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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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