# Flocculation into, and resuspension of, engineered nanomaterials from coagulated sludges

Jonathan Brant, Luke Ruff, Jérôme Labille

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Jonathan Brant, Luke Ruff, Jérôme Labille. Flocculation into, and resuspension of, engineered nano-materials from coagulated sludges. Environmental Nanotechnology, Monitoring & Management, 2019, 12, pp.100257. 10.1016/j.enmm.2019.100257. hal-02414696

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# Flocculation Into, and Resuspension of, Engineered Nanomaterials from Coagulated Sludges

Submitted to:

**Environmental Nanotechnology, Monitoring and Management**

Jonathan A. Brant, P.E., Ph.D. <fnref n="1" />*, Luke Ruff <fnref n="1" />, and Jerome Labille Ph.D. <fnref n="2" />

**Keywords:** Flocculation; engineered nanomaterials; nanoparticle tracking analysis, nanoparticle concentration

October 10, 2020

<footnote n="1">¹ University of Wyoming, Department of Civil and Architectural Engineering, 1000 E. University Avenue, Laramie, WY 82071</footnote>

<footnote n="2">² Centre Européen de Recherche et d'Enseignement des Géosciences de l'Environnement, Europole Mediterraneen de l'Arbois BP80 – 13545, Aix-en-Provence, cedex 4, France</footnote>

<footnote>* Corresponding author: Email: jbrant1@uwyo.edu</footnote>
Abstract

Engineered nanomaterials (ENMs), like nano-silver (nAg), cerium dioxide (CeO₂) and titanium dioxide (TiO₂), are garnering growing attention from an array of industries due to their unique attributes. As their commercial production has grown, so too has their environmental footprint. Ingestion of water containing ENMs has been identified as a primary human exposure route. This makes understanding ENM removal, and its permanence, in conventional water treatment systems imperative. In this study, the removal of three ENMs (nAg, CeO₂, TiO₂) by flocculation and clarification in complex media was evaluated using environmentally relevant concentrations. ENM homoaggregation and heteroaggregation with aluminum sulfate, natural organic matter and clay were characterized using a combination of light scattering and nanoparticle tracking analysis. This approach provided an unprecedented assessment of ENM aggregation and concentrations persisting in clarified waters distinct from previous works using more gross measures of ENM characteristics/concentrations. Under optimum conditions the mass removal of all ENMs was ≥84%; however, the particle concentrations in the clarified water were >10⁷ particles/mL. This corresponded to particle surface area values of 3.3, 25, and 2.0 cm²/L, respectively for the CeO₂, citrate-nAg, and TiO₂. Only the nAg showed a tendency to be released from the coagulated sludge through chemical and/or physical erosion processes suggesting that sludges containing ENMs may serve as point sources of ENM pollution during land application.
## 1.0 Introduction

Engineered nanomaterials (ENMs) are a significant factor in the material demographic, with widespread implementation into numerous commercial products. For example, titanium dioxide (TiO₂) ENMs are used in paints (Kiser, Westerhoff et al. 2009), cosmetics/sunscreens (Nohynek, Lademann et al. 2007), and as food additives (Lomer, Thompson et al. 2000). Nano-silver (nAg) and cerium dioxide (CeO₂) are two other ENMs that are widely used in commercial products. nAg is used as additives in textiles (Benn and Westerhoff 2008) and medical equipment to create anti-microbial surfaces/environments, while CeO₂ are used as fuel additives (Cassee, van Balen et al. 2011). According to Hendren et al., in 2011 the annual domestic production values for nano-TiO₂, nAg, and nano-CeO₂ were approximately (upper bound) 38,000, 20 and 700 tons/year, respectively (Hendren, Mesnard et al. 2011). In light of the expected growth in the nanotechnology market, and the corresponding increases in ENM production values, it would appear that the introduction of these novel materials into the environment is imminent.

Environmental introduction routes for ENMs may include direct discharge from municipal/industrial wastewater effluents (Gottschalk, Sonderer et al. 2009, Kiser, Westerhoff et al. 2009, Gottschalk, Sun et al. 2013, Hendren, Badireddy et al. 2013) and/or through the decomposition of nano-products (Bystrzejewska-Piotrowska, Golimowski et al. 2009). For example, nAg has been shown to leach from fabrics during washing (Benn and Westerhoff 2008) to subsequently enter sewerage. One outcome from this will be the presence of ENMs in some form in potable water supplies. Previous studies have measured variable levels of different ENMs in surface water systems, including 0.7-16 µg/L for TiO₂ (Mueller and Nowack 2008), 0.03-0.08 µg/L for nAg (Mueller and Nowack 2008), and 0.1-1 µg/L for CeO₂ (O'Brien and Cummins 2010). As this data suggests, ENM concentrations in potable water supplies are expected to be in the micro to nano-gram per liter range (Mueller and Nowack 2008). The presence of ENMs in potable water systems requires that we understand how, and to what
extent, they are removed by different treatment barriers. So called conventional treatment
processes, namely flocculation, clarification, and media filtration remain as the most widely used
processes in surface water fed treatment systems. Therefore, evaluating ENM removal by these
process is prudent. Also, of consequence is evaluating ENM removal under realistic
concentration scenarios (in the µg/L range), something that to date is absent from the literature.

