Essay Writing Service

1 Star2 Stars3 Stars4 Stars5 Stars (No Ratings Yet)
Loading...

A Review and Toxicological Evaluation of an Environmentally-Friendly Alternative to Silver Nanoparticles

do not necessarily reflect the views of UKDiss.com.

Within the past two decades, the rise of nanotechnology has provided various technological and
industrial sectors with avenues for significant growth and improvements to existing practices.
With the inherent qualities which make materials on the nanoscale unique in behavior and
function, there are limitless applications of nanotechnology.  One of the predominant issues in
the field is the lack of data addressing fate of nanomaterials, particularly in natural conditions.
This is primarily due to the complexity of nanomaterial-environmental reactions, as the small
size and large reactive surface area of nanomaterials significantly complicate modeling
processes.  In addition to gaps in the literature concerning fate of nanomaterials, the regulation of
nanomaterials are also of concern, as there are no specific provisions in United States law which
specifically addresses nanomaterials.  Although data gaps exist for many nanomaterials, silver
nanoparticles are one of the most well-studied nanomaterials.  Due to their antimicrobial
properties, silver nanoparticles are used widely in consumer products.  It has been demonstrated
that silver can continuously leach from the nanoparticle, and can enter wastewater streams, which
may pose a risk to sensitive aquatic life.  To potentially reduce the burden of silver release from
conventional silver nanoparticles, our collaborators engineered a lignin-core particle doped with
silver ions and surface stabilized with a polycationic electrolyte layer.  Our objective was to
determine whether any of the formulation components elicit toxicological responses using
embryonic zebrafish.  Ionic silver and free surface stabilizer were the most toxic constituents,
although when associated separately or together with the lignin core, toxicity of the formulations
decreased significantly.  Formulations containing silver had a significantly higher prevalence of
uninflated swim bladder and yolk sac edema.  Comparative analysis of dialyzed samples, which
intended to simulate post-consumer use, showed a significant increase in mortality as the samples
aged, in addition to eliciting significant increases in types of sub-lethal responses relative to the
non-dialyzed samples.  ICP-OES/MS analysis indicated that silver ion release from the particle
into solution was continuous and the rate of release was component-specific.  Overall, our study
indicates that the lignin core is an effective alternative to conventional silver nanoparticles for
potentially reducing the burden of silver released into the environment.
 
TABLE OF CONTENTS
Page
CHAPTER 1: Introduction to Nanotechnology and Silver Nanoparticles………..………………1
1.1 Impact and Origins of Nanotechnology and Engineered Nanomaterials……………..1
1.2 Important Inherent Properties of Nanomaterials and Risk Characterization…………5
1.3 Applications of Nanotechnology ……………………………………………………13
1.3.1 Nano-Foods and Food Packaging…………………………………………..14
1.3.2 Nanomedicine………………………………………………………………19
1.3.3 Consumer Products & Technological Applications……………………….23
1.3.4 Nanopesticides…………………………………………………………….27
1.4 Silver Nanoparticles: State of the Science…………………………………………..35
1.4.1 Use and Prevalence………………………………………………………..35
1.4.2 Mechanisms of Action…………………………………………………….36
1.4.3 Transformations……………………………………………………………40
1.4.4 Regulatory Status…………………..………………………………………43
1.4.5 Applicability of my Research……………………………………………..48
1.5 Figures and Tables…………………………………………………………………..50
Table 1. List of Naturally-Occurring Nanoparticles…………………………….50
Figure 1. Display of Important External Factors Which Impact the Fate of                                            Nanomaterials……………………………………………………………………51
Figure 2. Types of Nanopesticides Researched in the Literature up to October                                           2013………………………………………………………………………………51
Figure 3.  Species Sensitivity Distributions (SSDs) of both Silver Salt and                                                  Silver Nanoparticles………………………………………………………………52
