adsorption to advance the current state of knowledge.

Acknowledgments This research was supported by the Korea Ministry of Environment, ‘GAIA Project, 2015000540003’.

Appendix A

Symbols >

bulk adsorbent

As

Happel correction factor

AT

Temkin isotherm equilibrium binding constant [L3/M]

bT

Temkin isotherm constant

C

colloidal concentration in the liquid phase [Nc/L3]

Ce

equilibrium concentration of adsorbate [M/L3] 24

Ci

Helmholtz capacitance

D

diffusion coefficient [L2/T]

dc

mean diameter of the collector or porous medium

E

mean adsorption energy,

F

Faraday’s constant

I

ionic strength

Kad

Dubinin-Radushkevich isotherm constant [mol2/kJ2]

Kanion

anionic electrolyte binding constants

katt

first-order attachment coefficients [1/T]

Kcation

cationic electrolyte binding constants

kdet

first-order detachment coefficients [1/T]

Kf

Freundlich isotherm constant [M/M]

K i0

intrinsic acidity constant

KL

Langmuir isotherm constant [L3/M]

ks

adsorption isotherm coefficient [L3/M]

kstr

first order straining coefficient [1/T]

n

adsorption intensity

NA

attraction number

Nc

number of particles

NG

gravitational number

NPe

Peclet number

NR

interception number

Ns

surface site density

Nvdw

London-van der Waals attractive forces number

q

Darcy velocity [L/T]



2

25

qe

amount of metal adsorbed per gram of the adsorbent at equilibrium [M/M]

Qo

maximum monolayer coverage capacity [M/M]

qs

theoretical isotherm saturation capacity [M/M]

R

universal gas constant (8.314 J/mol/K)

RL

Equilibrium parameter (separation factor),

s

surface

Satt

colloidal concentration in the solid phase due to attachment [Nc/M]

Smax

maximum solid phase concentration [Nc/M]

Sstr

colloidal concentration in the solid phase [Nc/M]

T

absolute temperature [K]

t

time [T]

vp

pore-water velocity

x

distance traveled in the column [L]

x0

starting position of straining in the column

z

charge of ionic species





Greek symbols α

collision efficiency determined by the fraction of particles remaining attached after collisions

β

empirical fitting parameter for controlling the shape of the curve

i

ε

Freundlich adsorption isotherm coefficient Dubinin-Radushkevich isotherm constant, Langmuir adsorption isotherm coefficient [L3/M]

η0

single collector efficiency

θ

porosity

θw

volumetric water content 26

1



ξ

zeta potential

ρ

soil bulk density [M/L3]

σi

surface charge at position i

ψatt

dimensionless colloid attachment function

ψi

surface potential at position i

ψstr

dimensionless depth-dependent straining function

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40

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Fig. 1. Schematic diagram of the structure of triple layer model and the charge distribution at a solid-liquid interface representing the placement of ions in the electrostatic plane modified from [192-195].

41

Table 1. Summary of soil column experiments on the transport of selected ENMs. Nanomaterials

Transport system

Darcy velocity

Soil media

Solution chemistry

2.4-8.4 cm/min

spherical silicate glass beads d50 = 0.355 mm, θ = 0.43 ζ-potential = -29.8 mV water-saturated porous media, glass beads or Ottawa sand d50 = 0.36 mm spherical glass beads d50 = 0.355 mm, θ = 0.43 ζ-potential = -25.5 mV

0.01 M NaCl

Main findings

Ref.

nC60 were more mobile at higher flow velocities in soil, but little effect of flow velocity on the transport of nC60 was observed. nC60 particles were readily transported through water-saturated quartz sand at a low ionic strength of 3.05 mM. Aggregation with humic or fulvic acids and settling of nC60 particles resulted in limited mobility in the spherical glass beads of porous media with ionic strengths greater than 0.001 M.