ENM removal by flocculation and clarification processes has been addressed and
generally found to follow expectations based on coagulation theory. Research hurdles for ENMs
include measuring ENM removal and matching observations made at high ENM concentrations
with outcomes at more realistic concentrations. Hyung et al. used UV-vis and HPLC
measurements to quantify the removal of bare and natural organic matter (NOM) coated
Buckminsterfullerene (C60) by flocculation with aluminum sulfate (alum) (Hyung, Fortner et al.
2007). Removal, which ranged from 20 to 98%, was a function of solution pH and coagulant
(alum) dose in accordance with conventional expectations. The presence of natural organic
matter (NOM) resulted in an increase in coagulant consumption relative to that required in the
absence of NOM to achieve a similar C60 removal efficiency as that obtained in the absence of
NOM. This was attributed to coagulant consumption by the NOM, and to a lesser extent
electrosteric stabilization effects (Wiesner and Bottero 2007). Transformation of any ENM is a
function of the material and the dispersing media (Petosa, Jaisi et al. 2010). There is a great
body of literature detailing how NOM and its various fractions can modify the physical and
chemical characteristics of ENMs. Delay et al. and Chinnapongse et al. determined that NOM
reduced the propensity nAg to aggregate in high ionic strength solutions (Chinnapongse,
MacCuspie et al. 2011, Delay, Dolt et al. 2011). In separate studies Quik et al. and Li and Chen
observed that the aggregation tendency of CeO2 is reduced following the adsorption of humic
acid (Quik, Lynch et al. 2010, Li and Chen 2012). This finding has been confirmed by other
groups studying CeO2 (Hoecke, Schamphelaere et al. 2011) and TiO2 (Honda, Keene et al.
2014) aggregation. Exceptions to these observations arise when polyvalent cations are present
in solution. Polyvalent cations initiate bridging interactions with the negatively charged groups on the macromolecules resulting in enhanced aggregation (Baalousha, Nur et al. 2013). Kinsinger et al. studied the removal of TiO₂ by flocculation/clarification and concluded that approximately 5 mg/L of this material having a size <450 nm may persist prior to filtration (Kinsinger, Honda et al. 2015). The presence of humic acid reduced the effectiveness of the flocculation process. This study however used initial TiO₂ concentrations of 100 mg/L, which are several orders of magnitude higher than expected concentrations of this material in surface waters. The ENM concentration used is an important consideration as it affects the particle collision efficiency during aggregation.

A limitation of previous studies on ENM removal by flocculation/clarification processes has been the reliance on concentration measurements using techniques such as UV-vis absorption (Hyung and Kim 2009, Zhang, Zhao et al. 2014). Concentration measurements based on UV-vis absorption are perhaps the least accurate technique due to interferences from organic matter and other materials in the water samples. Another challenge is the lack of correlation between mass removal and ENM concentrations. From a risk standpoint, ENM toxicity is generally attributed to surface reactivity (Karakoti, Hench et al. 2006, Levard, Hotze et al. 2012, Kim, Saha et al. 2013), which in turn is a function of the characteristic large specific surface areas (m²/g) of ENMs. In other words, a relatively small ENM mass concentration may present a high surface area and/or particle number concentration. This is an omission from the literature that may have important consequences for the risk analysis of these emerging water contaminants.

The role(s) played by ENM material properties was illustrated in a study by Chalew et al., who determined that TiO₂ was more effectively removed than zinc oxide (ZnO) nanoparticles (Chalew, Ajmani et al. 2013). This was partly attributed to the dissolution of Zn to Zn(II), where the dissolved phase would not be removed during clarification. ENM heteroaggregation in the studied surface water samples was not considered in this work. Wang
et al. studied the removal of TiO₂ (d < 100 nm) by flocculation and clarification using ferric chloride and polyferric sulfate as coagulants (Wang, Qi et al. 2014). Removal was assessed using turbidity measurements and ranged from 0 to 80%. Removal decreased with decreasing alkalinity and/or increasing humic acid concentration. It was concluded that lower nanoparticle removals correlated with coagulant consumption by, and/or stabilization of, the nanoparticles by the humic acid. Of note, turbidity measurements are limited to particles having diameters ≥500 nm and would thus not detect nanoparticles in the finished water. Thus, a substantial fraction of nanoparticles may not have been measured in this study. The stabilizing effect(s) of humic acid have been observed in other studies using C₆₀ (Hyung and Kim 2009) and polyvinyl alcohol coated nAg (Sun, Li et al. 2013).

Surface waters, and to a lesser extent groundwater, are a mixture of dissolved and particulate materials including NOM and suspended inorganic solids like clay. These materials will interact with the ENMs in manners that are determined by the properties of the dispersing media and the interacting surfaces. As previously discussed, NOM can adsorb to the surfaces of ENMs and affect their aggregation behaviors. Particulates can also interact with ENMs to result in heteroaggregation; however, the influence of heteroaggregation on the removal of ENMs during flocculation/clarification processes has largely been ignored in the literature with the focus previously being placed on understanding ENM behavior in simple electrolyte solutions having one or. Nevertheless evidence from various studies points to the importance that heteroaggregation will play in determining ENM fate in natural waters. Praetorius et al. concluded that under favorable attachment conditions (pH 5) TiO₂ ENMs heteroaggregate with the clay (Praetorius, Scheringer et al. 2012). The propensity to heteroaggregate was reduced under unfavorable attachment conditions. These researchers also found that humic acid (>1.0 mg/L) hindered heteroaggregation at pH 8, even at high (100 mM) ionic strengths. Aggregation was reduced as a result of electrosteric repulsion. Conversely, heteroaggregation was observed in the presence of humic acid and divalent salts through bridging interactions. Similar results
were observed by Huynh et al. who studied ENM interactions with hematite (Huynh, McCaffery et al. 2012). Quik et al. determined that irrespective of the core ENM, they will heteroaggregate with clays in natural waters, both in the presence and absence of NOM (Quik, Velzeboer et al. 2014). Similar observations were made by Quik et al. in a study on the behavior of CeO₂ in surface water where 62% of the CeO₂ was removed through heteroaggregation with clay (Quik, Lynch et al. 2010).

Zhang et al. studied the removal of C₆₀ via flocculation with alum (Zhang, Zhao et al. 2014). These researchers determined that charge reversal can occur at high concentrations of alum resulting in a maximum removal of 93%. Note that sweep flocculation was not considered in this study. Furthermore, nanoparticle concentrations were measured using UV-vis adsorption, meaning that their actual values in the clarified water is debatable. Additionally, assuming that the C₆₀ in the decant have an average diameter of 100 nm, are spherical, and a density of 1.63 g/cm³ at a removal of 93% there would still be 14 µg/L of C₆₀ remaining in the decant. This translates to a total nanoparticle surface area of 5.15 cm²/L (10.3 cm²/day for average daily water consumption. Therefore, removal of ENMs by flocculation/clarification processes must be evaluated using both mass and particle counting techniques in order to fully quantify human exposure to these materials.