TABLE OF CONTENTS, Continued
Page
Figure 4. Transformations and Interactions of Silver Nanoparticles in                                                    Aquatic Media………………………………………………………..…………..53
Table 2. Tests Required for Production Volumes of Nanomaterials Under                                            REACH…………………………………………………………………………..53
1.6 References……………………………………………………………………………54
CHAPTER 2: Toxicological Assessment of a Lignin Core Nanoparticle Doped with Silver         as an Alternative to Conventional Silver Nanoparticles…………………………………………65
2.1 Abstract………………………………………………………………………………65
2.2 Keywords…………………………………………………………………………….65
2.3 Introduction…………………………………………………………………………..66
2.4 Materials and Methods……………………………………………………………….68
2.4.1 Materials and Characterization…………………………………………….68
2.4.2 Embryonic Zebrafish Assay……………………………………………….69
2.4.3 Toxicological Evaluations of Embryonic Zebrafish………………………69
2.4.4 Measurement of Dissolved Silver and Particle-Associated Silver………..70                                          2.4.5 Statistical Analysis………………………………………………………..71
2.5 Results and Discussion………………………………………………………………71
2.5.1 Particle Characterization………………………………………………….71
2.5.2 Analysis of Dissolved Silver and Particle-Associated Silver…………….72
2.5.3 Comparative Analysis of Formulation Toxicity………………………….73
2.5.3.1 Formulation Components………………………………………73
TABLE OF CONTENTS, Continued
Page
2.5.3.2 Dialyzed Formulations……………………………………………76
2.5.4 Analysis of Sub-Lethal Endpoints…………………………………………76
2.6 Figures………………………………………………………………………………..80
Figure 1. Concentration of silver associated with the filtrate and particle………80
Figure 2.LC50’s for formulation components (a) and dialyzed samples (b)…….80
Figure 3. Percent of zebrafish exhibiting significant sub-lethal responses………81   
2.6 Supplemental Information……………………………………………………………83
Figure S1. Representative images of zebrafish with and without significant                                           developmental impacts…………………………………………………………83
Figure S2. Average zeta potential and hydrodynamic diameter (HDD)                                            measurements for particle-containing formulations over a five-day period……83
Table S1. Metadata Associated with Zeta Potential Measurements…………….84
Figure S3. Concentration-response comparisons for formulation components                                               (a) and dialyzed materials (b) based on zebrafish mortality at 120 hpf…………85
Figure S4.Modeled concentration-response curve for the reference material                                             silver nitrate based on zebrafish mortality at 120 hpf……………………………86
Figure S5.Visual MINTEQ output for all silver-containing formulations………86   
2.7 References……………………………………………………………………………89
CHAPTER 3: Conclusion………………………………………………………………………100
CHAPTER 1: INTRODUCTION TO NANOTECHNOLOGY AND SILVER
 
NANOPARTICLES
 
1.1 Impact and Origins of Nanotechnology and Engineered Nanomaterials
Nanotechnology is broadly defined as an interdisciplinary area of research, development
and industrial activity that involves the manufacture, processing and application of materials that
have one or more dimensions between 1-100 nanometers (BSI Standards Publication, 2011;
National Nanotechnology Initiative, 2016; Nel, Xia, Mädler, & Li, 2006, ISO/TS 80004-2:2015).
Engineered nanomaterials are produced from these activities, and can be manipulated one atom
or molecule at a time, which can alter conductivity, reactivity, and optical sensitivity relative to
their bulk counterparts (Nel et al., 2006).  A nanometer is one billionth of a meter; when
materials are engineered on this scale, unique physiochemical properties emerge which include:
size (surface area and size distribution), chemical composition (purity, crystallinity, electronic
properties etc.), surface structure (surface reactivity, surface groups, inorganic or organic
coatings etc.), solubility and shape (Nel et al., 2006).  Engineered nanomaterials have a wide
variety of applications, ranging from incorporation in consumer goods, to drug delivery systems.
