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Nanoscale titanium dioxide (nTiO) transport in natural sediments: Importance of soil organic matter and Fe/Al-oxyhydroxides Leanne M. Fisher-Power, and Tao Cheng Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.7b05062 • Publication Date (Web): 27 Jan 2018 Downloaded from http://pubs.acs.org on January 28, 2018

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Environmental Science & Technology

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Nanoscale titanium dioxide (nTiO2) transport in natural sediments:

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Importance of soil organic matter and Fe/Al-oxyhydroxides

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Leanne M. Fisher-Power & Tao Cheng *

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Department of Earth Sciences, Memorial University St. John’s, Newfoundland and Labrador, A1B 3X5, Canada

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Corresponding author: [email protected]

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Abstract

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Many engineered nanoparticle (ENP) transport experiments use quartz sand as the transport

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media; however, sediments are complex in nature, with heterogeneous compositions that may

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influence transport. Nanoscale titanium dioxide (nTiO2) transport in water-saturated columns of

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quartz sand and variations of a natural sediment was studied, with the objective of understanding

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the influence of soil organic matter (SOM) and Fe/Al-oxyhydroxides and identifying the

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underlying mechanisms. Results indicated nTiO2 transport was strongly influenced by pH and

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sediment composition. When influent pH=5, nTiO2 transport was low, as positively-charged

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nTiO2 was attracted to negatively-charged minerals and SOM. nTiO2 transport was slightly

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enhanced in sediments with sufficient SOM concentrations due to leached dissolved organic

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matter (DOM), which adsorbed onto nTiO2 surface, reversing zeta potential to negative. When

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influent pH=9, nTiO2 transport was generally high since negatively-charged medium repelled

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negatively-charged nTiO2. However, in sediments with SOM or amorphous Fe/Al-

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oxyhydroxides depleted, transport was low due to pH buffering by the sediments causing

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attraction between nTiO2 and crystalline Fe-oxyhydroxides. This was counteracted by DOM

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adsorbing to nTiO2, stabilizing it in suspension. Our research demonstrates the importance of

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SOM and Fe/Al-oxyhydroxides in governing ENP transport in natural sediments.

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Abstract Art

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1. Introduction

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The widespread use of engineered nanoparticles (ENPs) has led to their release into the

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environment with limited knowledge about their transport and associated risks. A common way

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to study ENP transport is laboratory column experiments designed to simulate groundwater flow

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through saturated porous media 1–8. Most of these experiments use well-defined porous media,

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often quartz sand, in order to distinguish the influence of water chemistry (pH, ionic strength,

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cations and anions, etc.) on ENP transport. Due to technical difficulties in experiments 9, a

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limited number of studies have been conducted to investigate ENP transport in natural sediments,

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which are far more complicated compared to quartz sand. Sediment properties including

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composition, surface charge, and surface area are known to influence ENP transport 10–13. Fang

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et al. (2011) studied the stability of nTiO2 suspensions in natural soil and found that stability was

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positively correlated to dissolved organic carbon (DOC) and clay content, but negatively

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correlated to ionic strength, pH, and zeta potential (ZP) 14. Sun et al. (2015) demonstrated that

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zinc oxide nanoparticles were highly mobile in quartz sand columns under various water

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chemistry conditions, but less mobile in soil columns due to heterogeneous ZPs, blocking, and

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straining 15. Other studies used natural sediments to study ENP facilitated transport of

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contaminants 1,3,16,17. Though these studies contribute useful information of ENP transport in

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natural sediment, the roles of specific sediment components and associated mechanisms have not

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fully been identified, leaving knowledge gaps regarding ENP interaction with heterogeneous soil

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components 15,18.

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Common ubiquitous components of natural sediment that may influence transport include

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soil organic matter (SOM) and metal oxyhydroxides (Fe and Al). SOM is composed of both solid

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soil organic matter (SSOM) and dissolved organic matter (DOM) 19. While a large number of

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studies have shown that DOM enhances ENP stability and transport through adsorbing onto the

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ENP surface and altering surface charges through steric repulsion 20–24, little attention has been

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paid to SSOM-ENP interactions. It is intuitive to hypothesize that negatively-charged SSOM 25

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may attract positively-charged ENPs and remove them from the aqueous phase. Moreover,

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partial dissolution of SSOM releases DOM that may interact with ENPs. The overall effect of

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SOM (both solid and dissolved) in natural sediments on ENP transport is complicated and poorly

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understood.

