Journal of Hazardous Materials 275 (2014) 79–88

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Transport and retention of zinc oxide nanoparticles in porous media: Effects of natural organic matter versus natural organic ligands at circumneutral pH Edward H. Jones ∗,1 , Chunming Su 2 Ground Water and Ecosystems Restoration Division, National Risk Management Research Laboratory, Office of Research and Development, United States Environmental Protection Agency, 919 Kerr Research Drive, Ada, OK 74820, USA

h i g h l i g h t s

g r a p h i c a l

a b s t r a c t

• ZnO nanoparticles exhibited com•

• • •

plex surface chemistry as different Zn species dominate at different pH. Natural organic matter decreased attachment efficiency facilitating ZnO transport through sand columns. Natural organic ligands did not significantly affect ZnO transport. There were both favorable and unfavorable nanoparticle interactions. Results indicated significant deviation from classical colloid filtration theory.

a r t i c l e

i n f o

Article history: Received 19 October 2013 Received in revised form 4 March 2014 Accepted 26 April 2014 Available online 2 May 2014 Keywords: Nanoparticles Zinc oxide Natural organic matter Natural organic ligands Fate and transport

a b s t r a c t The potential toxicity of nanoparticles (NPs) has received considerable attention, but there is little knowledge relating to the fate and transport of engineered ZnO NPs in the environment. Column experiments were performed at pH 7.3–7.6 to generate effluent concentrations and retention profiles for assessing the fate and transport of ZnO NPs (PZC = 9.3, nominal size 20 nm) in saturated quartz sands (256 ␮m) in the presence of low natural organic matter (NOM) concentrations (1 mg/L humic and fulvic acids) and millimolar natural organic ligands (NOL) levels (formic, oxalic, and citric acids). At circumneutral pHs, ZnO NPs were positively charged and immobile in sand. The presence of NOM decreased the attachment efficiency facilitating ZnO transport through sand columns. Conversely, ZnO transport in the presence of formic and oxalic acids was only slightly improved when compared to ZnO in DI water; whereas, citric acid showed no improvement. The distinct difference between NOM and NOL may have important implications with regard to ZnO transport in the subsurface environment. Experimental results suggested the presence of both favorable and unfavorable nanoparticle interactions causes significant deviations from classical colloid filtration theory (CFT). © 2014 Elsevier B.V. All rights reserved.

∗ Corresponding author. Present address: Geosyntec Consultants Inc., 11490 Westheimer, Suite 150, Houston, TX 77077, USA. Tel.: +1 417 758 3563; fax: +1 281 920 4602. E-mail address: [email protected] (E.H. Jones). 1 National Research Council Resident Research Associate at the U. S. Environmental Protection Agency. 2 U.S. Environmental Protection Agency. http://dx.doi.org/10.1016/j.jhazmat.2014.04.058 0304-3894/© 2014 Elsevier B.V. All rights reserved.

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1. Introduction As nanoparticles are used in commercial applications, they will inevitably be released into the environment as a result of manufacture, transport, application, and disposal practices. Recent work has been conducted to ascertain the effects of velocity and ionic strength upon the fate and transport of engineered nanoparticles (NPs) such as fullerene [1–3], carbon nanotubes [4], and iron oxides [5,6], which ranged in size from ∼1 nm (fullerols and carbon nanotubes) to hundreds of nanometers (silica, fullerene). Metal oxide nanomaterials may also be potentially mobile in some environmental aquatic systems and research on a material-specific basis would assist in resolving this issue [7]. ZnO NPs are being used in personal care products, coatings and paints, on account of their UV absorption and transparency to visible light thus their potential harm to human health has attracted attention [8]. Toxicity studies indicated ZnO NPs were toxic to lolium perenne (ryegrass) and Pseudokirchneriella subcapitata microalgae [8,9]. Information will be required to evaluate exposure risks, to conduct life cycle analysis and to develop waste management strategies [3] as currently there is limited information about the transport of ZnO NPs in the subsurface. Physicochemical properties of NPs govern their transport, which influences their bioavailability and hence toxicity to living organisms [10,11]. In the natural environment, various components of natural waters may interact with NPs and interfere with their transport. Natural organic matter (NOM), which is ubiquitous in natural water, is known to readily adsorb onto colloidal particles. NOM has been found to facilitate the transport of naturally occurring particles [12], model colloids [13,14], and bacteria [15] in saturated porous media. Furthermore, natural organic ligands (NOL) like simple aliphatic organic acids with one to three carboxylic groups such as formic (HCOOH), oxalic (HOOC–COOH), and citric [(CH2 )2 COH(COOH)3 ] acids occur frequently in concentrations up to tens of micromoles per liter in anoxic subsurface environments from biological activities [16,17]. They alter chemical processes in soils through complexation reactions with metal ions in solution and ligand exchange reactions on soil surfaces [18]. Consequently, the presence of NOL may alter the surface properties of NPs facilitating their transportation in saturated porous media. Unfortunately, there has been no reported study on the impacts of NOL on NPs transport and retention. A few studies have examined nanoscale ZnO transport and deposition in aquatic systems. Recently Kanel and Al-abed [19], Petosa et al. [20], and Jiang et al. [21] conducted experiments to determine ZnO NP mobility in the presence of monovalent and divalent salts. Depending on the solution ionic strength, bare nanoscale ZnO elution was observed to increase or decrease over time, suggesting a complex interplay of mechanisms controlled deposition behavior. Nevertheless, the influence of NOM and NOL on ZnO transport has not been systematically evaluated and our current research aimed at bridging this knowledge gap. We present laboratory column transport and deposition experiments to investigate the fate and transport of ZnO NPs in saturated porous media to determine the controlling factors and extent of transport in the presence of NOM and NOL. Effluent concentration data and solid-phase retention profiles were obtained following pulse injections of ZnO NPs under different concentrations of NOM and NOL, which were used to investigate the mechanisms and causes of deviation from classical colloid filtration theory (CFT).

