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Nanomaterials and their Applications to Groundwater Remediation

Eric C. Plantenberg

GeoSci 750- Contaminant Hydrogeology

Dr. Tim Grundl

Fall, 2015

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Contents

Abstract Pg 3

Introduction- Nanomaterials and the Nano-effect Pg 4

Carbon-based Nanomaterials Pg 5

Nano-zero Valent Iron Pg 7

Mobility of Nanomaterials in a Porous Medium Pg 9

Potential Drawbacks of Nanomaterial Use for Groundwater Remediation Pg 11

Conclusion Pg 12

References Pg 13

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Abstract

Nanomaterials are an emerging technology that have found uses in many areas of science and

technology, including the remediation of contaminated groundwater. Because of their small size,

nanomaterials offer unique advantages that macro-scale media do not. This review describes the

current state of research concerning two nanomaterials that have demonstrated promise for

groundwater remediation. Single- and multi-walled carbon nanotubes (SWCNTs and

MWCNTs) are effective at sorbing metals and volatile organic molecules at higher rates than

commonly-used granular activated carbon. Nano-zero valent iron is a strong reducing agent that

shows potential for direct injection into a contaminant plume, allowing remediators to affect

contaminants that are difficult to reach using other technologies. Methods for applying these

new technologies are still being developed, and much research is underway to improve the

effectiveness and safety of these emerging technologies.

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Introduction- Nanomaterials and the Nano-effect

Engineered nanoparticles are an emerging technology with a diverse list of applications

in science and industry. A nanomaterial is defined as a material with at least one dimension less

than 100 nanometers (Borm et al, 2006). While nanomaterials such as ash particles, soil, and

biomolecules occur in nature, we now have the ability to create nanomaterials designed with

specific chemical and physical properties. For example, in the water industry, titanium dioxide

nanoparticles (nano-TiO2) have even been used to treat wastewater because of its photocatalytic

properties (Hofstadler et al, 1994), and polymer-composite nanomaterials drastically improve the

fouling resistance of filtration membranes (Tabara, 2009).

In many cases, nanomaterials have been shown to be superior to their counterparts that

are not nano-scale. This is due largely to what has been termed the nano-effect. Chemical and

physical interactions between two substances are affected in part by the relative free energies of

the surfaces where the two substances meet. In the macroscopic world, the surface area of an

object is relatively small when compared to that object’s volume, meaning that only the few

atoms at the surface of an object exert their effect of surface free energy. But when a substance

is reduced to the nano-scale, the surface area to volume ratio increases significantly, allowing for

a drastic increase in the effect of surface free energy of a substance (Dingreville and Cherkaoui,

2005). As we will see in the upcoming examples, this can drastically increase a materials

reactivity and sorptive capacity.

Unsurprisingly, nanomaterials have found applications in the field of groundwater

remediation, in some cases demonstrating their superiority to other used methods. So far, two

different nanoscale materials have shown promise as effective tools for groundwater

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remediation: carbon-based nanomaterials (CNMs), and zero-valent iron (nZVI). We will discuss

the current status of research concerning these promising nanomaterials for groundwater

remediation, the behavior of these particles in porous media, and some potential concerns and

drawbacks surrounding nanomaterial use in the environment.

Carbon-based Nanomaterials

Carbon has long been used as a medium in water treatment because of its capacity to sorb

large amounts of contaminants. In traditional pump-and-treat methods, contaminated

groundwater is removed from an aquifer and then treated before being injected back into the

aquifer. Activated carbon is common sorbent used to capture and sequester contaminants, and

can later be disposed of offsite. It can also be applied in situ, using methods such as permeable

reactive barriers (PRBs), in which a ditch is excavated in the subsoil of a contaminated aquifer,

downflow of a contaminant source, and then filled with sorbent materials such as activated

carbon. As groundwater moves though the sorbent layer via advection, the contaminant is

immobilized. This method has the benefit of requiring little maintenance after construction, but

the cost of installation is high. Further, because it requires excavation from the top of the

aquifer, is not ideal for reaching contaminants that might tend to congregate deeper in an aquifer,

such as dense non-aqueous phase liquids (DNAPLs) (Karn et al, 2009).

In the correct application, carbon nanomaterials have the potential to outperform

traditional granular activated carbon. So far, carbon nanomaterials have not been widely used in

the field, but they show promise for replacing granular activated carbon in many of its

applications (Matlochova, et al, 2013). In the lab, CNTs have been shown to be effective

sorbents of a wide range of contaminants. The effectiveness of a given sorbent material is

largely determined by the availability of reactive sites on the surface of sorbent particles. Due to

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the increased surface area to mass of CNMs, this results in more rapid equilibrium rates and

higher adsorption capabilities over a broader range of pH values than macro-scale carbon

adsorbents (Mauter and Elimelech, 2008).

So far, two forms of CNM have shown the most promise for environmental remediation:

single-walled carbon nanotubes (SWCNTs), and multi-walled carbon nanotubes (MWCNTs).