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Metal oxyhydroxides have a complex effect on ENP transport, as heterogeneous surface

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charges in natural sediments arise in part by metal oxide components. For example, quartz is

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negatively-charged in environmental conditions with a point of zero charge (PZC) around pH 2-

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3, Fe-oxyhydroxides have a PZC in the range of 8.1-8.5, and Al-oxyhydroxides have a PZC

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around 9.4 26. At variable pH, a complicated natural sediment will have an array of positive and

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negative surface charges, which will cause varying components to have different interactions

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with ENPs. Johnson et al. (1996) found increasing the percentage of Fe-oxyhydroxide coated

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surfaces increased deposition of silica microspheres 27, and Cornelis et al. (2013) found

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extractable Al in sediments was correlated with increasing attachment efficiency of silver ENPs

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ENP transport 28,29; however, coatings on quartz sand are artificial and may not behave the same

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as Fe-oxide coatings in natural sediments. Additionally, SOM and Fe/Al-oxyhydroxides often

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co-exist in natural sediments, therefore, research is needed to understand not only their

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individual effect, but also their synergistic effect when both are present in one heterogeneous

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system.

. Numerous other studies demonstrate the importance of Fe-oxide coating on quartz sand in


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The objective of this study was to examine the influence of SOM and Fe/Al-oxyhydroxides

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on ENP transport in a natural sediment and identify the underlying mechanisms. A natural

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sediment was separated into portions, treated to deplete specific components (Fe/Al-

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oxyhydroxides, SOM, or both), and used in systematic laboratory column experiments to

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determine the influence of sediment component(s) on ENP transport using nanoscale titanium

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dioxide (nTiO2) as a representative ENP. By comparing nTiO2 breakthrough curves and changes

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in water chemistry and particle property in different column experiments, mechanisms of how

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SOM and Fe/Al-oxyhydroxides influence particles retention and transport were proposed.

2. Materials and Methods 2.1 Transport Media The sediment was a diamicton from central Newfoundland, Canada; used and

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characterized in previous work 30,31. This sediment was chosen to represent aquifer materials as

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found in natural environments. It was glacial in origin, mainly composed of plagioclase (36%),

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quartz (31%), alkali feldspar (13%), and a variety of other silicates (Figure S1 panel A, SI). The

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sediment also contained important minor components including SOM as well as crystalline and

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amorphous metal oxyhydroxides. Before being used in any experiments, the sediment was air

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dried, sieved to less than 2 mm, homogenized, and separated into four representative portions: to

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be (1) untreated, (2) treated to deplete SOM using hydrogen peroxide (H2O2) based on a

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procedure by Jackson (1958)32, (3) treated to deplete amorphous metal oxyhydroxides using a

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modified ammonium oxalate under darkness (AOD) method 33, and (4) treated to deplete both

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SOM and metal oxyhydroxides using AOD extraction followed by H2O2 extraction. Details of

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the treatment processes are described in the Supporting Information (SI). 4 ACS Paragon Plus Environment


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After processing, each sediment was characterized to elucidate their properties and

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differences resulting from the treatments. Modal mineralogy from X-ray diffraction (XRD), pH,

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grain size distribution and ZP were determined; as well as the quantification of extractable Fe

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and Al oxides; soil organic carbon (SOC) (used to represent SOM); and leachable DOC (used to

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represent leachable DOM). Select characterizations of the untreated sediment were completed

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previously30,31, and additional characterizations were completed in this study, with detailed

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procedures in the SI.

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Additionally, quartz sand (0.252-0.355 mm diameter) from Univar Canada was used as

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control porous media, as nTiO2 transport in pure quartz columns is well documented. To remove

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impurities, the sand was soaked in 0.1 mol/L NaOH for 24 hours, followed by 24 hours in 0.1

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mol/L HCl, and rinsed with deionized water until the absorbance of the water mixed with sand

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was zero.

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2.2 Column Experiments

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To determine the influence of the natural sediment and individual sediment components

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on nTiO2 transport, systematic column experiments were conducted using quartz sand and each

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variation of the sediment as the transport medium (summary of experimental conditions in Table

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S1, SI). Sediments were packed into a column at a 1:4 sediment to quartz sand mass ratio, based

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on practical considerations (maintaining sufficient hydraulic conductivity and ensuring

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particulates were low as to not clog column stoppers). The sediments and quartz sand were

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mixed uniformly by shaking the dry materials before packing and mixing while packing the

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saturated columns.

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A Kimble Kontes Chromaflex ® chromatography column (2.5 cm inner diameter and 15 cm length) made of borosilicate glass was used to hold the transport medium. The column was 5 ACS Paragon Plus Environment

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fitted with 20 µm high-density polyethylene bed supports on each end and configured vertically,

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with Cole-Parmer Masterflex ® Tygon 0.8 mm tubing connected to either side. For each

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experiment, the column was packed with sediment and/or quartz sand, and nTiO2-free 1 mM

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NaCl background solution adjusted to pH 5 or 9. Small volumes of background solution

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followed by sediment/quartz sand were added incrementally until the entire column was full,

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tapping and packing to remove air bubbles and ensure uniform packing. The volume of

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background solution added to saturate the column and dry mass of porous media used was

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recorded, allowing for calculations of porosity and bulk density.