[138]

Carbon-based nanoparticles nC60

nC60

nC60

packed column D: 2.5 cm, H: 9.25 cm packed column D: 2.5 cm, H: 15 cm packed column H: 9.25 cm

0.19 cm/min 2.4 cm/min

nC60

nC60

nC60

packed column D: 2.5 cm, H: 9.25 cm packed column D: 2.5 cm, H: 15 cm

nC60

packed column D: 9 mm, H: 5 cm

nC60

packed column D: 2.65 cm, H: 10 cm

nC60

packed column D: 0.66 cm, H: 6.3-6.8 cm (Ottawa sand) and 6.0-6.2 cm (Lula soil)

2.4 cm/min 0.51 cm/min (40- to 50-mesh OS) or 0.55 cm/min (100- to 140mesh OS) 0.061.97 cm/min

spherical glass beads d50 = 0.355 mm, θ = 0.43 ζ-potential = -29.8 mV saturated Ottawa sand (40- to 50mesh and 100-140 mesh) d50 = 0.335 and 0.125 mm, θ = 0.363-0.395

1.0 mM CaCl2 0.1, 0.01 and 0.001 M NaCl 1 to 300 mM NaCl 0.1 to 1.0 mM CaCl2 0.01 M NaCl

3.05 and 30.05 mM NaCl or CaCl2

The deposition rate of nC60 particles increased as the electrolyte concentrations increased, and the electrophoretic mobility was more effectively reduced by Ca2+ than Na+ ion. The transport of nC60 particles in principle quite hydrophobic and nonreactive in water was interrupted through the porous medium. The deposition of nC60 particles in porous media increased as ionic strength increased in the finer water-saturated Ottawa sand column. nC60 particles were retained in the finer water-saturated Ottawa sand column at higher ionic strength of 30.1 mM, regardless of the type of electrolytes (NaCl or CaCl2).

[102]

[143]

[136]

[123]

[129]

Lula soil (0.27% organic carbon) d50 = 0.25 mm, θ = 0.4 specific surface area = 1.24 m2/g ρb = 2.62 g/cm3

0.01 M NaCl/0.01 M NaN3

a favorable deposition of nC60 particles was obtained at the lowest flow velocity of 0.38 m/d on the soil column packed with Lula soil.

[139]

7.2 cm/min (fast) or 2.4 cm/min (slow)

spherical silicate glass beads d50 = 0.36 mm, θ = 0.36

Higher ionic strength in the presence of polysaccharide-type organic matter favors retention of nC60 particles, whereas lower ionic strength in the presence of humic-like substances will tend to favor transport.

[135]

0.050.69 cm/min

Ottawa sand (mainly pure quartz) and Lula soil (45% sand, 36% silt, and 19% clay) d50 = 0.25 and 0.12 mm, θ = 0.36-0.38 and 0.44-0.46 specific surface area = 8.38 m2/g ρb = 1.63-1.67 and 1.40-1.45 g/cm3

0.01-0.6 M NaCl, Na2SO4, (NH4)2SO4, NaNO3, CaCl2 and MgCl2 1-10 mM NaCl, 0.5 mM CaCl2, 0.5 mM Na2SO4 and 5.0 mg/L fulvic acid

Increased deposition of nC60 was observed in both sand and soil columns when the background solution was switched from 1.5 mM NaCl to 0.5 mM CaCl2.

[85]

42

MWCNTs

packed column D: 2.5 cm, H: 10 cm

0.013.06 cm/min

four soil samples d50 = 0.036-0.89 mm, θ = 0.560.67 ζ-potential = -11.6--51.3 mV

MWCNTs

packed column D: 1.0 cm, H: 15 cm

5.64 cm/min

quartz sand d50 = 1.22 ± 0.21, 0.89 ± 0.19, and 0.44 ± 0.09 mm, θ = 32.2 ± 0.1 - 33.6 ± 0.4

CaCl2 increased from 0 to 0.88 mM

MWCNTs

packed column D: 2.5 cm, H: 15 cm

0.41 cm/min

1 and 10 mM

MWCNTs

packed column D: 2.5 cm, H: 5 cm

MWCNTs

packed column D: 2.5 cm, H: 10 cm

2.9 x 103 , 5.3 x 10-3 and 2.9 x 104 cm/min 0.2 cm/min

saturated glass beads and quartz sand d50 = 0.1–0.2 mm (fine) and 0.5– 0.6 mm (medium), θ = 0.38 (fine) and 0.40 (medium) Four quartz sands d50 = 0.35 mm ζ-potential = -50 mV