2. Clean-bed filtration theory The most common model for describing filtration of colloidal particles is the clean bed filtration model. Particles suspended in

groundwater contact porous media surfaces (collector) due to three governing mechanisms: diffusion, gravitational sedimentation, and interception [22]. When these conditions occur, concentrations of aqueous-phase particles C(x,t), and retained particles, S(x,t), at column depth x and time t are described by a 1D advection-dispersion equation (ADE) with a first order kinetic term [23]: 2

∂C b ∂S ∂ C ∂C −v + =D w ∂t ∂t ∂x2 ∂x

(1)

b ∂S = katt C w ∂t

(2)

where, v is interstitial particle velocity, D is hydrodynamic dispersion coefficient,  w is the volumetric water content, b is porous media bulk density and katt is particle deposition rate coefficient. According to clean bed filtration theory, the particle attachment rate can be related to single collector efficiency (0 ) and the attachment efficiency factor (˛), which represents the fraction of collisions between suspended particles and collectors leading to attachment by: katt =

3(1 − w )v 0 ˛ 2dc

(3)

For most applications, the system is considered steady state, hydrodynamic dispersion is negligible and katt is spatially and temporally invariant (i.e. a single value is specified). For a continuous particle injection at concentration C0 (at x = 0) and time period t0 , the solution for a column initially free of particles is [22]:

 k  att

C(x) = C0 exp − S(x) =

v

(4)

x



t0 w katt t0 w katt C0 katt C(x) = exp − x b b v

 (5)

Column retention profiles were analyzed using equation (5) to determine katt , which was also calculated from column breakthrough data using ˛BTC . For particles to be effectively transported almost all collisions must be unsuccessful. Attachment efficiency can be estimated from column breakthrough data (˛BTC ) using equation (6) [24]: ˛BTC = −

2dc ln(CL /C0 ) 3(1 − ε)0 L

(6)

where dc is the average diameter of collector particles, ε is porosity, L is column length, C0 and CL are influent and effluent particle concentrations. Values (˛BTC ) approach 1.0 when colloidal interactions are favorable and are «1 when unfavorable conditions dominate (i.e. a net repulsive force between particles and collectors). Attachment efficiency can also be determined from retention profile data (˛Ret ) by comparing the fitted katt from the retention profiles to katt calculated under favorable conditions (kfav ) using equation (3) (i.e. ˛ = 1 and  = 0 ). Values for 0 were determined using equation (7) [20]: 0 = 2.4AS 1/3 NR −0.081 NPe −0.715 NvdW 0.052 + 0.55AS NR 1.675 NA 0.125 + 0.22NR −0.24 NG 1.11 NvdW 0.053

(7)

The effects of the three filtration mechanisms remain separable: the additive terms represent filtration by particle diffusion, interception, and sedimentation, respectively, where AS is the Happel model parameter, NR is an aspect ratio, NPe is the Peclet number, NvdW is the Van Der Waals number, NA is the attraction number and NG is the gravity number, which are calculated from soil porosity, particle size, soil grain size, Darcy (or superficial) velocity, particle density, fluid temperature, viscosity, and particle Hamaker constant. The density of ZnO was taken as 5606 kg/m3 .

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Table 1 ZnO transport experiments conducted in saturated sand columns. Test

Description

C0 (mg/L)a

ZnO PVb /LTc (PV)

pH

Zeta potential (mV)

Pre-flush/post-flush solution

1 2 3 4 5 6 7 8

DI Water DI Water + 1 mg/L Humic Acid DI Water + 1 mg/L Humic Acid DI Water + 1 mg/L Fulvic Acid DI Water + 1 mg/L Fulvic Acid DI Water + 0.004 M Citric Acid DI Water + 0.008 M Oxalic Acid DI Water + 0.016 M Formic Acid

5.035 8.190 8.613 8.165 7.405 6.609 4.655 4.332

8/14 8/14 8/14 8/14 8/14 8/14 8/14 8/14

7.56 7.34 7.52 7.33 7.45 7.39 7.62 7.54

+22.00 −28.70 −26.30 −27.40 −22.30 −14.10 +15.30 +23.70

DI Water DI Water 1 mg/L Humic Acid DI Water 1 mg/L Fulvic Acid 0.004 M Citric Acid 0.008 M Oxalic Acid 0.016 M Formic Acid

a b c

Influent ZnO concentration. ZnO pulse width expressed in dimensionless pore volumes. Number of pore volumes flushed through column (ZnO and DI water).