As with non-nano scale carbon, CNMs can be altered through the process of activation. In

general, carbon nanomaterials that have been activated through oxidation, which results in the

removal of impurities such as ineffective, amorphous carbon, tend to be more effective as

sorbents (Gotovac et al, 2007).

When applied to the removal of metal contaminants, including zinc, cadmium, nickel,

and lead, nano-scale carbon was shown to be more effective at sorbing contaminants than typical

activated carbon (Ruparelia et al, 2008). One study found that carbon nanotubes possessed that

capacity to adsorb 2.08 milligrams of Cu2+ per gram sorbent, compared to just 0.316 mg/g of

activated carbon. For Ni2+ removal, SWCNTs and MWCNTs sorbed 47.85 and 38.46 mg/g,

compared to only 26.39 mg/g for granular activated carbon (Pyrzyńska and Bystrzejewski,

2010).

Carbon nanomaterials show similar promise for the removal of organic pollutants. Shao

et al (2010) demonstrated the customizable nature of nanomaterials by grafting β-cyclodextrin to

the walls of MWCNTs, thereby increasing their affinity for sorbing polychlorinated biphenyls.

The result was a higher sorption capacity than unmodified MWCNTs at a broader pH range.

Other research demonstrated that CNTs, establishing equilibrium concentrations of

trihalomethane after only three hours (Lu et al, 2005). While there are multiple other studies

confirming the superiority of CNMs as sorbents, it is important to note that the majority of these

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studies have been conducted in a laboratory setting, and as of yet, little is known of how they

will behave in the environment. Before their widespread application, it is necessary to confirm

that CNMs are effective in real-world applications, as well as cost-effective and environmentally

benign (Matlochova, 2013). It is further important to note that sorbents, such as activated carbon

and CNMs do not degrade or transform contaminants, merely sorb and immobilize them.

Nano Zero-valent Iron

Iron compounds are some of the most widely used substances in groundwater

remediation due to their diverse forms and varying reactivity. Ferrous iron(Fe2+), for example,

plays an important role as a reducing agent, as it acts as an electron donor in its transformation to

Ferric iron (Fe3+). Various mineral forms of ferrous iron also have relatively high sorptive

capacities (Cundy et al, 2008).

Zero valent iron (Fe0, or ZVI) is the most commonly used granular compound in PRBs,

mainly due to its ability to degrade, reduce, or immobilize a wide range of contaminants (Tosco,

2014). When exposed to the oxygen, ZVI is rapidly oxidized to Fe2+, but in anaerobic

environments, it becomes a powerful reducing agent. In a three step process, electrons are

directly transferred from Fe0, before catalyzed hydrolysis by H/H2 and reduction by the Fe2+

species resulting from Fe0 corrosion (Chicgoua et al, 2011). Cundy et al (2008) further describes

how ZVI acts as a reducing agent for chloride compounds:

Fe0 + RCl + H3O+ → Fe2+ + RH + Cl− + H2O (1)

Nano-scale zero valent iron (nZVI) consists of particles of a size ranging from 50-200

nm, providing them with an effective surface area of 33.5 m2/g (Lien et al, 2006). In many cases,

particles can be modified by the addition of a surface coat, which changes the chemical

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properties of nanomaterials to improve characteristics such as mobility in a porous medium

(Kanel et al, 2007).

Like other nano-scale technologies, nano-zero valent iron (nZVI) shows enhanced levels

of reactivity and mobility in porous media due to its small size and increased surface area (US

EPA, 2007). Like its granular counterpart, nZVI is capable of degrading a wide array of organic

contaminants, including chlorinated methanes, ethanes, benzenes, and chlorinated biphenyls, as

has been demonstrated in laboratory settings (Elliot and Shang, 2001). Unlike CNMs, nZVI has

already been successfully applied in the field for in situ remediation. A 2006 study by O’Hara et

al reported relatively high remediation efficiencies, with a reduction in soil concentrations of the

DNAPL trichloroethane (TCE) of more than 80%, and reduction in groundwater TCE of 60-

100%.

As with other nanomaterials, nZVI’s increased surface area results in an increased rate of

reaction. Lowery and Johnson (2004) reported degradation rates of PCBs 1.4 to 38 times greater

than larger particles of ZVI. Similarly, Crane et al (2011) applied a combination of nZVI and

magnetite nanoparticles for the removal of uranium in water from a carbonate-rich aquifer.

Uranium was removed to less than 2% of its original concentration within the first hour of the

reaction period.

The potential benefits of nZVI

compared to other iron based

remediation tools comes from the

versatility of nZVI’s delivery method.

The nanomaterials can be injected into a

given aquifer as part of a slurry

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

Figure 1. Diagram of nanomaterials being delivered to the bulk of a DNAPL pool at the bottom of an aquifer. (Tratnyek and Johnson, 2006).

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consisting of water, a hydrophobic fluid, and an emulsifying agent to prevent agglomeration of

the particles in an aqueous environment. This delivery system allows remediators to deliver

nanoparticles directly to the source of the contaminant without costly and intrusive digging

methods. It also allows direct treatment of contaminants which are unreachable by digging, such

as DNAPLs (Cundy et al, 2008). The nanomaterials then move through the porous medium as

part of the normal advective flow, and thus degrade contaminants moving as part of a plume

(Figure 1).