saturated Ottawa sand (40-50 mesh) d50 = 0.36 mm

SWCNTs

packed column D: 1.6 cm, H: 4.3 cm

0.080.17 cm/min

SWCNTs

packed column D: 1.6 cm, H: 6.3 cm

SWCNTs

packed column D: 2.5 cm, H: 15 cm

SWCNTs

packed column D: 2.5 cm, H: 15 cm

SWCNTs

packed column D: 1.5 cm, H: 7.5 cm

0.18-2.8 cm/min

porous medium θ = 0.45 specific surface area = 0.789 m2/g

SWCNTs, MWCNTs

packed column D: 2.5 cm, H: 15

1.0-2.0 cm/min

acid-cleaned, baked, and natural sand

The positively charged CNT suspensions stabilized by cetylpyridinium chloride (CPC) were fully retained in the soil columns due to electrostatic attraction to and/or precipitation on the grain surfaces with the straining effect. Soil organic matter was less likely to control the mobility of CNTs than soil texture in quartz sands. Transport of SDBS, octyl-phenol-ethoxylate (TX-100) and CPC-functionalized MWCNTs was unfavorable in sands which had greater specific surface areas and larger numbers of sites for sorption and retention of MWCNTs. Reduction in moisture content promoted the retention of SWCNTs in the porous media for which grain surface attachment and thin-water film straining were responsible.

[121]

[141]

[115]

The retention of MWCNTs in soil columns had greater dependency on grain size due to the effect of pore structure in a greater number of retention locations.

[140]

6 and 75 mM

MWCNTs functionalized with 4-ethoxybenzoic acid (4-EBAc) were stable in suspension and readily transported through water-saturated quartz sand. In the saturated porous media, retention of SWCNTs was not observed, but in unsaturated porous media retention occurred near the column inlet at low moisture content (< 0.10) and was sharply reduced with travel distance in the effluent of 40-50 mesh Ottawa sand.

[101]

Cheshire fine sandy loam (sand 58%, silt 13% and clay 29%) d50 = 0.42-1.0 mm specific surface area = 3.67 m2/g

0.1 to 100 mM KCl and 0.03 to 10 mM CaCl2

The deposition rate of carboxyl-functionalized SWCNTs increased as ionic strength increased, and divalent cations were more effective in the retention of SWCNTs in a natural soil.

[134]

1.09 (± 0.012) cm/min

clean quartz sand d50 = 263 mm, θ = 0.37

0.1-55 mM KCl and a mixture of 7 mM KCl with 1 mM CaCl2

[132]

0.487.11 cm/min

quartz sand d50 = 0.1-0.2 (fine), 0.5-0.6 (medium), 1.4-1.6 mm (coarse), θ = 0.38-0.40 ζ-potential = -19.73 mV

The transport of GONPs was inhibited significantly by complexing with the surface Ofunctionalities of both GONPs and soil components. The increased ionic strength or the addition of divalent cations caused higher deposition of carboxyl functionalized SWCNTs in clean quartz sand. Sodium dodecylbenzene sulfonate (SDBS) dispersed SWNTs were highly mobile in quartz sand, and surface deposition was low due to negatively charged interactive surfaces. Oxidization and surfactant coating on SWCNTs were all effective in solubilizing and stabilizing in aqueous solutions and highly mobile in packed sand columns.

[115]

Greater retention of sodium dodecylsulfate (SDS)wrapped SWCNTs was observed at lower SDS concentrations (0.001 - 0.05% w/v) with the decreased critical aggregation concentration in porous medium.

[110]

The electrostatic properties were controlled by the increased system pH reducing surface deposition

[124]

granular Quartz sand d50 =0.1-0.2 (fine), 0.5-0.6 (medium) and 1.4-1.6 mm (coarse), θ = 0.4 ζ-potential = -19.59 (fine), -19.73 (medium) and -27.56 mV (coarse)

43

10 mM NaCl and 1.0 mM CaCl2

[111]

cm

d50 =0.5-0.6 mm, θ = 0.4 ζ-potential = -19.4, -41.1 and 45.2 mV at pH 5.6, 8 and 10 (acid-cleaned sand), -15.3, -35.4 and -38.8 mV at pH 5.6, 8 and 10 (baked sand) and -15.1, -33.3 and -36.4 mV at pH 5.6, 8 and 10 (natural sand)

CNTs

of ENMs on natural sand porous media.

porous media

0.01-10 mM KCl and CaCl2

Humic acid (HA)-coated SWCNTs and MWCNTs were readily transported through porous media, and were stable in aqueous solution.