3. Materials and methods 3.1. Porous media The porous media were composed of 40/60 mesh high purity (>99.7% SiO2 ) quartz sands (Accusand, Unimin Corporation, New Canaan, CT), with an average diameter of 256 ␮m. The column was uniformly wet packed in 1 cm increments and tamped to remove air bubbles giving a porosity of 0.44. The hydraulic conductivity was determined using the falling head technique [25] with a value of 3.20 × 10−3 m/s calculated. 3.2. ZnO nanoparticles

on an orbit shaker (Cole-Parmer Rotatest 110 V) at 150 rpm for the duration of the pulse input. Borosilicate glass columns (Ace Glass Inc., NJ) were 300 mm in length with a 15 mm internal diameter. The influent and effluent ends were sealed with Teflon end fittings and connectors, which had two way valves allowing flow (Ace Glass Inc., NJ). A 14-mm diameter glass filter disc with a 145–174 ␮m pore diameter (Ace Glass Inc., NJ) was located in each end fitting to support the solid phase and promote uniform flow. All experiments were conducted at room temperature (22 ± 1 ◦ C). 3.6. Transport experiments

ZnO NPs (99.5% purity, nominal size 20 nm, 30–50 m2 /g, data from vendor) were obtained from Nanostructured and Amorphous Materials Inc. (Los Alamos, NM). All suspensions were prepared by weighing 10 mg (Mettler PM400, Highstown, NJ) ZnO NPs, which were added to 1 L of deionized (DI) water. Initial average concentration was about 6.4 mg/L with these lower than theoretical concentrations attributed to nanoparticle settling from suspension. This indicated that the nanoparticles present in suspension comprised of both individual nanoparticles and nanoparticle aggregates. The pH range was 7.34 to 7.56 (with and without NOM or NOL). No buffer was used as we were concerned its composition may affect results. Dissolution was considered negligible as previous work indicated dissolution of ZnO NPs in the pH range studied were 2.2% and 1% for 15 nm and 241 nm NPs, respectively [26].

Eight column experiments were conducted to ascertain the transport and retention of ZnO NPs in water-saturated columns (Table 1). Each packed column was flushed for 8 h with DI water or the NOM/NOL being studied to remove background turbidity and provide a uniform collector surface charge. Once flushed, the nanoparticle suspension was introduced. All solutions were applied at a constant flow rate of 1.0 mL/min, which corresponded to a Darcy velocity of 8.14 m/day. Background electrolyte and nanoparticle suspensions were drawn into the column using a peristaltic pump (Masterflex L/S Digital Standard Drive Model 7919–15), with upward flow maintained to ensure the column remained fully saturated and sustained steady-state flow. All tubing was Masterflex chemical durance tubing. An 8 pore volume (PV) pulse of nanoparticle solution was introduced, with the effluent solution collected continuously in 20 mL liquid scintillation vials every 0.33 PV using a Retriever 500 fraction collector (Teledyne ISCO, Lincoln, NE).

3.3. Natural organic matter

3.7. Recovery experiments

Suwannee River humic and fulvic acid standards were used as model NOM compounds. Individual stock solutions of 1 mg/L humic acid and 1 mg/L fulvic acid (both IHSS Suwanee River Standard II, 2S101H, International Humic Substances Society, St. Paul, Minnesota) were prepared by dissolving them in DI water containing 10 mg/L ZnO.

To investigate stability and potential recovery of attached NPs, experiments were conducted with an extended elution phase. To establish particle retention and achieve steady state conditions, an 8-PV pulse of the nanoparticle suspension was introduced, followed by a 6-PV pulse of nanoparticle-free DI water or electrolyte solution. Table 2 shows the most salient properties of the packed quartz sand columns used in the ZnO NPs deposition tests.

3.4. Natural organic ligands 3.8. Analytical methods Individual stock solutions of 0.016 M formic acid (Sigma Aldrich, St. Louis, MO), 0.004 M citric acid, and 0.008 M oxalic acid (both J.T. Baker, Phillipsburg, NJ) were prepared by dissolving these acids individually in 1 L of DI water containing 10 mg ZnO nanoparticles. 3.5. Experimental setup Prior to testing, the particle suspension was placed in an Ultrasonic bath (Fisher Scientific, Howell, NJ) and sonicated for 6 h to ensure the suspension was well mixed and aggregates were dispersed. Once completed, the nanoparticle suspension was placed

X-ray diffraction (XRD) was undertaken using a Ringaku Miniflex X-ray diffractometer (30 kV and 15 mA) to investigate ZnO NP structure. Apparent zeta potential (ZP) and particle size distribution were measured using a Zetasizer Nano-ZS ZEN3600 analyzer (Malvern Instruments, Scarborough, MA, USA) utilizing the Dynamic Light Scattering (DLS) technique. Column effluent samples, pristine ZnO NPs, and impurities in sand samples underwent total elemental analysis by ICP-OES (Perkin Elmer Optima 3300DV ICP) after microwave-assisted HNO3 (10%) digestion. Transmission electron microscopy (FEI CM20 Field emission S/TEM)

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Table 2 Properties and parameters of the packed quartz sand columns used in the ZnO nanoparticle deposition tests. Property

Value

Nanoparticle hydrodynamic diameter (␮m) Collector diameter, dc (mm) Particle density, p (kg/m3 ) Fluid density, f (kg/m3 ) Fluid viscosity,  (kg/m s) Temperature, T (K) Hamaker constant, A (J) Volumetric water content ( w ) Column length (cm) Column Diameter (cm) Happel model parameter

0.145a 0.256 5600 1000 0.001005 298 1.89 × 10−2 ◦ b 0.44 30 1.50 29.91

a b

Based on number mean values from particle size analyzer. Bergstrom, L. (1997).