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Aeroxide® P25 titanium (IV) oxide powder from Arcos Organics with minimum 99.5%

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purity was used to prepare the nTiO2 suspensions. Previous analysis determined the powder was

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90% anatase and 10% rutile 23. For all experiments, 50 mg/L nTiO2 suspensions were prepared in

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1 mM NaCl in nanopure water and sonicated by Branson Digital Sonifier® (Crystal Electronics)

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at 200W, 30% amplitude, for 30 minutes. The pH of nTiO2 suspensions were adjusted to pH 5 or

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9 after sonication using small volumes of 1.0 mol/L NaOH or HCl solution.

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Prior to nTiO2 injection, 1 mM NaCl background solution with pH 5 or 9 (to match the

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pH of the influent suspensions) was injected for approximately 16 hours at a rate of 1 mL/min

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using a Masterflex® peristaltic pump from Cole-Parmer to ensure that electrical conductivity,

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light absorbance, and pH of the effluent became stable. 50 mg/L nTiO2 suspension was then

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injected for 7 pore volumes before influent was switched back to nTiO2-free background

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solution. The column experiments were completed when no nTiO2 was detectable in the effluent.

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During particle injection and elution stages, effluent samples were collected using a

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Spectrum® Labs C-F2 Fraction Collector and monitored for pH and electrical conductivity.

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Light absorbance of the samples were analyzed by a Thermo Scientific GENESYS™ 10S UV6 ACS Paragon Plus Environment


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Vis Spectrophotometer at wavelength 368 nm, and nTiO2 concentration was determined using a

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calibration curve generated by diluting the 50 mg/L nTiO2 suspension. At the midpoint of nTiO2

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injection (~3.5 pore volumes) effluent samples were collected for ZP and hydrodynamic

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diameter (HDD) analysis by a Malvern Nano ZS Zetasizer. DOC concentration of column

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effluents was analyzed by an OI Analytical Aurora 1030 Total Organic Carbon Analyser. nTiO2

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recovery in each experiment was calculated as the ratio of the sum of nTiO2 mass in the effluent

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samples and the total injected nTiO2 mass. Influent pH was checked over time (adjusting as

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necessary to keep stable at pH 5 or 9), and light absorbance of influent suspensions was

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monitored to ensure particle stability. All column experiments were conducted in triplicate.

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3. Results & Discussion

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3.1 Sediment properties

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The treatment processes were successful in removing their target components (Fe/Al-

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oxyhydroxides, SOM, or both) (Figure 1) while generally preserving the sediment’s other

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properties (Table S2, SI). Fe/Al-oxyhydroxide concentrations varied according to sediment

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treatment, with the greatest amounts in the untreated sediment (Figure 1a and 1b). Processing by

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AOD and AOD-H2O2 to deplete metals was 96-97% effective in removing AOD extractable Al

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and 81-88% effective for Fe; however, treatment by H2O2 for SOM depletion also removed 42

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and 62% Al and Fe respectively (Figure 1a). This was due to degradation and removal of Fe/Al-

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oxyhydroxides naturally associated with SOM in the sediment during the H2O2 procedure 34,35.

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Dithionite-citrate-bicarbonate (DCB) extractable Fe was less affected by the treatment processes,

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as it includes some crystalline Fe (not extractable by AOD) in addition to amorphous Fe 36, with

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74-75% removal for AOD and AOD-H2O2 treated sediments and 48% for H2O2 treated sediment 7 ACS Paragon Plus Environment

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(Figure 1b). Again, Fe depletion for the H2O2 treated sediment was a result of Fe-oxyhydroxides

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associated with SOM being removed. SOC concentration in the sediments also changed

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according to treatment type (Figure 1c). Treatments that targeted SOM (H2O2 and AOD-H2O2)

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removed most of the sediment’s original SOC, reducing by 90-94%; however, the AOD

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treatment (which targeted metal oxyhydroxides) also removed a substantial percentage (79%) of

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SOC (Figure 1c). This was a result of leaching and degradation of SOC associated with Fe/Al-

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oxyhydroxides in the sediment, which were consequently removed during the treatment process

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30,34,35

.



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Figure 1. Proportions of Al, Fe, and SOC in the sediments.