[116]

GONPs were highly mobile at low ionic strengths, but retention increased as ionic strength increased mainly through secondary-minimum deposition.

[128]

GONPs

packed column D: 2.5 cm, H: 16.5 cm

0.2 cm/min

porous media (quartz sand) d50 = 0.5-0.6 mm ζ-potential = -63.82, -54.37, -33.9 mV at 1, 10, 100 mM NaCl

1, 10 and 100 mM NaCl

GONPs

packed column D: 2.5 cm, H: 15 cm

0.31 cm/min

saturated sand packs with quartz sand d50 = 0.254 mm, θ = 0.37 specific surface area = 8.91 x 10-3 m2/g ρb = 1.8 g/cm3

1, 5, 20 and 100 mM NaCl

The retention of GONPs was strongly influenced by ionic strength change in saturated quartz sand packs.

[133]

GONPs

packed column D: 6.6 mm, H: 10 cm

0.69, 0.35 and 0.07 cm/min

saturated quartz sand d50 = 0.26 mm (50–70 mesh, 0.21–0.30 mm), θ = 0.42 ± 0.01 ρb = 1.53 ± 0.03 g/cm3

10-50 mM NaCl

[91]

GONPs

packed column D: 6.6 mm, H: 5 cm

11.4 cm/min

quartz sand d50 = 275 μm, θ = 0.45 ± 0.02

GONPs

packed column D: 6.6 mm, H: 10 cm

0.69, 0.35 and 0.07 cm/min

GONPs

packed column D: 4 cm, H: 20 cm

0.66 ± 0.002 cm/min

Lula soil (45% sand, 36% silt, and 19% clay) and Sigma sand d50 = 120 mm (Lula soil) and 260 mm (Sigma sand), θ = 0.42-0.47 (Lula soil) and 0.40-0.43 (Sigma sand) saturated porous media θ = 0.36 ± 0.005 ρb = 1.62 g/cm3

0.3–30 mM MgCl2, 1.510 mM NaCl and 0.1–10 mM CaCl2 10-50 mM NaCl, 0.30.5 mM CaCl2

The mobility of GONPs dispersed mostly was enhanced in the presence of 10 mg/L Suwannee River HA (SRHA), and significantly inhibited stacking of GO flakes at high ionic strength of 35 mM NaCl in saturated quartz sand. The presence of NOM increased the transport and stability of GONPs at 0.1-10 mg/L KCl and CaCl2 salts.

GONPs

packed column D: 2.5 cm, H: 16.5 cm

0.2 cm/min

clean Quartz sand porous media d50 = 0.55 mm, θ = 0.44

loamy sand d50 = 1, normal adsorption; 1/n < 1, cooperative adsorption The smaller 1/n, the greater the expected heterogeneity Linearly decrease in heat of adsorption of all molecules in the layer with coverage

Ref. [188]

[157, 189, 190]

[159, 160]

ln

DubininRadushkevich

 

Properties KL refers to the bonding energy of adsorption related to free energy and net enthalpy If RL > 1, unfavorable; RL = 1, linear; 0 < RL < 1, favorable; RL = 0, irreversible

sinh(zFψd/2RT) ψ 0 = ψd Requires estimation of three parameters (K10, K20, Ns) from a limited set of experimental data



σ0 = C 1 × ψ 0



Requires estimation of four parameters (K10, K20, C1, Ns) from a limited set of experimental data



σd = -0.1174 × I1/2 × sinh(zFψd/2RT)



σ0 = (ψ0 - ψβ) × C1



σ0 + σβ = (ψβ - ψd) × C2 = -σd



Requires estimation of seven parameters (K10, K20, Kanion, Kcation, C1, C2, Ns) from a limited set of experimental data



CCM incorporates only a single plane



Classic 2-pK models assume one unit charge per bond, leading to MeO-, MeOH0, and MeOH2 Metal ions with finite size placed on two planes simplified from the Stern layer model



a

All variables are defined in Appendix A.

46

Temperature dependent If E < 8 kJ/mol, physisorption dominates; E is between 8 and 16 kJ/mol, the adsorption process follows chemical ion-exchange Metal ions behave as point charges Only innersphere complexes are accounted for

[161, 162, 183, 191]

[192]

[192]

[192-195]