was used to collect NP images to determine particle size and morphology. Further details are provided in the Supporting Information. 4. Experimental results 4.1. Apparent ZP Apparent ZP results (Fig. 1) indicate the point of zero charge (PZC) of the bare ZnO NPs in DI water was ∼ pH 9.35, which was comparable to previous work [27]. Results indicate a charge

reversal at ∼ pH 7, which is in agreement with previous research [28] for colloidal ZnO. The ZP of ZnO in DI water can be explained by equations S8 to S13. Only H+ and OH− ions are present in DI water and the NP surface charge is only affected by the chemical reaction between the surface and liquid. The adsorption of the resulting zinc species affects both the sign and magnitude of the nanoparticle surface charge. In the pH range 7–8 the Zn2+ (aq) ions become the dominant species. At high pHs, the dominant species are Zn(OH)2(aq) which could precipitate giving a new solid phase, Zn(OH)2(s) . The charge reversal cannot be explained by an increase in concentration of Zn2+ (aq) species with decreasing pH, but can be a consequence of negatively charged impurities (Table S3: Supporting Information). This would indicate the specific adsorption of anions such as bicarbonate (HCO3 − ). Due to NP preparation and sonication prior to column injection, negatively charged impurity ions (HCO3 − ) from the initial ZnO and dissolution of atmospheric CO2 gas maybe present, which could contribute to charge reversal. Previous work indicated ZnO particles have similar ZP–pH relationships but the ZP value at different pH depends upon ZnO purity [28]. That study indicated the higher the concentration of Zn2+ ions, the higher the ZP of ZnO in the pH range 6.50–10.00. Their results showed the maximum positive surface charge increases from +4 mV for 99.99% ZnO to +20 mV for less pure ZnO (99.60%). Our XRD Results (Figure S1) indicated the ZnO NPs were only composed of zincite crystals, while ICP–OES analysis of the 20 nm ZnO NPs indicated the ZnO was 99.40% pure (Table S3) with the major minor impurities being titanium, sodium, and calcium. Although undetected by XRD, trace amounts of NaHCO3 and CaCO3 could be present in the pristine ZnO sample. Composition of the ZnO NPs gave a maximum positive surface charge of ∼+20 mV, which corroborates previous results. Application of 0.008 M oxalic acid and 0.016 M formic acid did not significantly alter the ZP–pH relationship for bare ZnO NPs, in that the PZC for ZnO NPs in the presence of these NOLs was ∼pH 9.30 (Fig. 1b). Results indicate a charge reversal at ∼ pH 7.20, which is comparable to bare ZnO NPs in DI Water. The presence of 1 mg/L humic acid, 1 mg/L fulvic acid or 0.004 M citric acid provided a negative charge to the ZnO NPs in the pH range studied. Adsorbed NOM molecules lowered the ZP of the NPs such that for 1 mg/L humic acid at pH 8.0, the ZP of the ZnO NPs changed from +20 mV to −30 mV. The presence of 0.008 M oxalic acid did not significantly change the ZP–pH relationship for the ZnO NPs compared to results obtained for bare ZnO. A similar ZP–pH relationship to the bare ZnO NPs was obtained when 0.016 M formic acid was applied to ZnO. 4.2. Particle size measurements

Fig. 1. Zeta potential of the zinc oxide nanoparticles in (a) DI water and (b) in the presence of natural organic matter and natural organic ligands.

TEM images (Fig. 2) indicate ZnO NPs were generally spherical with an average diameter between 100 nm and 200 nm. These sizes are much larger than stated by the manufacturer. A possible explanation is that when adding dry NPs to water, they are initially at high local concentrations leading to enhanced collision frequency and rapid aggregation. This is also a thermodynamic-driven process in which particles reduce their surface energy by forming larger aggregates [29]. However, application of 1 mg/L humic and fulvic acid appears to reduce the average size of individual ZnO NPs compared to just ZnO NPs in DI water. Furthermore, in the presence of NOL, TEM images indicate average individual nanoparticle diameters decreased below 100 nm, possibly from NOL-promoted dissolution of primary ZnO particles. Previous work found dissolution rates of ZnO increased below pH 7.50 (yielding Zn2+ ions) and above pH 12.00 [28]. These results suggest NOM at low concentrations affects the size and morphology of individual NPs due to dissolution of

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Fig. 2. TEM image of ZnO NPs at pH 7.50 in (A) DI water, (B) 1 mg/L fulvic acid, (C) 1 mg/L humic acid, (D) 0.016 M formic acid, (E) 0.004 M citric acid, and (F) 0.008 M oxalic acid.

metal oxides by steric protection or ligand-promotion processes [26]. This may have important environmental implications and suggests individual ZnO NPs dissolved in low pH environments or where low concentrations of organic acids exist may change their particle size and morphologies altering their subsurface fate and transport characteristics. These images do not provide detailed information on the levels of aggregation of individual NPs and therefore should be interpreted in conjunction with DLS particle size results. Results from the particle size analyzer indicated the average number mean

hydrodynamic diameter of ZnO NPs in the presence of NOM and NOL was ∼120–145 nm without pH adjustment (pH 7.5) (Fig. 3A). A range of sizes were present comprising individual NPs and nanoparticle aggregates. Furthermore, hydrodynamic particle size measurements (Fig. 3B) for aqueous ZnO suspensions show that with increasing pH, the hydrodynamic diameter of ZnO NPs in the presence of DI water, 0.004 M citric acid, and 0.008 M oxalic acid increased. This was attributed to the pHpzc for these suspensions being near pH 9.30. Increasing the pH from pH 7.50 to pH 9.30 decreased the

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Fig. 3. (A) Number weighted mean particle size distributions of the ZnO suspensions without pH adjustment (pH ≈ 7.5), (B) changes in ZnO hydrodynamic particle size with varying pH.