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Despite changes in Fe/Al-oxyhydroxide and SOC concentrations, XRD spectra showed

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that mineralogy of each treated sediment did not greatly differ from the untreated sediment,

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indicating that bulk mineralogy was not significantly affected by the treatment processes (Figure

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S1, SI). The pH of the sediments was not strongly influenced by the treatments (ranging from

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5.74 to 5.86 for the untreated, AOD treated, and H2O2 treated sediments) except for the AOD-

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H2O2 treated sediment which increased to 6.35 (Table S2, SI). Grain size distribution changed

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slightly for the AOD and AOD-H2O2 treated sediments; the treated sediments had a higher

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proportion of sand-sized particles but lower proportion of clay and silt relative to the untreated

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sediment (Table S2, SI). A larger change was observed for the H2O2 treated sediment grain size

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distribution, as the proportion of sand size particles increased and fine particles decreased (Table

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S2, SI), presumably due to the breakdown and loss of fine particles during treatment.

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Quartz sand columns had 40% porosity, whereas mixed sediment and sand columns had

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slightly less (36-37%) due to grain size variation (Table S2, SI). Similarly, bulk density was 1.73

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g/cm3 for quartz sand, and slightly greater for mixed columns (1.77-1.83 g/cm3) (Table S2, SI).

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Pore space was greater in quartz sand columns, due to increased uniformity compared to mixed

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columns with variable grain sizes; whereas specific surface area (SSA) and surface roughness of

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quartz sand was much less than that of the natural sediment 23,37. Differences in physical

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properties between quartz sand and mixed sediment/sand columns (all of which suggest

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enhanced transport in quartz sand columns) needed to be considered in the interpretation of

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nTiO2 transport behaviour.

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3.2 nTiO2 transport nTiO2 breakthrough curves were determined a function of pH and porous medium in the

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column experiments (Figure 2; Figure S2, SI with error bars). Effluent nTiO2 concentrations

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were low for all experiments with influent pH 5; and typically much higher for experiments with

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influent pH 9 (Figure 2). Columns containing only quartz sand had 0% recovery of nTiO2 for

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experiments with influent pH 5 (Figure 2a); whereas almost all of the injected nTiO2 (96%) was

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recovered for experiments with influent pH 9 (Figure 2b). Breakthrough patterns for the

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sediments at influent pH 5 varied according to sediment type (Figure 2a), with the untreated

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sediment having the highest effluent recovery (~3.6%). AOD treated sediment had significant

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but low nTiO2 recovery (~1%) at influent pH 5, while H2O2 and AOD-H2O2 treated sediments

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had zero or near zero nTiO2 recovery, similar to the quartz sand (Figure 2a).

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Figure 2. nTiO2 breakthrough curves for experiments with influent pH 5 and 9. C/C0 is the ratio

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of recovered nTiO2 concentration to nTiO2 concentration in the influent.

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nTiO2 transport was also dependent on sediment type for the pH 9 influent experiments (Figure 2b). Two of the sediments had high nTiO2 recovery; the AOD-H2O2 treated sediment 11 ACS Paragon Plus Environment

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with ~93%, and the untreated natural sediment with ~80%. In contrast, low recovery of nTiO2

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was observed for the AOD and H2O2 treated sediments; the AOD treated sediment had ~6%

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recovery, whereas no nTiO2 was recovered from the H2O2 treated sediment (Figure 2b).

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Blocking shapes (characteristic of deposited nTiO2 preventing further deposition as collector

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surfaces become covered 27,38–40) were observed for the untreated sediment and for the AOD-

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H2O2 treated sediment to a greater extent, as less deposition sites were available due to the AOD

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and H2O2 treatment processes.

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3.3 Zeta potential of nTiO2 and sediments

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The zero recovery of nTiO2 in quartz sand at pH 5 and nearly 100% recovery at pH 9 can be

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explained by comparing the ZP of nTiO2 and quartz at pH 5 and 9 (Figure 3). At pH 5, nTiO2

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was positively-charged and quartz sand had negative ZP (Figure 3), leading to electrostatic

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attraction and zero recovery in column effluents (Figure 2a). Conversely, at pH 9 nTiO2 was

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negatively-charged as was quartz sand (Figure 3), and therefore electrostatic repulsion occurred,

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leading to nearly 100% recovery (Figure 2b). This logic also partially explained nTiO2 transport

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behaviour in the sediments; however, it disagrees with some of the experimental results. As

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expected, positive nTiO2 at pH 5 was attracted to the untreated, AOD treated, and AOD- H2O2

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treated sediments with negative ZPs (Figure 3), accounting for the low recoveries (Figure 2a);

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and negative nTiO2 at pH 9 was repelled by the untreated and AOD-H2O2 treated sediments with

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negative ZPs (Figure 3), agreeing with the high recoveries (Figure 2b). In contrast, the H2O2

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treated sediment at pH 5 had a positive ZP suggesting it should repel positive nTiO2 (Figure 3),

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however, no nTiO2 was recovered from the H2O2 treated sediment columns at pH 5 (Figure 2a).