The overall goal of this study was to improve our understanding of how heteroaggregation with clay and NOM influences the removal of ENMs in flocculation and clarification processes. Also studied, was the release of the ENMs from the coagulated sludges and the conditions favoring such release. This was done using three commercially relevant ENMs, nAg, TiO₂, and CeO₂, having distinct characteristics in order to understand ENM removal as a class of contaminants, rather than as a singular substance. Removal was assessed in terms of mass and particle number concentrations so as to quantify the significance of measured removal efficiencies. Experiments were done at ENM concentrations in the µg/L range in agreement with predicted environmentally relevant values for these materials.
## 2.0 Materials and Methods

### 2.1 Reagents

All salts were acquired from Fisher Scientific (Pittsburg, PA) and were reagent grade or better. Ultrapure water, having a resistivity of 18 mΩ·cm and an average pH=7.5, was used to make all test solutions. Aluminum sulfate (alum) was used as the coagulant in this investigation. A stock solution of alum was made having a concentration of 1 g/L and unaltered pH=2.4. Tests were done using simple and complex electrolyte solutions. Simple electrolyte solutions were made using different concentrations of sodium chloride, where pH was adjusted using hydrochloric acid or sodium hydroxide. The simulated surface water was made to have the following electrolyte composition, which was made to be representative of a fresh surface water system: 64 mg/L MgCl₂, 9 mg/L NaCl, 93 mg/L calcium chloride, and 168 mg/L sodium bicarbonate. The average final pH of the simulated surface water was 7.9±0.2. Bentonite clay (Colloid Environmental Technologies Company, Arlington Heights, IL) and Suwannee River natural organic matter (SRNOM, International Humic Substances Society, MN) were used to evaluate their impact(s) on ENM flocculation. The SRNOM had a carbon content of 52.47% and was supplied as a dry powder.

### 2.2 Engineered Nanomaterials (ENMs)

Three ENMs were evaluated and were selected based on their unique physicochemical characteristics and existing use in commercial products. Titanium dioxide (TiO₂) nanoparticles were acquired from Nanostructured and Amorphous Materials Inc. (Houston, TX). The TiO₂ nanoparticles were 99% anatase and were supplied as a concentrated aqueous mixture (15% TiO₂ w/w) and had a reported diameter between 5 to 30 nm. The TiO₂ was used without further purification. Stock solutions were made using ultrapure water to achieve a TiO₂ concentration of 200 mg/L and a final pH=2.5. Cerium dioxide (CeO₂) nanoparticles (Nanostructured and Amorphous Materials Inc.) were supplied as a dry powder with a reported purity of 99.5% with a reported diameter between 15 to 30 nm. A stock solution was created by dispersing the CeO₂
powder in ultrapure water at pH=2.5 to achieve a concentration of 100 mg/L. A pH of 2.5 was selected for the TiO₂ and CeO₂ stock solutions as both are strongly positively charged at this pH, and thus stable in suspension. The stock solutions were stored in brown glass jars at room temperature (T=23°C) and constantly mixed using a rotary table. Nano-silver (nAg) was acquired from NanoComposix (San Diego, CA) as an aqueous mixture having a concentration of 20 mg/L. The nAg had a citrate coating (citrate-nAg) and a reported mean diameter of 20 nm. The nAg stock solution was stored at 4°C until used. All ENM stock solutions were sonicated in a sonication bath for 10 mins prior to use to break-up any large aggregates.

## 2.3 NOM adsorption onto ENMs

The physical and chemical transformation of the three different ENMs following their interaction with SRNOM was evaluated. NOM carries a net negative charge in most natural aqueous systems as a result of the carboxylic and phenolic moieties that exist along their molecular chains (Summers and Roberts 1988). Each of the ENMs were coated with SRNOM by mixing them in a 10 mM sodium chloride solution at pH 2.5 in order to minimize repulsive electrostatic interactions between the macromolecules and the ENM surfaces. At pH 2.5 the CeO₂ and TiO₂ ENMs were positively charged, while the citrate-nAg was weakly negatively charged (Figure S1). The ENM and SRNOM were mixed in the electrolyte solution at a concentration ratio of 10:1 (NOM:ENM). The mixture was agitated for 30 mins at 50 RPM on a rotary mixer (Reax 2, Heidolph, Elk Grove Village, IL). The mixture was then filtered through a modified polyethersulfone membrane with a molecular weight cutoff (MWCO) of 500 kDa (Spectrum labs, Rancho Dominguez, CA) using a tangential flow filtration apparatus (KrosFlo Research II, Tangential Flow Filtration System, Spectrum Labs, Rancho Dominguez, CA). This MWCO was chosen because the average size of NOM has been reported to be ≤ 500 kDa (Pelekani, Newcombe et al.). During filtration the e SRNOM coated ENMs were concentrated in the retentate stream, while the un-adsorbed SRNOM entered the filtrate stream. The TFF was operated at a transmembrane pressure (TMP) of 0.5 psi and a flowrate of 40 mL/min. The ENM
suspension was filtered until the retentate volume was approximately 3 mL. The retentate was then removed and its volume reconstituted to original volume using unbuffered ultrapure water and adjusted to a pH of 4.

Filtrate samples from the TFF were analyzed using a total organic carbon (TOC) analyzer (InnovOx, GE Power and Water, Boulder, CO) to determine the equilibrium concentration of SRNOM after adsorption, C<sub>e</sub> (mg C/L). This value was then used to calculate the equilibrium concentration of SRNOM adsorbed per unit mass of ENM, q<sub>e</sub> (mg C/mg ENM), according to Eq. 1:

$$q_e = \frac{C_o - C_e}{D} \qquad (1)$$

where C<sub>o</sub> (mg C/L) is the initial concentration of SRNOM measured in terms of total carbon and D (mg ENM/L) is the ENM dosage.