ZP of the NPs thereby lowering the repulsive interactions, which were electrostatic in nature. This resulted in the NPs aggregating together when the pH = pHpzc . When the pH was further increased from pH 9.30 to pH 10.35 the ZP decreased to −30 mV or lower, increasing the electrostatic repulsion between NPs leading to enhanced stability in suspension. The highest hydrodynamic particle sizes were achieved between pH 7.5 to 9.5 in 0.004 M citric acid though it was not clear why citric acid promoted aggregation of ZnO NPs. 5. Column test results: effluent concentration and retention profiles Pulse injections of ZnO suspensions yielded asymmetrical BTCs that gradually increased to a maximum value before declining sharply to relative concentrations (C/C0 ) approaching zero. Due to the positive surface charge of ZnO NPs in DI water at pH 7.50, concentrations near the inflow were high, approaching zero 15–20 cm into the column (Fig. 4). Results also indicated re-entrainment of ZnO NPs from the sand was negligible suggesting irreversible particle attachment. The ZP of the quartz sand ranged from −31.40 mV to −46.00 mV (Table S1 and Figure S3). When ZnO NPs were surface modified with 1 mg/L humic or fulvic acid, NOM stabilized the ZnO NPs by imparting a substantial negative surface charge. This provided electrostatic stabilization between the NPs and negatively charged collector, resulting in greater mobility of ZnO NPs. This was expressed by increased transport resulting in C/C0 concentrations approaching 0.4 and a lower attachment coefficient (˛). The BTCs indicate a delayed

Fig. 4. Effluent concentrations (A) and retention profiles (B) after pulse injection of 10 mg/L ZnO nanoparticles in columns packed with 256 micron sand in the presence of 1 mg/L humic and fulvic acid: Pre- and post-flushed with DI water.

breakthrough of the ZnO nanoparticles in the presence of NOM suggesting that attachment and displacement of the ZnO nanoparticles onto the collector surfaces maybe occurring. Retention profiles for ZnO NPs in NOM solutions indicate concentrations increased initially near the inlet but after this, concentrations decreased linearly along the remainder of the column, with ratios of sand concentration to influent concentration being approximately 2 for humic and fulvic acids near the outlet. The NOM imparted a negative surface charge and electrosteric stabilization to the ZnO NPs due to binding of carboxylic and phenolic functional groups to the particle surface as well as the molecular size effect of NOM, which facilitated ZnO transport through the column. Electrosteric stabilization is a combination of pure electrostatic repulsion and polymeric repulsion, where the relative importance of each is related to the segment density profile at the interface. Addition of NOM leads to stronger electrostatic repulsion and steric hindrance, which increases the stability of the suspension preventing aggregation [10]. The concentrations of NOM used for surface modification may have improved ZnO transport by blocking strong adsorption sites (i.e. positively charged metal oxides) on collector grains. This was considered unlikely as results obtained for NOM coated ZnO NPs were identical when the system was pre- and post-flushed with DI water or the NOM of interest (Fig. 5). While transport of ZnO NPs improved in the presence of NOM, when 0.008 M oxalic acid and 0.016 M formic acid were studied, the ZP of the ZnO NPs was still positive; this contributed to their rapid deposition. Retention profile data indicated retained ZnO concentrations were greatest near the inlet but generally declined with distance for 0.008 M oxalic acid and 0.016 M formic acid (Fig. 6).

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Fig. 6. Effluent concentration (A) and retention profiles (B) after pulse injection of 10 mg/L ZnO nanoparticles in columns packed with 256-␮m sand in the presence of natural organic ligands in the near neutral pH range.

6. Discussion Fig. 5. Effluent Concentrations (A) and retention profiles (B) after pulse injection of 10 mg/L ZnO nanoparticles in columns packed with 256-␮m sand in the presence of 1 mg/L humic and fulvic acid: Pre- and post-flushed with DI water and NOM.

Nevertheless, transport of ZnO in the presence of formic and oxalic acids were slightly improved compared to ZnO in DI water with higher concentrations of ZnO recorded along the full length of the column. Transport rates for ZnO NPs did not improve when 0.004 M citric acid was applied, with analysis indicating very low ZnO concentrations in the effluent aqueous samples. This result was unexpected, as the apparent ZP of ZnO in the presence of 0.004 M citric acid was ∼−15 mV indicating there should have been some electrosteric repulsion of the ZnO NPs with the sand. However, the total interaction energy might remain favorable and the concentration of citric acid applied was insufficient to prevent NP attachment to the collector. Furthermore, ZnO NPs in the presence of citric acid were less stable compared to other NOLs. This was evidenced by DLS measurements (Fig. 3A), which showed ZnO NPs had a larger mean particle size distribution. Retention profile data indicated high retained ZnO concentrations near the inlet which decreased along the column length indicating rapid attachment of ZnO NPs to the quartz sand once they entered the column.

Characterization studies of ZnO NPs indicated their surface chemistry in DI water was complex with different Zn ion species dominating at different pHs. Determining the characteristics of NOM and NOL adsorption onto ZnO and the associated layer thickness under different geochemical conditions and surface species was beyond the scope of this study but previous work [31] indicated the effect of NOM can be directly measured through surface potential changes, adsorbed layer thickness and mass. Furthermore, solution pH and CaCl2 strongly affect the polymeric characteristics by varying the adsorption affinity of NOM onto surfaces and protonation of functional groups and by neutralizing surface charge and chemically bridging NOM strands together [32]. Continuation of this research to address these parameters is required to determine the effects of steric stabilization of NOM and NOL on ZnO NPs under a wide range of geochemical conditions. Currently there is little information in the literature detailing the fate and transport of ZnO NPs in porous media, and the purpose of this study was to address this data gap. Results indicated naturally occurring polyelectrolytes such as fulvic and humic acids improved transport rates of ZnO NPs by sorbing to particle surfaces and reducing attachment efficiencies through electrostatic or steric stabilization. These results are comparable to those seen in previous work [2,33]. However, in the presence of NOL (citric, formic, or oxalic acid), transport rates of ZnO NPs were poor.