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Additionally, the negative ZP of the AOD and H2O2 treated sediments at pH 9 suggested that

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negatively-charged nTiO2 should be repelled, leading to high recovery (Figure 3), however, low 12 ACS Paragon Plus Environment


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nTiO2 recovery was observed in these experiments (Figure 2b). A limitation of the sediment ZP

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measurement was that it may not be representative of all the surface sites in the sediment,

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tending to disproportionally favour certain components. For example, amorphous Al and Fe,

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clays, and DOM were more likely to be mobilized and represented by these measurements in

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comparison to larger particles in the sediment, such as quartz and other minerals. The H2O2

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treated sediment leached a higher proportion of positively-charged Fe/Al-oxyhydroxides relative

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to negatively-charged DOM due to SOM depletion, accounting for the reported positive ZP at

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pH 5. Despite this positive ZP, the H2O2 treated sediment also contained negatively-charged

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quartz and other minerals which were less represented by the ZP measurement, and attracted

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nTiO2 accounting for the nil recovery.

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Figure 3. Zeta potential of quartz sand, sediments, and nTiO2 at pH 5 and 9. Zeta potential of quartz sand determined previously by Wu and Cheng (2016).

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Discrepancies in predicted and observed nTiO2 transport based upon ZPs indicated that

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individual sediment component ZPs under experimental pH conditions need to be considered.

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Although effluent pH remained approximately constant during each experiment (Figure S3, SI),

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pH of the column effluents varied from the influent pH of 5 or 9 as the porous media buffered

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suspension pH as it travelled through the column, which could have large effects on ZP and

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therefore nTiO2 transport (Figure 4a). Xu et al. (2012) reported pH buffering by acidic soils was

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correlated to cation exchange capacity and deprotonation of SOM, consistent with the sediment

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used in this study 41. Effluent pH was buffered closer to the natural pH of each sediment (Table

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S2, SI), and was similar for experiments at both 5 and 9, slightly higher for pH 9 influents

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(Figure 4a). pH buffering by the sediments in the pH 9 experiments was stronger than in the pH

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5 experiments, as the sediments were naturally acidic (Table S2, SI).



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Figure 4. Column effluent pH, DOC, and zeta potential for each experiment.

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Differences in influent and effluent pH indicated that pH of the pore water changed along the

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column length, and consequently influenced ZP of nTiO2 and sediment components and therefore

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the nature of their electrostatic interactions. At either end of the column existed two pH end

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members: the influent, with pH adjusted to 5 or 9; and the effluent, with pH influenced by the 15 ACS Paragon Plus Environment

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transport media. To determine zones of theoretical attractive or repulsive electrostatic

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interactions between nTiO2 and various sediment components, pH of each column influent and

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effluent was plotted (pH 5 influent experiments at the top, and pH 9 at the bottom) along with

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charges of representative sediment components and nTiO2 (positive or negative, based on each

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component’s PZC) in Figure 5. Quartz was selected to represent the bulk minerals in the

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sediment, because the majority of minerals had strongly negative ZP under the pH range of our

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experiments 42–44. Fe oxide was representative of the metal oxyhydroxides in the sediment, due to

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its variable ZP with changing pH; whereas the major constituent of SOM, humic substances, was

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chosen for SOM. Although a wide range of PZC values have been reported for nTiO2 45, 6.2

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represents an average previously determined for nTiO2 under the same water chemistry

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conditions 22. If the true PZC of nTiO2 is slightly higher or lower than 6.2, and the line in Figure

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5 is shifted slightly, the explanation of results remains the same due to the range of charges

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experienced for each component along column length.



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Figure 5. Column effluent pH and ranges of positive or negative surface potentials for nTiO2,

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Quartz, SOM, and Fe oxides based on previously determined PZC (quartz and Fe oxide 26,

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nTiO2 46,47, SOM 25)

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For the pH 5 influent experiments, column effluent pH remained within a range where nTiO2

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was expected to be positively-charged, except for the AOD-H2O2 treated sediment, with effluent

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pH 6.31 (Figure 5). Although the AOD-H2O2 treated sediment column effluent pH was above the

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PZC of nTiO2 (and any nTiO2 that eluded the column was expected to be negatively-charged), no

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nTiO2 was detected in the column effluent because attraction to quartz and other minerals near

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the influent site caused retention of all of the nTiO2 in the column (before the pH was buffered

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above the PZC of nTiO2). Homoaggregation and straining was also possible under these

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conditions, however, the rate of aggregation in pore spaces could be much lower than deposition 17 ACS Paragon Plus Environment

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48

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component, accounting for the low recovery of nTiO2 for the pH 5 influent suspensions as

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attractive forces were expected, and nTiO2 was retained in the column. Repulsive forces existed

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between nTiO2 and Fe oxides in this pH range; however, the Fe oxides were in much lower

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concentration relative to the minerals in the sediment, and were unlikely to cause repulsion under

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these conditions.