Due to their increasing use in the world market, careful consideration is needed to determine if
there is undue risk to either the environment or biological entities.
Although nanotechnology is now a prolific area of study, the definition for what
parameters constitute a nanomaterial remains a debatable topic amongst researchers.  Most
definitions include a reference to size and how many dimensions must qualify.  However, there
are several cases where a dimension may exceed the upper limit of 100nm proposed in most
definitions (nano-plates, nanopesticides etc.).  Although the nano- prefix is used, in these cases it
is only associated with the novelty or enhanced activity of a material (M. Kah, Beulke, Tiede, &
Hofmann, 2013).  This is perhaps why some definitions do not mention ‘unique’ properties in
comparison to bulk counterparts.  When ‘nano’ is used, it is inferred that because of their
nanoscale, nanomaterials should exhibit properties and behavior that differ from, or are
additional to, those of coarser bulk materials with similar chemical compositions (M. Kah et al.,
2013).  Two examples of organizations that address these issues directly in providing more
thorough definitions on what constitutes a nanomaterial are the USDA and the American Society
for Testing and Materials (ASTM).  The USDA provides a definition which makes an exception
to size and function: ‘a nanomaterial generally encompasses a size range of 1-100nm along at
least one dimension but, they may exceed that size, and be defined by physical or chemical
characteristics or behavior that distinguish them from bulk materials’ (McEvoy, 2015).  ASTM’s
definition addresses unique properties in their definition directly—but it is not a mandatory
requirement for a material to be classified as nano, as the defining factor is size (ASTM, 2012).
Although there are differing definitions of what may constitute a nanomaterial, it is helpful when
organizations clearly define these parameters and abandon ambiguity by not relying on possible
inherent meanings of terms.
In addition to engineered nanomaterials that are synthesized by researchers, nanoscale
materials can also be found in nature.  Cellular activity occurs at the nanoscale, in addition to
several biological structures, like hemoglobin (5.5 nanometers in diameter) and DNA at 2
nanometers in diameter.  Volcanic ash, sea spray, smoke from fire and by-products of
combustion from burning of fuels, such as coal and petroleum also contain nano-sized materials
(Bystrzejewska-Piotrowska, Golimowski, & Urban, 2009; National Nanotechnology Initiative,
2016).  Table 1 adapted from (Handy, Owen, & Valsami-Jones, 2008) describes in more detail
what kinds of nanoscale materials are found in nature, and their origins.  However, for the
purposes of this review, nanomaterials that are engineered will be focused upon due to their
prevalence and application in a wide variety of products and technologies.
The ideas and concepts behind nanoscience and nanotechnology started with a talk
entitled “There’s Plenty of Room at the Bottom” by physicist Richard Feynman at an American
Physical Society meeting at the California Institute of Technology on December 29, 1959
(Keiper, 2003; National Nanotechnology Initiative, 2016).  Feynman described a process in
which scientists would be able to manipulate and control individual atoms and molecules, and
how it would be an inevitable practice in the future (Keiper, 2003; National Nanotechnology
Initiative, 2016).  Although Feynman did not coin the term ‘nanotechnology’ at the time, over a
decade later, Professor Norio Taniguchi of Tokyo University of Science suggested it to describe
technology that strives for precision at the level of about one nanometer (Keiper, 2003).  Then, in
the 1980’s, the scanning tunneling microscope was invented, which allowed researchers to move
and manipulate atoms.  This event, along with the discovery of the “buckyball” in 1985 and the
carbon nanotube in 1991, began a rapid increase in in the implementation of nanotechnologies
(Keiper, 2003).
With the advent of these technologies and materials, interest in nanotechnology at the
federal level has increased significantly in the following years.  Various agencies, including the
U.S. Naval Research Laboratory had an interest in nanotechnology as early as the 1980s.  By
1997, the federal government was annually investing $116 million in nanotechnology, with the
funding doubling to $232 million by 1999 (Keiper, 2003).  In 2000, the Clinton Administration
pushed for more subsidies for nanotechnology and the creation of a National Nanotechnology
Initiative (NNI) that would coordinate the nanotechnology work of six different agencies
(Keiper, 2003; Maynard, 2006).  The NNI was approved with an initial budget of $422 million.