## 2.3 Characterization of Nanoparticle Size, Zeta Potential and Aggregation State

ENM surface charge was characterized in terms of a zeta potential, which was calculated from electrophoretic mobility values that were measured by the laser Doppler technique. Electrophoretic mobility was measured using a ZetaSizer Nano ZS (Malvern Instruments Ltd, Worcestershire, UK). Zeta potential was calculated from electrophoretic mobility using the Smoluchowski approximation (Elimelech, Gregory et al. 1995). The Smoluchowski approximation was valid for the ENMs and conditions studied here as the Henry Function was >> 1 (Tadros 2012). The Henry function is the ratio of the inverse Debye screening length (κ) and the particle radius (r). Electrophoretic mobility was measured as a function of solution ionic strength and pH through automatic titration of the test solution using concentrated solutions of sodium chloride, sodium hydroxide, and/or hydrochloric acid. For the pH titration experiments the test solution was titrated from an approximate solution pH=7.0 to either an acidic (pH=3), or to a basic (pH=9) value. This was done to avoid over diluting the sample and substantial modification of the solution ionic strength.
Nanoparticle size and aggregation were characterized using dynamic light scattering (DLS, ZetaSizer Nano ZS, Malvern Instruments Ltd, Worcestershire, UK) and nanoparticle tracking analysis (NTA, NanoSight NS500, Malvern Instruments Ltd, Worcestershire, UK). The NS500 can characterize particles in a size range of 10 nm to 2000 nm and a concentration range of 10⁶ to 10⁹ particles/mL. The NS500 visualizes individual particles based on the amount of light (λ=532 nm) that they scatter, which is detected by a digital camera. Movement, or Brownian motion, is measured and converted into a sphere equivalent hydrodynamic diameter, $d_H$ from the measured diffusion coefficient using the Stokes-Einstein equation. Size distributions are generated in terms of a particle concentration, which is determined based on video tracking of individual particles. The DLS, which has a measurement range of 0.3 nm to 10 µm, similarly calculates a $d_H$ size distribution using the Stokes-Einstein equation; however, the size distribution is reported in terms of the intensity of scattered light (λ=633 nm) or a number weighted distribution. All background electrolyte solutions were filtered through a syringe filter having a nominal pore size of 0.2-µm. Prior to DLS and NTA measurements, samples were vortex mixed using a Digital Vortex Mixer (Fisher Scientific, Pittsburg, PA) for 10-secs at 3000 RPM and allowed to stabilize under the relevant conditions for 5-mins. All tests were done at a solution temperature of 25°C and in a minimum of triplicate to calculate an average result.

## 2.4 Jar Tests

ENM removal, and transformation, through flocculation and clarification was assessed using jar tests. The jar tester (Phipps Bird, Model PB-700, USA) was equipped with 6 square acrylic reactors having a maximum working volume of 2-L. Unless otherwise noted all jar tests were done with an initial ENM concentration of 0.1 mg/L. This concentration was used as it approximates the predicted concentrations for these materials in natural waters (Nowack, Ranville et al. 2012). Select tests were done using a SRNOM concentration of 1 mg/L and/or a clay concentration of 24.6 mg/L (turbidity=5 NTU). The SRNOM and clay concentrations were
selected to be representative of that found in surface water systems (Crittenden, Trussell et al. 2012). Test solution components, except alum, were added to a square reactor and mixed at 120 RPM for 3-mins. The mixing speed was then increased to 212 RPM (G≈1000 sec⁻¹) at which point the alum was added and mixed for 20-secs. Flocculation was done at a mixing speed of 12 RPM (G≈15 sec⁻¹) for 30-mins. After flocculation, the suspension was subject to clarification for 1-hr. Samples of the clarified water were withdrawn from the upper 3 cms of the reactor and stored in the dark at 4°C until they were analyzed. Jar test conditions were optimized with respect to solution pH and coagulant dose for each ENM, with optimized conditions being those that produced the lowest ENM concentration and/or turbidity in the clarified water. All tests were done at room temperature (T=23±1°C).

## 2.5 Heteroaggregation Kinetics

ENM, SRNOM and clay heteroaggregation kinetics were characterized using time-resolved laser diffraction (MasterSizer 3000, Malvern Instruments Ltd, Worcestershire, UK). The MasterSizer 3000 allows for samples to be recycled from a 200 mL reactor through the measurement cell. Tests incorporated a rapid mix phase followed by a flocculation phase. Rapid mixing of the clay mixture (clay=24.6 mg/L, ENM=18/mg/L, SRNOM=1mg/L), made using the complex salt solution, was done using a rectangular paddle wheel mixer at a speed of 350 RPM for 50-secs. This was done to fully disperse the clay in suspension to facilitate their interaction with subsequently added ENMs. After this phase the ENMs were then added to the clay mixture and mixed at 350 RPM for 50-secs. Flocculation was then done at a mixing speed of 80 RPM for a minimum reaction period of 30-mins. Throughout the entire duration of the tests sample is recycled through the MasterSizer and the reaction cell to characterize the particle size evolution over time. Prior to each test the clay was equilibrated with the electrolyte solution at specified pH for a minimum of 12-hrs to ensure full hydration of the clay.
## 2.8 Measurement of Nanoparticle Concentration

The mass concentrations of ENMs in water samples were measured using ICP-OES in standard mode (Optima 8300, Perkin Elmer, Akron, OH) and ICP-MS using collision mode in the presence of helium gas (Nexion 300X, Perkin Elmer, Akron, OH). Mass concentration values from ICP-OES and ICP-MS analyses were used solely for jar test experiments with different digestion procedures. Three different digestion procedures were used: a water extraction method (ICP-OES), digestion with hydrofluoric, nitric, and sulfuric acid (ICP-MS), and varying digestion procedures optimized for dissolution of each metal/metal oxide (ICP-OES). Only ICP-MS measurements were used for quantitative analyses, while ICP-OES measurements were used for qualitative analyses and trend development. Samples for ICP-MS measurements were prepared using the following digestion procedure. A 3.5 mL sample was mixed with 1.5 mL of 69% nitric acid, 0.5 mL of 40% hydrofluoric acid and 0.05 mL of sulfuric acid in a 7 mL Teflon container. The container was sealed, and heated to 110°C for 24-hrs. After heating, samples were cooled to 65°C and allowed to evaporate. The containers were then filled to a volume of 5 mL with a 2% nitric acid solution and then were analyzed in the ICP-MS. Ultrapure water was used as a blank correction.