. In all of the sediments, quartz and other negatively-charged minerals were a major

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Like quartz, forces between nTiO2 and SOM were also attractive under these conditions,

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therefore in sediments with sufficient SOM concentrations, nTiO2 was also attracted to SOM in

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the sediment. The SOM existed as either SSOM or leached as DOM; although SSOM may have

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attracted nTiO2 and removed it from pore water, DOM is known to facilitate the transport of

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ENPs. Wu and Cheng (2016) found that DOM, in concentrations as low as 0.13 mg/L, prevented

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nTiO2 attachment to quartz sand by adsorbing onto nTiO2 surfaces, thereby modifying nTiO2’s

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surface charge from positive to negative, and stabilizing it in suspension. Similar results have

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been reported indicating that DOM can play a major role in enhancing ENP transport

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experiments, the untreated and AOD treated sediments, which had not undergone the H2O2 SOM

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degradation process, contained the greatest concentrations of SOC (Figure 1) and were the only

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experiments with nTiO2 recovered in the column effluent (Figure 2a), as DOM leached from

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these sediments, attached to nTiO2 surface, and facilitated its transport. The highest

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concentrations of leached DOC in column effluents from the pH 5 experiments came from the

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untreated and AOD treated sediment (~ 0.20 mg/L) (Figure 4b), correlated with greatest nTiO2

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recovery (Figure 2a), whereas the quartz sand, H2O2, and AOD-H2O2 treated sediments had

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lower leached DOC concentrations (Figure 4b), and therefore facilitation did not occur, and all of

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the injected nTiO2 was retained (Figure 2a).

22,24

. In our



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Although ZP of influent nTiO2 was positive at pH 5, effluent samples from the untreated and

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AOD treated sediments had negative ZP, -28 to -30 mV (Figure 4c), correlated to DOM

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adsorption onto nTiO2 reversing its surface charge to negative. Effluent ZP was not measured for

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the H2O2 and AOD-H2O2 treated sediments in the pH 5 experiments due to no nTiO2 being

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detected in column effluents, correlated with lower DOC concentrations. Decreasing nTiO2

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recovery with time was observed for the untreated and AOD treated sediments (Figure 2a),

350

consistent with ripening behaviour; where nTiO2 already deposited in the column attracted nTiO2

351

with reversed charges, providing further evidence that DOM had adsorbed onto nTiO2 in these

352

experiments 24,38,49. HDD of all column effluents with nTiO2 recovered was smaller than that of

353

the influent, indicating that aggregation was not occurring for the eluted nTiO2 (Figure S4, SI).

354

In summary, two competing processes were responsible for nTiO2 behaviour in the pH 5

355

experiments; attachment of positively-charged nTiO2 to SOM, quartz sand, and other negatively-

356

charged minerals, removing it from the pore water, versus attraction of leached DOM to nTiO2,

357

which reversed the ZP of nTiO2 and facilitated transport. The majority of nTiO2 was retained in

358

the column (>95%), indicating that attachment was the dominant process, whereas DOM

359

influenced a small percentage. Physical differences between sediment and quartz sand (such as

360

sediment having smaller pore sizes, rougher surfaces) suggested that quartz sand columns would

361

yield greater transport of nTiO2, however, our results showed the opposite, indicating influence

362

of these physical properties are secondary.

363

pH buffering of influent suspensions by the sediments was much stronger in the pH 9

364

experiments, and a larger range of pH was experienced along the column lengths than in the pH

365

5 experiments (Figure 5). Within this range, as pH decreased from 9, ZP of nTiO2 changed from

366

negative, to near neutral, and positive when pH was below its PZC ~6.2 46,47. Additionally, 19 ACS Paragon Plus Environment

Page 21 of 32


367

surface charges on Fe oxides changed from negative to positive when pH was below their PZC

368

of ~8.3 26, whereas SOM and quartz remained negative across the whole range (Figure 5). These

369

large changes in pH and ZPs of nTiO2 and Fe oxides indicated that electrostatic forces between

370

nTiO2 and various sediment components were variable along the column lengths, leading to

371

complex nTiO2 transport behaviour.