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Table 3 Experimentally determined particle deposition rates, attachment coefficients and filtration lengths. Test

1 2 3 4 5 6 7 8 a b c *

Description

DI Water DI Water + 1 mg/L Humic Acid DI Water + 1 mg/L Humic Acid DI Water + 1 mg/L Fulvic Acid DI Water + 1 mg/L Fulvic Acid DI Water + 0.004 M Citric Acid DI Water + 0.008 M Oxalic Acid DI Water + 0.016 M Formic Acid

BTC

Retention profile

Mass balance (%)

katt a (s−1 )

˛b

FLc (m)

katt a (s−1 )

˛b,*

FLc (m)

0.00822 0.00067 0.00064 0.00067 0.00066 0.00525 0.00525 0.00795

0.8538 0.0700 0.0663 0.0692 0.0681 0.5453 0.5453 0.8259

0.18 2.20 2.32 2.22 2.25 0.28 0.28 0.19

0.00397 0.00041 0.00043 0.00047 0.00054 0.00100 0.00056 0.00090

0.4126 0.0427 0.0446 0.0486 0.0558 0.1039 0.0579 0.0930

0.37 3.60 3.45 3.16 2.75 1.48 2.66 1.65

121.80 87.45 94.88 93.42 102.82 106.56 104.45 122.97

Particle deposition rate coefficient. Attachment coefficient. Filtration length (99.9% removed). Retention profile ˛ determined from katt /kfav where kfav occurs when ˛ = 1.

To quantitatively compare NP transport and interaction behavior, ˛ and katt were determined from both BTC and retention profile data, which were compared to classical CFT. Determining if CFT could adequately explain the experimental results and mechanisms involved was considered the first step to modeling the experimental data. A summary of the results are shown in Table 3, which indicates a discrepancy between ˛ values calculated from retention profile and BTC data. Data analysis indicated katt (and subsequent ˛ values) were in poor agreement with classical CFT. Variations in particle deposition behavior can arise from a number of sources not accounted for in classical CFT. DLVO theory considers the sum of the LondonVan der Waals (VDW) attraction and electrostatic double layer repulsion (EDL) and can be used to determine interaction energy profiles between nanoparticles and collectors. The total interaction energy, namely the sum of the VDW and EDL interactions, was determined by considering the particle–quartz grain system as a sphere–plate interaction. Theoretical expressions [34] were used to calculate VDW interactions accounting for the effect of retardation where a Hamaker constant of 1.89 × 10−2◦ J was selected for the ZnO–water–quartz system. Constant potential EDL interactions were calculated where the ZP of the ZnO particles and quartz sand were used in place of their respective surface potentials [35]. However, it could be argued that DLVO theory is not adequate to describe the behavior of nanoparticles. Experimental results suggest the main cause of the deviation from CFT was attributed to the ZnO population heterogeneity, and pertains to the range of individual particle sizes and aggregates determined during this study. However, multiple potential sources of population heterogeneity could lead to a distribution of deposition rate coefficients. Particle size and TEM results indicated a range of particle sizes were present suggesting nanoparticle aggregates existed rather than individual particles, which is indicated in the mean particle size results–there is a normal distribution of particle sizes. At pH 7.5, the mean nanoparticle aggregate size for all solutions ranged from 95 nm to 250 nm. This lead to a range of particle/aggregate sizes (evidenced by TEM) causing a difference in particle retention. This is important as previous work demonstrated that particle–surface interactions are sensitive to particle size [36–38]. The range of nanoparticle aggregate sizes result in differing repulsive energies and secondary minima leading to a distribution of interaction energies. Calculated interaction energies using the sphere–plate model indicate a very low energy barrier of −0.5 kT in the secondary minima for the NOM solutions studied suggesting that the majority of NPs were not being retained in the secondary minimum (Figs. 7 and 8). In the presence of citric acid, an energy barrier of approximately −1.5 kT in the secondary minima was calculated for the ZnO nanoparticles which could have contributed to the lower than expected recovery in the effluent

solutions as nanoparticles were retained in the secondary minimum. Bradford et al. [39] indicated variations in ␣ can also be attributed to (1) surface charge heterogeneities on NPs or collectors, (2) a fraction of the particles overcoming a repulsive energy barrier to reach the primary minimum or (3) nanoparticles retained

Fig. 7. (A) Calculated DLVO interaction energy plotted as a function of separation distance between the quartz sand and ZnO nanoparticles in the presence of natural organic matter at pH 7.50, (B) the secondary energy attractive region.

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7. Conclusions A number of laboratory column transport and deposition experiments were undertaken to investigate the fate and transport of ZnO NPs in saturated porous media containing known concentrations of NOM and NOL to determine both the controlling factors and extent of transport. From the study it can be concluded that, (1) naturally occurring polyelectrolytes such as fulvic and humic acids enhance ZnO transport even at low concentration of 1 mg/L by sorbing to particle surfaces and reducing attachment efficiencies through electrostatic and/or steric stabilization; (2) naturally occurring small organic ligands at millimolar levels do not significantly promote ZnO transport; (3) katt determined from retention profiles and effluent concentrations was in poor agreement with classical CFT; and (4) nanoparticle population heterogeneity was likely the main cause of the deviation from CFT leading to a range of attachment rate coefficients. Further systematic studies are required that address (i) steric stabilization and adsorption kinetics of NOM and NOL on ZnO NP surfaces at different pHs, surface species and ionic strengths, and (ii) characteristics of the collector surfaces and NP population heterogeneity, which is essential for the identification of key mechanisms governing particle transport and depositional behavior. In-depth modelling of laboratory results using different physical and chemical non-equilibrium models ranging from classical models simulating flow and transport, to relatively mobile-immobile water physical and two-site chemical non-equilibrium models is required. Acknowledgements

Fig. 8. (A) Calculated DLVO interaction energy plotted as a function of separation distance between the quartz sand and ZnO nanoparticles in the presence of natural organic ligands at pH 7.50, (B) the secondary energy attractive region.