Sample preparation for ICP-OES analyses consisted of an initial water extract in a pH range of 6 to 8. Due to the refractory nature of oxides, additional ENM-specific dissolution procedures were used. CeO₂ containing samples were ashed in a crucible at 500°C for 8 hrs. After cooling, 1 ml of 69% nitric acid was added to the ash residue and allowed to sit for 3 hrs to ensure dissolution of the CeO₂. Ultrapure water was added to reconstitute the sample back to the original 10 mL volume. nAg containing samples were prepared by adding 5 mL of 69% nitric acid to 5 ml of the nAg sample. The mixture was mixed for 12 hrs in a glass beaker. The sample was simultaneously heated at 50°C until 1 mL of sample remained. The solution was then cooled to room temperature and reconstituted to a volume of 5 mL using 2% nitric acid. TiO₂ containing samples were digested according to the method outlined elsewhere (Khosravi,
Hoque et al. 2012). TiO₂ containing samples were placed in a crucible and dried to a residue on a hot plate. The residue was covered with 1 g of ammonium persulfate. The crucible was then heated with a Bunsen burner until fuming ceased at which point the TiO₂ was fused with the ammonium persulfate. The fused residue was soaked in 10 mL of 2% nitric acid solution for 30 min, transferred to a beaker, and heated to a slow boil for 10 mins on a hot plate.

Method detection limits (MDL) for ICP-OES were determined by analyzing trace metal grade 2% nitric acid blanks at the respective wavelengths for each ENM ($\lambda_{CeO_2}$ = 418.660 nm; $\lambda_{nAg}$ = 328.068 nm; $\lambda_{TiO_2}$ = 336.121 nm). The blank was analyzed seven times with five replicates per analysis, and a standard deviation was taken of the seven averaged replicates. For this study, based on the mass concentrations observed, an MDL was defined as the average of the seven replicates plus one standard deviation. MDLs for the ICP-MS measurements were approximated by a conservative estimation, which was the theoretical instrument detection limit ×100. Thus, the ICP-MS detection limits for CeO₂, nAg, and TiO₂ were 0.4, 0.1, and 0.1 µg/L, respectively.

## 3.0 RESULTS AND DISCUSSION

### 3.1 ENM Characteristics in Relevant Media

Zeta potentials for each of the studied ENMs are reported as a function of pH in **Figure 1**. Isoelectric points (pH$_{iep}$) of 6.5 and 4.0 were respectively measured for the TiO$_2$ and CeO$_2$, in general agreement with literature reported values (Labille, Feng et al. 2010). Notably, the pH$_{iep}$ for unmodified CeO$_2$ has been reported to be as high as pH 8 to 9 (Pelletier, Suresh et al. 2010). No pH$_{iep}$ was measured for the citrate capped nAg; however, Roh et al. reported a pH$_{iep}$ of 2.7 for this ENM (Roh, Umh et al. 2013). While the specific surface charge characteristics for each ENM were unique, the overall magnitude of the measured zeta potentials for each ENM was similar at pH ≥ 7 at ≤ -20 mV.
<figure><img src="image_2.png" /><figcaption>**Figure 1.** (a) Zeta potential as a function of solution pH for the (a) unmodified and (b) NOM surface modified ENMs. All values are reported at a solution ionic strength of 1 mM NaCl (T = 25°C, n ≥ 3).</figcaption></figure>

The specific adsorption capacities of the unmodified CeO₂ and TiO₂ for NOM were 1.78±0.5 and 2.35±0.1 mg NOM/mg ENM, respectively and were statistically similar (p-value >0.05). That measured for the citrate-nAg was significantly different from those values, and higher, at 5.60±0.04 mg NOM/mg ENM. At the pH used in the adsorption tests (=2.5) the ENMs and NOM carried net positive surface charges so adsorption primarily occurred through hydrophobic and van der Waals interactions. NOM will adsorb to the TiO₂ and CeO₂ surfaces through interactions with phenol and carboxyl groups present on the side chains of the various macromolecules that make up NOM. Adsorption of NOM to the citrate-nAg surface resulted from substitution of the citrate groups by the larger MW fractions of NOM. Citrate is a weakly bound ligand that can be displaced by fulvic and humic acids (Levard, Hotze et al. 2012), which are common components of NOM. The higher specific adsorption for NOM on the citrate-nAg, relative to the other ENMs, was attributed at least in part to the former's higher, by at least an order of magnitude, specific surface area (m²/g). This resulted from the more narrow size distribution and smaller mean particle size that was measured for the citrate-nAg (Figure S1). NOM adsorption resulted in the zeta potentials of all of the ENMs converging to similar values over the studied pH range (Figure 1b).
Coating the ENMs with NOM affected their size distributions in a 10 mM solution at pH 7 did not substantially alter their respective size distributions (Figure S1). Increasing the solution ionic strength to 100 mM did reveal a stabilizing effect on behalf of the SRNOM surface coating (Figure S2). Both the TiO₂ and citrate-nAg demonstrated a reduced tendency to aggregate as the ionic strength was increased. For the TiO₂, the SRNOM surface coating reduced the mean particle diameter from 240 nm to 153 nm, while that for the citrate-nAg decreased from 106 nm to 61 nm in the 100 mM solution. This suggested that the SRNOM stabilized the two ENMs at the relevant solution chemistry in agreement with previous findings (Petosa, Jaisi et al. 2010). No change was observed for the CeO₂ indicating that the SRNOM did not affect the stability of this ENM under the noted solution chemistry.