Mobility of Nanomaterials in a Porous Medium

Once injected into the porous medium, colloids of nanomaterials have the potential to

flow with the surrounding groundwater, though this process is complex and still relatively poorly

understood. The motion of nanomaterials in a porous medium can be affected by physical

factors, such as interception (contact with grains of the medium), normal diffusion rates, and

gravitational sedimentation. Further, because nanomaterials are inherently chemically active,

factors such as adsorptive site density and blocking, charge heterogeneities in the medium, and

variations in the surface charge of the nanoparticles can all effect mobility (Liu et al, 2009). Due

to these factors, and diversity of nanomaterials currently being investigated, modeling and

understanding the aspects of mobility of

nanomaterials in a porous medium, such as

retardation and dispersion, is still in its

infancy.

Liu et al (2009) conducted column

experiments to measure several factors

effecting mobility of MWCNTs, in porous Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

Figure 2 Fitted and observed breakthrough curves of MWCNTs at a pore water velocity of 0.42 m/d in quartz sand and glass bead columns. Ionic strengths of the aqueous phase are 10 mM and 0.1 mM. (Liu et al, 2009).

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medium. They observed that breakthrough of MWCNTs was significantly affected by flow rate,

with lower flow rates resulting in a greater degree of retardation. Also, the ionic charge of pore

water influenced the rate of adsorption to the medium, with higher ionic strengths reducing

retardation, and allowing MWCNTs to more closely resemble those of a chloride tracer (Figure

2).

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Tosco et al (2014) further

illustrated the complexities of

nanomaterial mobility in porous

medium by summarizing the

results of several column

experiments with nZVI, which

tends to be relatively immobile

in porous media. This is largely

due to the tendency of nZVI particles to aggregate to one another as a result of attractive particle-

particle interactions, leading to pore clogging and mechanical straining of additional particles.

Particles may also exhibit the phenomenon or “ripening”, in which particles bound to the

medium attract and aggregate with particles currently in suspension, producing a plaque-like

deposit on the walls of pores (Figure 3). Conversely, when particle-particle interactions are

repulsive, the process of “blocking” results from the deposition of nZVI particles onto the

medium, which due to their repulsive nature toward one another, have little effect on the motion

of other particles. The clogging mechanisms were shown to be mitigated by the use of

stabilizing compounds, such as carboxymethyl cellulose (Kocur et al, 2013), and anionic

surfactants (Saleh et al, 2007). Research into the mobility of nanomaterials in porous media is

ongoing, and as of now, modeling techniques for understanding such phenomenon are limited

but improving.

Potential Drawbacks of Nanomaterial Use for Groundwater Remediation

While nanomaterial demonstrate great potential as a tool for groundwater remediation,

because the technology is still developing, many aspects are still poorly understood. For

Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014). Figure 3-Examples of mechanisms effecting mobility of nZVI particles in a porous medium. (Tosco et al, 2014).

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example, Grieger et al (2010) describes our lack of understanding of how injection of nZVI

effects biological and geochemical characteristics of an aquifer. In addition to potential clogging

of the porous medium, which can alter the flow of groundwater near the injection site, injection

of nZVI tends to yield a strongly alkaline reducing environment, and can induce shifts in the

microbial populations.

Additionally, many questions have been raised as to the safety of these chemicals. In

many cases, these concerns stem from a lack of knowledge rather than from clear evidence of

toxicity. Currently, there is no good methods for measuring nanoparticle concentration in an

aquifer after the particles have been introduced (Hassellov et al, 2008). This is important

information for understanding the particles’ tendency to migrate away from the injection site,

and for understanding the particles’ lifespan in porous media. So far, measurement of nZVI

concentrations after treatment have used indirect methods, such as monitoring of geochemical

parameters like pH (Mace et al, 2006).

Similarly, questions have been raised concerning the potential persistence of nZVI in the

environment. Little is known about the ability of bacteria to degrade the particles or the various

surface coatings which can be used to modify the particles’ reactivity (Geiger et al, 2010). This

directly relates to concerns over toxicity. One study by Li et al (2009) demonstrated that as little

as 0.5 mg/L of nZVI caused lipid peroxidation, a sign of toxicity caused by oxidative stress, in

the model organism Japanese rice fish. Carbon nanotubes have also been shown to cause in

increase in reactive oxygen species, a sign of toxicity caused by oxidative stress, in model

organisms (Pulskamp et al, 2007).

Conclusion

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While nanomaterials continue to find applications in science and industry, it is important

to consider their potential drawbacks, as well as their benefits. CNMs and nZVI have great

potential for immobilizing, removing, and degrading harmful contaminants, such as metals and

volatile organic compounds, from groundwater. Due to their small size and large surface area to

volume ratio, nanomaterials exhibit much higher efficiencies than are possible by their macro-

scale counterparts. This opens the door to new methods and new technologies. But as with any

new technology, it is important to be cautious, and to put forth the necessary research to ensure

that in attempting to remove hazardous chemicals from the environment, we are not creating

additional hazards.

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