This research was performed while the first author held a National Research Council Resident Research Associateship Award at the Robert. S. Kerr Environmental Research Center, US EPA, Ada, OK. Laboratory sample digestion analysis was undertaken onsite by Shaw Environmental Inc. The work upon which this paper is based was supported by the U.S. Environmental Protection Agency through its Office of Research and Development. This work has not been subjected to agency review and, therefore, doesn’t necessarily reflect the views of the agency and no official endorsement should be inferred. Any product or trade name mentioned here is for information purposes only and not to constitute endorsement. Appendix A. Supplementary data

in the secondary energy minimum (not the case herein). DLVO theory assumes NPs and collectors have a constant surface potential, which was not the case herein (Table S2). Chemical analysis of the 256-␮m sand indicated aluminum and iron as main impurities. The aluminum and iron oxides on collector surfaces will be neutrally charged or carry a slightly positive charge in the pH range studied (dependent upon pHPZC ) when compared to the bulk SiO2 surface (negatively charged), which provides favorable deposition sites for nanoparticles in the primary energy well. Tests conducted with columns pre- and post-flushed with DI water and the selected NOM or NOL produced similar results suggesting the effects of surface charge heterogeneities were negligible. The effects of straining could also contribute to the poor agreement with CFT. Recent studies suggest straining occurs when the ratio of particle size to grain size diameter is 0.002 [40] to 0.008 [39]. In the 256-␮m sand, the ratio was 0.00055 indicating straining was probably not a major mechanism for deposition. Whilst beyond the scope of this study, it was anticipated surface roughness of collectors and nanoparticles decreased the repulsive energy barrier for smaller nanoparticles facilitating their deposition on collector surfaces.

Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.jhazmat. 2014.04.058. References [1] H.F. Lecoanet, J.Y. Bottero, M.R. Wiesner, Laboratory assessment of the mobility of nanomaterials in porous media, Environ. Sci. Technol. 38 (2004) 5164–5169. [2] K.L. Chen, M. Elimelech, Interaction of fullerene (C60 ) nanoparticles with humic acid and alginate coated silica surfaces: Measurements, mechanisms, and environmental implications, Environ. Sci. Technol. 42 (2008) 7607–7614. [3] Y. Wang, Y. Li, K.D. Pennell, Influence of electrolyte species and concentration on the aggregation and transport of fullerene nanoparticles in quartz sands, Environ. Toxic. Chem. 27 (2008) 1860–1867. [4] H. Hyung, J.D. Fortner, J.B. Hughes, J.H. Kim, Natural organic matter stabilizes carbon nanotubes in the aqueous phase, Environ. Sci. Technol. 41 (2007) 179–184. [5] H.F. Lecoanet, M.R. Wiesner, Velocity effects on fullerene and oxide nanoparticle deposition in porous media, Environ. Sci. Technol. 38 (2004) 4377–4382. [6] K.A.D. Guzman, M.P. Finnegan, J.F. Banfield, Influence of surface potential on aggregation and transport of titania nanoparticles, Environ. Sci. Technol. 40 (2006) 7688–7693. [7] N.T. Loux, N. Savage, An assessment of the fate of metal oxide nanomaterials in porous media, Water, Air, Soil Pollut. 194 (2008) 227–241.