## 3.1 Flocculation and Clarification of Unmodified ENMs

The flocculation, and subsequent removal via clarification, of each unmodified ENM was investigated with respect to solution pH and coagulant dose. Removal of the CeO₂ was ≥96% independent of the coagulant dose used within a range of 10 to 22 mg/L alum (Table 1). In contrast, the highest removal efficiencies (≥91%) for the citrate-nAg and TiO₂ were achieved at an alum dose of 18 mg/L. Removal decreased for both ENMs at alum doses below, and above, this value (Table 1). Based on the pH and alum concentrations used for the TiO₂ and CeO₂ the principle flocculation mechanism that occurred was sweep flocculation, which is a combination of surface interactions (charge-charge) and physical entrapment. That for the citrate-nAg was a combination of sweep flocculation and adsorption-charge neutralization.
<table><caption>**Table 1.** Removal of unmodified ENMs as a function of pH and coagulant (alum) concentration (T=21°C; n ≥ 3). Coagulant concentration was varied while holding pH at 7.0. Removal was measured based on NTA particle concentration measurements</caption><thead><tr><th>ENM</th><th>Alum Concentration, mg/L</th><th>Removal %</th><th>pH</th><th>Removal %</th></tr></thead><tbody><tr><td rowspan="4">CeO<sub>2</sub></td><td>-</td><td>-</td><td>5.0</td><td>39±12</td></tr><tr><td>10</td><td>96±3</td><td>6±0.1</td><td>43±1</td></tr><tr><td>18</td><td>98±2</td><td>7±0.1</td><td>99±0.1</td></tr><tr><td>22</td><td>99±1</td><td>8±0.1</td><td>100±0.2</td></tr><tr><td rowspan="3">Citrate-nAg</td><td>14</td><td>59±17</td><td>6±0.1</td><td>92±2</td></tr><tr><td>18</td><td>91±3</td><td>7±0.1</td><td>91±3</td></tr><tr><td>22</td><td>86±5</td><td>8±0.1</td><td>16±14</td></tr><tr><td rowspan="4">TiO<sub>2</sub></td><td>10</td><td>89±4</td><td>5±0.1</td><td>0±2</td></tr><tr><td>14</td><td>87±5</td><td>6±0.1</td><td>12±12</td></tr><tr><td>18</td><td>93±3</td><td>7±0.1</td><td>93±3</td></tr><tr><td>22</td><td>80±3</td><td>8±0.1</td><td>88±3</td></tr></tbody></table>

CeO₂ removal decreased from 100% to 43% as the pH decreased from 8 to 5, where a substantial drop-off in removal occurred from pH 6 to 5. Removal of the citrate-nAg decreased as the pH was increased from 6 to 8, ranging from ≥ 91% to approximately 16%. For the TiO₂ similarly high removals were measured at pH ≥ 7, while removal at pH ≤ 6 was negligible. In a pH range of 6 to 8 all of the ENMs carried a net negative surface charge of approximately the same magnitude (Figure 1) suggesting that all would be affected similarly by adsorption-charge neutralization. Note, that even an effective removal, >90%, there was still a large number concentration of ENMs in the clarified water. For the optimized conditions (pH and alum concentration), there were on average 2.8×10⁵, 2.6×10⁸, and 9.5×10⁷ particles/mL for CeO₂, citrate-nAg, and TiO₂ in the decant, respectively. This amounts to a surface area of 0.18 cm²/L for CeO₂ (primary peak=145 nm), 320 cm²/L for citrate-nAg (principal peak=200 nm) and the surface area for TiO₂ was 47 cm²/L (primary peak=125 nm). The importance of these area values arises from the ability of nanoparticles to adsorb, as a result of their characteristically high specific surface area values, and subsequently serve as transport vectors for, other aqueous contaminants. The question of surface area is important because many toxicological
studies are dependent on surface area which can increase the production of reactive oxygen
species and cellular uptake which may result in cytotoxicity (Carlson, Hussain et al. 2008). In
comparing the removals for the three ENMs, the densities of the constituent particles must be
considered as this affects the terminal settling velocity of the floc, which were 7.1, 2.1, and 3.9
g/cm³ for the CeO₂, citrate-nAg, and TiO₂, respectively. Neglecting the actual densities of the
flocs themselves we would expect the CeO₂ to be the most well-removed ENM and the citrate-
Ag to be the least.

## 3.2 ENM Heteroaggregation with Clay and NOM

Flocculation tests for the clay-ENM-NOM mixtures were done at different pH values corresponding to the optimum pH that was determined for each system. In the clay suspension, each of the ENMs were characterized by similar aggregation rates (8 to 11 µm/min) and final aggregate sizes ($D_{50}$=100 to 160 µm) (Figure 2). ENM addition to the clay suspension did not alter the clay aggregation rate, or $D_{50}$, irrespective of pH. When 1.0 mg/L of NOM was added to the mixture, the aggregation rates and $D_{50}$ values were reduced for the citrate-nAg (19 to 1.5 µm/min, pH 6) and TiO₂ (11 to 1.8 µm/min, pH 7) mixtures; however, the aggregation rate for the CeO₂ mixture was relatively unchanged (8.1 to 7.6 µm/min, pH 8). This was attributed to preferential adsorption of the NOM to the clay at pH≤7. Clay is comprised of flat plate-like structures, where the faces have a constant net negative charge, while the edges are amphoteric (pH<sub>lep</sub>=7). At pH≤7 the clay edges have a net positive charge, which can serve as preferential adsorption sites for the negatively charged sites on the NOM. Attachment of NOM to ENMs has been shown to reduce the favorability of aggregation in various mixtures (Petosa, Jaisi et al. 2010). There was no decrease in the heteroaggregation rate or $D_{50}$ of the CeO₂ mixture at pH 8 because all constituents were negatively charged resulting in unfavorable attachment conditions. Similar results on the behavior of clay and ENMs were observed by Praetorius et al. who observed reduced heteroaggregation as inter-particle charge interactions became more substantial (Praetorius, Scheringer et al. 2012). It was also found that the humic
acid only partially slow the heteroaggregation between SiO₂ and TiO₂ since the clay, humic and TiO₂ were all negatively charged. From these results, the SRNOM increased the stability of the clay at pH>6. At environmentally relevant ENM concentrations (0.1 mg/L or an ENM:clay concentration ratio=0.41%), clay heteroaggregation with the ENMs is not measurably enhanced.