372

Effluent pH of the quartz sand experiments at influent pH 9 (8.77) was not strongly

373

buffered along column length, and with negative surface potentials on both nTiO2 and quartz at

374

this pH, repulsive electrostatic forces were expected (Figure 5). This was consistent with nearly

375

100% nTiO2 recovery observed in quartz sand columns at pH 9 (Figure 2b). Like quartz sand, the

376

AOD-H2O2 sediment had low concentrations of SOC and Fe/Al-oxyhydroxides (Figure 1),

377

leaving a composition of mainly negatively-charged minerals, and therefore negatively-charged

378

nTiO2 was repelled and eluded from the AOD-H2O2 sediment columns as well (Figure 2b). High

379

recovery (93%) for the AOD-H2O2 treated sediment alike the quartz sand confirmed that nTiO2

380

homoaggregation and straining was not a major factor in our experiments at this pH; and

381

observed differences for the pH 9 experiments should be due to changes in Fe/Al-oxyhydroxides

382

and SOM concentrations.

383

About 80% of injected nTiO2 was recovered in untreated natural sediment column effluent

384

(Figure 2b). This high recovery was attributable to electrostatic repulsion between nTiO2 and

385

transport medium. Across the entire pH range of the untreated sediment column, nTiO2 carried

386

negative charges and was repelled by like-charged SOM, quartz, and other minerals (Figure 5).

387

Although anticipated electrostatic forces between nTiO2 and SOM were repulsive at pH 9

388

(Figure 5), a fraction of DOM may adsorb onto nTiO2 surfaces even at this high pH; and the

389

magnitude of DOM adsorption would increase with decreasing pH along the column length 23. It 20 ACS Paragon Plus Environment


Page 22 of 32

390

was therefore possible that in our experiments some of the leached DOM adsorbed onto nTiO2,

391

facilitating transport and contributing to high nTiO2 recovery. This process became more

392

important as pore water pH decreased along the column, which lowered energy barriers for

393

DOM adsorption onto nTiO2. The effluent pH (6.33) was very close to the PZC of nTiO2, and

394

therefore ZP was expected to be near zero. However, effluent ZP was found strongly negative at

395

-32 mV (Figure 4c), indicating that nTiO2 eluding the column could be covered by negatively-

396

charged DOM. Effluent DOC of the untreated sediment column with pH 9 influent was found at

397

0.33 mg/L, the highest of all of the experiments (Figure 4b), confirming that DOM could have a

398

strong effect on nTiO2.

399

To explain the 20% of nTiO2 retained in the untreated sediment, two mechanisms were

400

considered; attraction by Fe/Al-oxyhydroxides and the influence of surface roughness. Despite

401

electrostatic forces between nTiO2 and Fe oxides being repulsive at the pH 9 influent site, as the

402

suspension travelled through the column and was buffered by the sediment, it passed the PZC of

403

Fe oxides, and Fe oxides became increasingly important away from the influent site as

404

electrostatic forces became attractive and some of the nTiO2 was attached to oxides surfaces and

405

removed from suspension (Figure 5). Additionally, although pH ranges of 9 to 6.2 for SOM and

406

9 to 8.3 for Fe oxides suggested repulsive electrostatic forces with nTiO2 (Figure 5), SSOM and

407

metal oxides have irregular rough surfaces which are known to drastically reduce interaction

408

energies between ENPs and porous media, and therefore these sites may have attracted some of

409

the nTiO2 23,50,51. Bayat et al. (2015) found the rough surface of dolomite grains accounted for

410

10% Al2O3 ENP deposition under unfavourable conditions 51, whereas Wu and Cheng (2016)

411

reported that up to 15% of nTiO2 was deposited onto Fe oxide coating of quartz sand under

412

unfavourable electrostatics at pH 9 23. Surfaces on Fe/Al-oxyhydroxides and SSOM within


Page 23 of 32


413

sediments are rough in nature 23,52, and therefore may have similarly led to increased nTiO2

414

deposition even under unfavoured conditions. Repulsive forces between nTiO2 and

415

quartz/minerals/SSOM and DOM facilitation were the dominant processes governing nTiO2

416

transport in the untreated sediment, counteracted by attraction to Fe/Al-oxyhydroxides and rough

417

surfaces.

418

In contrast to the other pH 9 experiments with high nTiO2 mobility, no nTiO2 was

419

recovered from the H2O2 treated sediment column (Figure 2b). Buffered pH and complex

420

competing interactions between nTiO2 and varying proportions of Fe/Al-oxyhydroxides and

421

SOC accounted for the low recovery. The H2O2 treated sediment had SOC and therefore DOC

422

depleted (Figure 1c), and most likely retained all the injected nTiO2 in the buffered pH range of

423

8.3 to 6.2, where attractive electrostatic forces existed between nTiO2 and Fe oxides (Figure 5).