88

E.H. Jones, C. Su / Journal of Hazardous Materials 275 (2014) 79–88

[8] D. Lin, B. Xing, Root uptake and phytotoxicity of ZnO nanoparticles, Environ. Sci. Technol. 42 (2008) 5580–5585. [9] V. Aruoja, H-C. Dubourguier, K. Kasemets, A. Kahru, Toxicity of nanoparticles of CuO, ZnO and TiO2 to microalgae Pseudokirchneriella subcapitata, Sci. Total Environ. 407 (2009) 1461–1468. [10] B. Xie, Z. Xu, W. Guo, Q. Li, Impact of natural organic matter on the physicochemical properties of aqueous C60 nanoparticles, Environ. Sci. Technol. 42 (2008) 2853–2859. [11] X. Liu, G. Chen, A.A. Keller, C. Su, Effects of dominant material properties on the stability and transport of TiO2 nanoparticles and carbon nanotubes in aquatic environment: From synthesis to fate, Environ. Sci.: Process. Impacts 15 (2013) 169–189. [12] A. Amirbahman, T.M. Olson, Transport of humic matter coated hematite in packed beds, Environ. Sci. Technol. 27 (1993) 2807–2813. [13] A. Franchi, C.R. O’Melia, Effects of natural organic matter and solution chemistry on the deposition and reentrainment of colloids in porous media, Environ. Sci. Technol. 37 (2003) 1122–1129. [14] A.J. Pelley, N. Tufenjki, Effect of particle size and natural organic matter on the migration of nano- and microscale latex particles in saturated porous media, J. Colloid Interface Sci. 321 (2008) 74–83. [15] W.P. Johnson, B.E. Logan, Enhanced transport of bacteria in porous media by sediment-phase and aqueous-phase natural organic matter, Water Res. 30 (1996) 923–931. [16] P.B. McMahon, F.H. Chapelle, Microbial production of organic acids in aquitard sediments and its role in aquifer geochemistry, Nature 349 (1991) 233–235. [17] I.M. Cozzarelli, R.P. Eganhouse, M.J. Baedecker, Transformation of monoaromatic hydrocarbons to organic acids in anoxic groundwater environment, Environ. Geol. Water Sci. 16 (1990) 135–141. [18] C. Su, R.W. Puls, Nitrate reduction by zerovalent iron: Effects of formate, oxalate, citrate, chloride, sulfate, borate, and phosphate, Environ. Sci. Technol. 38 (2004) 2715–2720. [19] S.R. Kanel, S.R. Al-abed, Influence of pH on the transport of nanoscale zinc oxide in saturated porous media, J. Nanopart. Res. 13 (2011) 4035–4047. [20] A.R. Petosa, S.J. Brennan, F. Rajput, N. Tufenkji, Transport of two metal oxide nanoparticles in saturated porous media: Role of water chemistry and particle coating, Water Res. 46 (2012) 1273–1286. [21] X. Jiang, M. Tong, R. Lu, H. Kim, Transport and deposition of ZnO nanoparticles in saturated porous media, Colloid. Surface. A: Phys. Eng. Aspects 401 (2012) 29–37. [22] N. Tufenkji, M. Elimelech, Deviation from the classical colloid filtration theory in the presence of repulsive DLVO interactions, Langmuir 20 (2004) 10818–10828. [23] D. Grolimund, M. Elimelech, M. Borkovec, K. Barmettler, R. Kretzschmar, H. Sticher, Transport of in situ mobilized colloidal particles in packed sand columns, Environ. Sci. Technol. 32 (1998) 3562–3569.

[24] K.M. Yao, A.T. Habibian, C.R. O’Melia, Water and waste water filtration— Concepts and applications, Environ. Sci. Technol. 5 (1971) 1105–1109. [25] C. Liu, Soils and Foundations, in: Upper Saddle River, Prentice Hall, New Jersey, 2001. [26] S-W. Bian, I.A. Mudunkotuwa, T. Rupasinghe, V.H. Grassian, Aggregation and dissolution of 4 nm ZnO nanoparticles in aqueous environments: Influence of pH, ionic strength, size and adsorption of humic acid, Langmuir 27 (2011) 6059–6068. [27] D. Zhou, A.A. Keller, Role of morphology in the aggregation kinetics of ZnO nanoparticles, Water Res. 44 (2010) 2948–2956. [28] A. Degen, M. Kosec, Effect of pH and impurities on the surface charge of zinc oxide in aqueous solution, J. Eur. Ceram. Soc. 20 (2000) 667–673. [29] R. French, A. Jacobson, B. Kin, S. Isley, R. Penn, P. Baveye, Influence of ionic strength, pH, and cation valence on aggregation kinetics of titanium dioxide nanoparticles, Environ. Sci. Technol. 43 (2009) 1354–1359. [31] V.L. Morales, W. Zhang, B. Gao, L.W. Lion, J.J. Bisogni Jr., B.A. McDonough, T.S. Steenhuis, Impact of dissolved organic matter on colloid transport in the vadose zone: Deterministic approximation of transport deposition coefficients from polymeric coating characteristics, Water Res. 45 (2011) 1691– 1701. [32] R.F. Domingos, Z. Rafiei, C.E. Monteiro, M.A.K. Khan, K.J. Wilkinson, Agglomeration and dissolution of zinc oxide nanoparticles: Role of pH, ionic strength and fulvic acid, Environ. Chem. 10 (2013) 306–312. [33] E.H. Jones, C. Su, Fate and transport of elemental copper (Cu0 ) nanoparticles through saturated porous media in the presence of organic materials, Water Res. 46 (2012) 2445–2556. [34] J. Gregory, Approximate expressions for retarded van der Waals interaction, J. Colloid Interface Sci. 83 (1981) 138–145. [35] R. Hogg, T.W. Healy, D.W. Fuerstenau, Mutual coagulation of colloidal dispersions, Trans. Faraday Soc. 62 (1966) 1638–1651. [36] G.M. Litton, T.M. Olson, Particle size effects on colloid deposition kinetics, Colloids Surf. A: Phys. Eng. Aspects 107 (1996) 273–283. [37] T. Phenrat, H.J. Kim, F. Fagerlund, T. Illangasekare, R.D. Tilton, G.V. Lowry, Particle size distribution, concentration, and magnetic attraction affect transport of polymer-modified Fe0 nanoparticles in sand columns, Environ. Sci. Technol. 43 (2009) 5079–5085. [38] T.D. Phenrat, Y. Liu, R.D. Tilton, G. Lowry, Adsorbed polyelectrolyte coatings decrease Fe0 nanoparticle reactivity with TCE in water: Conceptual model and mechanisms, Environ. Sci. Technol. 43 (2009) 1507–1514. [39] S.A. Bradford, J. Simunek, M. Bettahar, Y.F. Tadassa, M.T. Van Genuchten, S.R. Yates, Straining of colloids at textural interfaces, Water Resour. Res. 41 (2005), W10404. [40] S.P. Xu, B. Gao, J.E. Saiers, Straining of colloidal particles in saturated porous media, Water Resour. Res. 42 (2006), W12S216.