<figure><img src="image_3.png" /><figcaption>**Figure 2.** (a) Aggregation rate and (b) final mean aggregate size ($D_{50}$) of clay in different mixtures of ENMs, SRNOM, and aluminum sulfate in the complex salt solution.</figcaption></figure>

**3.3 ENM Heteroaggregate Removal through Flocculation and Clarification**
The addition of alum increased the aggregation rate and D₅₀ substantially for the CeO₂ and TiO₂ mixtures; however, no changes were observed for the citrate-nAg relative to that measured for the ENM/clay system (Figure 2b). The aggregation rate of CeO₂ mixture increased from 7.9 to 319 µm/cm and the D₅₀ increased from 130 µm to 550 µm. Similarly, the TiO₂ mixture aggregation rate increased from 10.4 to 176 µm/min and the D₅₀ increased from 110 to 580 µm. Conversely, the aggregation rate and D₅₀ for the citrate-nAg with alum addition (aggregation = 7.6 µm/cm, D₅₀=150 µm), which was comparable to those values for ENM-clay mixture (aggregation=11.2 µm/cm, D₅₀ =160 µm). This however, was greater than the aggregation rate and final D₅₀ when there was SRNOM in the solution (aggregation=1.3 µm/cm, D₅₀ =50 µm). This variation in final aggregate size was not a surface charge magnitude phenomenon because the differences in charge was minimal between the ENMs, CeO₂ ζ ≈ |20 mV|, nAg ζ ≈ |16 mV|, TiO₂ ζ ≈ |12 mV|. Rather, the difference was attributed to the difference in available surface area, which was different for CeO₂, citrate-nAg, and TiO₂. The dependency of coagulating salts on the surface area and surface charge density of colloids, particularly through charge neutralization has been observed previously (Dentel 1988). Hence, the aggregation rate and final aggregate D₅₀ is much lower for citrate-nAg because of the abundant surface area. Recall, the size distribution of the ENMs (Figure S1), where citrate-nAg was substantially more dispersed than the TiO₂ and CeO₂. As the number concentration increases, the surface area will also increase. Therefore, calculating the surface area based on the number concentration and the most prominent peak results in surface areas (cm²/L) for CeO₂ (Figure 3a), citrate-nAg (Figure 4a), and TiO₂ (Figure 5a) were 3.3, 25, and 2.0 cm²/L, respectively. There was ~10x more surface area of citrate-nAg than TiO₂ and CeO₂ played a significant role in aggregation rate and final aggregate size.

ENM removal during the flocculation/clarification process for the complex salt/clay/SRNOM mixture, was statistically the same (p-value>0.05) for all three materials when using the associated optimized conditions; however, differences between the ENMS were
observed. Based on the ICP-MS analyses, average removal was lowest for the citrate-nAg (84±13%) and highest for the TiO₂ (96±5%). That for the CeO₂ was 89±19%. Reductions in turbidity for all tests were ≥96%, with no statistically significant differences between any of the tests.

ENM removal through flocculation and clarification followed conventional theory, where removal occurred via enmeshed in the coagulant floc and/or heteroaggregation with the clay (Chalew, Ajmani et al. 2013). Also, removal occurred by adsorption of positively charged hydrolytic species (coagulant) that neutralizes the surface charge through adsorption-charge neutralization mechanisms (Petosa, Jaisi et al. 2010, Chalew, Ajmani et al. 2013). These observations were supported by the static light scattering measurements, where the D₅₀ for all ENM mixtures, regardless of solution chemistry, was ≥30 µm. Similar removal efficiencies were observed by Chalew et al. (R≈60-100%) and Hyung et al. (R≈60%), though specific ENM size fractions were not reported in either study (Hyung, Fortner et al. 2007, Chalew, Ajmani et al. 2013). An outstanding, and important, question was thus the particle concentrations that persisted in the clarified waters.

In a suspension of CeO₂ and clay only, the CeO₂ largely accumulated in the sludge zone following clarification. This was due to heteroaggregation with the clay and/or settling of homoaggregates. Approximately 15% of the initial CeO₂ mass remained in the clarified water with a fraction of that being smaller than 100 nm (Figure 3b). The addition of SRNOM to the mixture reduced the amount of CeO₂ entering the sludge zone, consequently increasing the amount of ENM in the clarified water; however, the nano-fraction remained unchanged. Alum addition marginally altered the removal of the CeO₂ to the sludge zone with approximately 30% of the CeO₂ still present in the clarified water. Similar observations were made for the citrate-nAg. When just citrate-nAg and clay were in the mixture there was 25% of nanoparticles remaining in the clarified water (Figure 4b), but the concentration of ENMs smaller than 100 nm was very low (~1 µg/L). When SRNOM was added to the solution the amount of nAg that
remained nano-sized increased from 1 µg/L to 9 µg/L. This was expected because of heteroaggregation, where the aggregation rates and final aggregation sizes decreased for the nAg solution because the NOM stabilizes the nanoparticles and the clay. Because of an increased amount of nanoparticles in the clarified water, the removal decreased from 74% to 62% (Figure 33). However, similar to CeO₂ once the alum was added the amount of ENMs remaining in the clarified water decreased and the amount that remained nano-sized approached zero. Heteroaggregation between the clay and the nanoparticles and homoaggregation of the nanoparticles will cause increases in particle size and subsequent removal during clarification. The third particle, TiO₂ had dissimilar observations. Of the three solutions, TiO₂ and clay was the only test that had any nanomaterials remaining in the clarified water (2 µg/L), however this concentration was not nano-sized. The latter two solutions (ENM, clay, and NOM, and ENM, clay, NOM, and alum) resulted in no fraction of nanoparticles in the decant (Figure 34). The concentration of the colloidal fraction (>220 nm) and the nano-sized fraction approached zero in both test solutions. Again, these removals are likely due to homo and heteroaggregation of the TiO₂ nanoparticles with the clay materials.
<figure><img src="image_4.png" /></figure>

<figure><img src="image_5.png" /><figcaption>**Figure 3.** (a) Size distributions of CeO₂ measured with NTA of pre- and post-clarification using optimal flocculation conditions (n=3, IS = 5 mMol complex salt; 24.6 mg/L bentonite, and 1.0 mg/L SRNOM). (b) Mass concentrations of cerium (Ce) following clarification in the clarified water and sludge.</figcaption></figure>

The concentration and location of the nanoparticles in the solution correlated positively with heteroaggregation and NTA analysis. Thus, the amount of ENMs remaining in the clarified water was small, but further, the amount of ENMs that remained nano-sized approached zero for all three ENMs. From the heteroaggregation studies, the ENMS were expected to interact with the SRNOM and clay during flocculation, thus affecting their flocculation efficiency. There
was further support of the low concentration of ENMs in the clarified water by the NTA analysis.
With CeO₂, there were particles smaller than 100 nm (Figure 3a), but these particles may have
been heteroaggregates of NOM, alum species, dispersed clay, CeO₂, or any amalgamation of
the four. This was similarly observed for the citrate-nAg (Figure 4a) and TiO₂ (Figure 5a),
where the concentration of particles smaller than 100 nm was extremely small. It can be
concluded from heteroaggregation and removal analysis that the mass concentration of ENMs
that are nano-sized that will pass to clarification will be small (<10 µg/L).