By 2010, the NNI’s research budget totaled an estimated $1.78 billion (Kessler, 2011), with
approximately 95% allocated to basic research into nanomaterial behavior, research facilities,
and developing nanoscale devices and systems, while the other 5% is allocated toward
environmental, health and safety research.   Since the founding of the NNI in 2000, more than 60
nations have established similar programs, and by 2010, worldwide annual public and private
sector funding for nanotechnologies was $17.8 billion (Sargent, 2016).  Globally, the
nanotechnology market is poised to grow at a compound annual growth rate of around 18.1%
over the next decade to reach approximately $173.95 billion by 2025 (Accuray Research LLP,
2016).
With an exponential increase in funding over the years due to the growing potential of
nanotechnology, there has been a parallel increase in global nanotechnology products.  Most
common among these nanotechnology products are engineered nanomaterials, including
nanoparticles.  (Piccinno, Gottschalk, Seeger, & Nowack, 2012) estimates that 9571 tons of
nanomaterials (the 10 most popular) are to be engineered each year.  In 2004, annual production
of nanomaterials amounted to about 1000 tons, and at that time, it was estimated that there were
already more than 800 products based on nanotechnologies in everyday use (Maynard, 2006).
Per the Woodrow Wilson Center’s Project on Emerging Nanotechnologies (PEN), between $60-
70 billion in nano-related products were sold in 2007 (Kessler, 2011; Maurer-Jones, Gunsolus,
Murphy, & Haynes, 2013).  To track the marketing and distribution of nano-enabled products
in the commercial marketplace, the same organization developed a broader international
coalition and formed the Nanotechnology Consumer Products Inventory (CPI) in 2005.  The
revised inventory was released in 2013, which listed 1814 consumer products that contained
nanotechnology from 622 companies in 32 countries (Vance et al., 2015).  An estimate indicated
that by 2014, more than 15% of all products on the global market will have nanotechnology
incorporated into their manufacturing process (Dawson, 2008).  As the global production of
nanoparticles increase, nano-enabled products are projected to grow to over half a million tons
by 2020 (Maurer-Jones et al., 2013), with the most popular including nanosilver, various forms
of carbon, zinc oxide, titanium dioxide, and iron oxide (Bystrzejewska-Piotrowska et al., 2009).
Further discussion of the types of sectors nanotechnology has had an impact on is included after
the next section: discussing the important inherent properties and risk characterization of
nanomaterials.
1.2 Important Inherent Properties of Nanomaterials and Risk Characterization
Nanomaterials possess inherent characteristics which make them attractive candidates for
utilization in numerous sectors.  The most obvious characteristic is size, which can influence
physical and chemical interactions with their environment due to increased surface area to
volume ratio in comparison to bulk materials.  Other important characteristics include shape and
surface structure (surface reactivity, surface groups, and inorganic or organic coatings), among
several others.  Although these inherent characteristics provide desirable uses in many
applications, they are also problematic as they impose an extra layer of complexity in terms of
determining fate and subsequent risk characterization; they are a chemical as well as a
nanostructure.  Besides inherent properties, there are also external factors which can impact the
behavior and subsequent fate of nanomaterials in both biological and ecological systems which
must be considered in order to build successful models for risk characterization.