424

AOD-Al, AOD-Fe and DCB-Fe were relatively high for the H2O2 treated sediment compared to

425

the other treatments (except the untreated sediment) (Figure 1), whereas SOC was depleted and

426

DOC concentrations in the column effluent were low (Figure 4b), therefore deposition of all the

427

injected nTiO2 within the H2O2 treated sediment was most likely due to Fe/Al-oxyhydroxides

428

playing an important role in attracting nTiO2 in the absence of SOM13,27.

429

For the AOD treated sediment column effluent, 6% of the injected nTiO2 was recovered,

430

higher than nil recovery by the H2O2 treated sediment, but much lower than the other sediments.

431

The reason for low nTiO2 recovery for the AOD treated sediment was similar to that of the H2O2

432

treated sediment; DCB-Fe strongly attracted nTiO2 under reduced SOM/DOM concentrations

433

once pH was buffered below the PZC of Fe oxides. Even though AOD-Al and Fe were depleted

434

by the AOD treatment process, DCB-Fe was still 26% of the original sediment’s (Figure 1a and

435

1b). Additionally, AOD treatment reduced SOC by 79% (Figure 1c), which reduced DOM 22 ACS Paragon Plus Environment


Page 24 of 32

436

concentration in the column pore water, evidenced by lower DOC concentration in the column

437

effluent (0.19 mg/L, 57% relative to the untreated sediment) (Figure 4b). The nTiO2 that eluted

438

the column (6%) was strongly negatively-charged (-35 mV) at pH 6.26 (Figure 4c), indicating

439

that DOM had attached to and stabilized this small fraction of nTiO2. However, the DOM

440

covered nTiO2 would have also been attracted to positively-charged DCB-Fe, limiting the

441

facilitation of nTiO2 transport by DOM.

442

The pH 9 experiments revealed nTiO2 transport was influenced by pH buffering of the

443

sediments, which could cause electrostatic forces to change from repulsive to attractive when

444

charges on nTiO2 and Fe reversed at their PZC. Quartz and other negatively-charged minerals

445

always repulsed nTiO2 under these conditions, whereas Fe/Al-oxyhydroxides and rough surfaces

446

led to nTiO2 attachment. Additionally, leached DOM may facilitate nTiO2 transport. Comparable

447

nTiO2 transport in quartz sand and AOD-H2O2 treated sediment columns indicated variation in

448

physical properties (i.e., pore size, and surface roughness) were not responsible for the

449

differences in nTiO2 transport behaviour, whereas varying concentrations SOM and Fe/Al-

450

oxyhydroxides had substantial effects, emphasizing their important influence on ENP transport.

451

3.5 Environmental implications

452

ENPs in natural subsurface environments may travel through sediments with complex

453

compositions including SOM and Fe/Al-oxyhydroxides. Results of our column experiments

454

demonstrated that SOM and Fe/Al-oxyhydroxides could have big influence on nTiO2 retention

455

and transport, even though they were only minor components (< 1% by mass) in the sediment.

456

pH was a key factor that determined the role of SOM and Fe/Al-oxyhydroxides, since pH

457

strongly influenced ZP of nTiO2 and Fe/Al-oxyhydroxides, and therefore interactions of nTiO2

458

with various components in the sediment. Leached DOM, even at low concentrations, may 23 ACS Paragon Plus Environment

Page 25 of 32


459

considerably promote transport by adsorbing to and reversing ZP of nTiO2. Differences in nTiO2

460

transport in quartz sand and natural sediment columns emphasize the necessity of considering

461

influences of SOM and Fe/Al-oxyhydroxides when evaluating ENP transport in natural

462

sediments.

463 464

4. Acknowledgements

465

The authors acknowledge funding from the Natural Sciences and Engineering Research Council

466

of Canada's Discovery Grant (No. 402815-2012) and Canada Foundation of Innovation's Leaders

467

Opportunity Fund (No. 31836). Ms. Fisher-Power is a grateful recipient of the 2017-2018

468

Chevron Canada Limited Rising Star Award, which provided financial support throughout this

469

project.

470 471

Supporting Information (SI). Sediment treatment and characterization details, summary of

472

experimental conditions (Table S1) complete sediment properties (Table S2), sediment

473

mineralogy (Figure S1), nTiO2 breakthrough curves with error bars (Figure S2), column effluent

474

pH with time (Figure S3), and HDD of column effluents (Figure S4) are given in the SI.

475 476

Notes. The authors declare no competing financial interest.

477 478 479

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