<figure><img src="image_6.png" /></figure>

<figure><img src="image_7.png" /><figcaption>**Figure 4.** (a) Size distributions of citrate-nAg measured with NTA of pre- and post-clarification using optimal flocculation conditions (n=3, IS = 5 mMol complex salt; 24.6 mg/L bentonite, and</figcaption></figure>
1.0 mg/L SRNOM). (b) Mass concentrations of silver (Ag) following clarification in the clarified water and sludge.

<figure><img src="image_8.png" /></figure>

<figure><img src="image_9.png" /></figure>

**Figure 5.** (a) Size distributions of TiO₂ measured with NTA of pre- and post-clarification using optimal flocculation conditions (n=3, IS = 5 mMol complex salt; 24.6 mg/L bentonite, and 1.0 mg/L SRNOM). (b) Mass concentrations of titanium (Ti) following clarification in the clarified water and sludge.

**3.4 ENM Redispersion from Coagulated Sludges**

In conventional water treatment facilities coagulated sludges are collected from the clarification process and ultimately disposed of through land application and/or landfilling. An
important question then is the degree to which ENMs are fixed in the sludges, as these sludges
may over time serve as a point source for ENM pollution. To study this question the coagulated
sludges from jar tests for the different ENMs were subjected to two different conditions. In the
first condition, the sludge was collected and then filtered through a membrane having a nominal
pore size of 0.2 µm. This condition simulated physical disruption of the sludge. In the second
condition, the sludge was dried and then exposed to water of varying pH values. This condition
represented chemical disruption of the sludge structure. Application of the pH solution was done
so that the sludge was not physically disrupted. Samples were collected after 72-hrs of
exposure to the pH solution.

Physical disruption of the coagulated sludges showed no significant leaching of the TiO₂
or CeO₂; however, there was leaching of the citrate-nAg based on ICP-OES analysis (Figure 6).
Because the ICP-OES is detecting silver it is possible that what is being measured is a
combination of ionic silver and nAg. Adjusting the pH of the sludge prior to the filtration step
affected the amount of silver that was released, increasing it from approximately 1 µg/L to 5
µg/L. Similarly, chemical disruption of the dried sludges only resulted in the resuspension of nAg
irrespective of the solution pH. Like the physical disruption measurements, the highest
concentration of silver was measured when the sludge was exposed to pH 7 (Figure 6). Though
statistically different, the concentration measured at pH 3 for the chemical disruption tests was
comparable. This suggests that the release of the silver is not dictated by the charge character
of the citrate-nAg or the SRNOM-citrate-nAg alone, as both particles are negatively charged at
both pH values. Similarly, greater silver release was not observed when the pH was adjusted to
9 in either tests, where the coated nAg was most strongly negatively charged. Therefore, the
reaction within the sludge matrix is complex and but hinges on the unique interfacial interaction
conditions, chemically and geometrically based on the size of the nAg core particles, presented
by the nAg as compared to the TiO₂ and CeO₂.
<figure><img src="image_10.png" /><figcaption>**Figure 6.** Silver (Ag) concentrations following physical and chemical disruption of the coagulated sludge consisting of alum, clay, SRNOM, and citrate-nAg (n=3). The sludge was coagulated using optimum jar test conditions for citrate-nAg conditions.</figcaption></figure>

Floc strength is dependent upon the inter-particle bonds between each of the components of the aggregate (Bache, Johnson et al. 1997). The fractal nature of an agglomerate can be a good indicator of floc strength. Aggregates with dense agglomerate structures (high fractal dimension) are stronger than those with ramified structures (low fractal dimension) (Li, Zhu et al. 2007). In a diffusion limited aggregation regime particles rapidly form aggregates, resulting in low fractal dimension of branched structure. A low fractal dimension is the measure of a ramified agglomerate structure. Conversely in the reaction limited aggregation regime the particles formed aggregates characterized by a higher fractal dimension, implying a more dense aggregate structure. Aggregates with a low fractal dimension are directly proportional to lower agglomerate strength factors and are more prone to shear forces, such as those applied during the physical disruption tests. Conversely aggregates with a high fractal dimension are more prone to slow surface erosion (Li and Logan 2001). Based on the ENM erosion measurements done here, the TiO₂ and CeO₂ were not prone to erosion through shear and/or chemically induced erosion. This would imply a high fractal dimension for the associated
floc structures. The nAg on the other hand was susceptible to chemical erosion implying a lower fractal structure. Further consideration of this question is warranted to better understanding the potential of ENM containing sludges to serve as future point sources for these materials.

**4.0 Conclusions**

The ENMs studied here were effectively, based on mass concentrations, removed through flocculation/clarification using alum as a coagulant at determined optimum pH values. Removal measured in terms of particle, and subsequently surface area, concentrations revealed the presence of substantial ENMs in the finished waters when starting with environmentally relevant concentrations (~µg/L range). Therefore, previously reported removals for ENMs based on conventional measures of particle concentration, like UV-vis absorption or turbidity, are likely incomplete. This suggests that risk analyses for ENM exposure through ingestion of potable water should be revisited. Only nAg displayed a tendency to be released from the coagulated sludges through physical or chemical erosion processes. It is therefore possible that residuals from water treatment facilities removing ENMs may serve as point sources of ENM pollution. This will be a function of ENM characteristics, e.g., size, density, surface coating, and the nature of the floc formed.

**5.0 Acknowledgements**

This study was supported by the National Science Foundation under Grant No. 1134231. Any conclusions, findings, recommendations, or suggestions expressed in this paper are those of the authors and may not reflect those of the National Science Foundation. Funding was also provided by the Partner University Fund and Centre Européen de Recherche et d'Enseignement des Géosciences de l'Environnement, Aix en Provence, France.

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