In the past decade, modeling efforts have been expanded upon greatly which incorporate
transformation reactions, dissolution, phase transformations, heteroaggregation and
homoaggregation.  However, rapid changes of the initial form of the nanomaterial may lead to
overestimation of the initial forms of the nanomaterial (Gottschalk, Sun, & Nowack, 2013), and
there is a need to increase the complexity of systems and move from individual species or
environmental condition models to more complex mesocosms to determine the effects of
engineered nanomaterials in aquatic systems (Maurer-Jones et al., 2013).  Not only do we need
to consider the numerous intrinsic factors associated with the nanomaterial, but we also need to
consider the impact of factors in the surrounding environment.  For example, there can be
interactions between nanomaterials and non-nano pollutants, as they could adsorb organic
pollutants to the outer surface of the particle.  Nanomaterials can also potentially sorb metal ions,
which can increase transport and toxicity effects of these metals (Maurer-Jones et al., 2013).
Most publications within the past decade cite the importance of including these inherent and
external factors, but there has been difficulty modeling these processes, both in the laboratory
and in environmental media.  The primary difference in modeling fate in environmental media
compared to synthetic media is largely due to the dominant presence of natural nanoscale
particles and colloidal materials (Peijnenburg et al., 2015).  Additionally, environmental
exposure conditions are not homogenous (Gottschalk et al., 2013).
Another source of uncertainty is due to a lack of standardized reporting on
physiochemical characteristics.  (Handy et al., 2008) notes that detailed records would enable
accurate comparisons between data sets from different laboratories, or on different species with
the same materials.  Coupled with that is the uncertainty of these properties at all stages of use
(post-production, during use and after final treatment).  This complicates the process of obtaining
predicted environmental concentrations of nanomaterials, as nanomaterial production volumes
and emission rates from products are not reported (Furtado, Bundschuh, & Metcalfe, 2016;
Gottschalk, Sonderer, Scholz, & Nowack, 2009; Gottschalk et al., 2013; Maurer-Jones et al.,
2013; Peijnenburg et al., 2015).
Although the parameters that drive fate and behavior of nanomaterials in different
compartments are not yet fully understood, we may be able to draw from existing knowledge to
build more successful models for estimating nanomaterial fate (Peijnenburg et al., 2015).  Over
the past few years, computational modeling has emerged as a reliable tool to estimate the
underpinning parameters that control properties and effects of chemical substances via
quantitative structure activity relationships (QSARs).  QSARs read across available data on
structurally or functionally similar compounds that have already been tested to fill data gaps on
unknown nanomaterials.  In addition to QSARs, both (Harper et al., 2011; Liu et al., 2013) notes
that we may also collect data on nanomaterial-biological interactions to further build these
nanostructure-activity relationships to predict nanomaterial properties and activity in the absence
of empirical data.  With this knowledge, toxic responses may be able to be minimized by
manipulating the features of the nanomaterial.
The most distinguishing feature of nanomaterials is size.  As previously discussed, the
definition of ‘nanomaterial’ generally dictates that at least one dimension must be between 1-
100nm to be considered in the nanoscale—but this may not always be the case as some
nanomaterials exhibit unique properties when exceeding 100nm in one of their dimensions, but
are still considered a nanomaterial.  Although there are discrepancies on the size range at which a
material is considered nano, generally the impact of materials in this size range is primarily
associated with an increase in surface area to volume ratio compared to bulk materials of the
same composition.  As the size of a particle deceases, its surface area increases and allows a
greater proportion of its atoms or molecules to be displayed on the surface rather that the interior
of the material (Nel et al., 2006).  With a greater surface area, there is potential for improved
reactivity, which has been instrumental in creating catalysts for use in the automotive, chemical
and oil industries, as well as in applications for environmental remediation.
Some nanomaterials may affect biological behaviors at varying levels of organization,
as they can readily travel throughout the body, deposit in target organs, penetrate cell
membranes, lodge in mitochondria, and may even trigger injurious responses due to their small
size (Nel et al., 2006).  Also, cellular uptake efficacy is high for nanomaterials.  Sizes suitable for
uptake range from 10-500nm with an upper limit of 5mm; large particles are most likely to be
engulfed via micropinocytosis, particles that are ~100nm are taken up by clathrin-mediated
endocytosis and particles that are 60-80nm are taken up by caveolae-mediated endocytosis (Shin,
Song, & Um, 2015).  (Shin et al., 2015) concludes from their review that by decreasing particle
size and the resulting increasing surface area, biological activity increases substantially due to
greater reactivity.  Additionally, if particles are smaller, there is potential for a larger number of
particles to occupy a unit area, which then increases available surface area further.  These uptake
mechanisms are important particularly in the realm of drug delivery, as nanomaterials are
typically used as carriers to provide a dose of drug to an affected site.  However, the increased
reactivity of these particles can elicit reactions with other particles or chemicals, which can result
in possibly harmful effects when present in consumer products (Nel et al., 2006).
An increase in sorption capabilities due to a large surface area to volume ratio is also an
important feature of nanomaterials.  Nanomaterials can bind or carry a variety of molecules,
whether engineered by researchers to do so or are modified as a result of external factors and
interactions (De Jong, Borm, & others, 2008).  Therefore, physiochemical properties such as
hydrophobicity and charge of the nanomaterial can be altered which can affect stability and
mobilization of nanomaterials in the environment and in biological systems.  For example,
nanomaterials frequently interact with dissolved organic matter which then forms a nanoscale
surface coating and/or the replacement of existing surface coatings, which alters surface charge
of the particle (Maurer-Jones et al., 2013; Peijnenburg et al., 2015).  These coatings that form are
also impacted by other environmental factors, such as pH, ionic strength and temperature, to
name a few.  Alteration of the surface charge can lead to aggregation events, changes in
sedimentation rate and changes in dissolution.  Similarly, surface charge can impact both uptake
and distribution of the particle in biological systems; positively-charged nanoparticles are taken
up at a faster rate and are cited to be more toxic, primarily by damaging the cell membrane
(Albanese, Tang, & Chan, 2012; Juganson, Ivask, Blinova, Mortimer, & Kahru, 2015).
Therefore, it is critical that surface characteristics and sorption ability are considered as
parameters in future modeling efforts, as bioavailability and toxicity of the nanomaterial are
greatly influenced by these changes.
Another critical nanomaterial feature is shape.  Diffusion rates will change with the
aspect ratio of the material, and the physical shape may also make it difficult for particles to
approach each other due to steric hindrance (Handy et al., 2008).  However, the addition of
detergents or surfactants could coat the particle and change their shape or surface charge.  Shape
can also impact uptake—it was observed that for nanomaterials larger than 100nm, rod-shaped
nanomaterials had preferential uptake into cells, however; when the size was 50nm, spherical
nanomaterials were preferentially taken up (Albanese et al., 2012).  This corresponds with (Shin
et al., 2015)’s review that nanomaterials of this size range are compatible with the cell’s
machinery for uptake.  Although shape is commonly mentioned as an important parameter
affecting cellular toxicity, (Juganson et al., 2015) mentions that only 33% of publications
characterize the shape of the particle, with most particles tested being spherically-shaped.
Besides the inherent properties of nanomaterials, we must also consider external factors
and how they impact numerous processes when building models to characterize risk.  A few
important external factors include ionic strength and dissolved organic matter.  These external
factors can impact several processes, which have been alluded to already, including: aggregation,
sedimentation, and dissolution (Gottschalk et al., 2013; Lapresta-Fernández, Fernández, &
Blasco, 2012).  These processes are eloquently displayed in Figure 1, adapted from (Peijnenburg
et al., 2015).  As these processes and factors are not exclusive, they must be considered in
tandem to determine the fate of the nanomaterial.  Currently, there are major knowledge gaps
pertaining to how both external factors and processes impact nanomaterial fate, however;
partitioning behavior and degradation by chemical, physical and/or biological means are thought
to be the most important processes.
Ionic strength of the medium is an important factor in determining fate of nanomaterials
once released in the environment.  Additions of salt into the medium may provide charge
shielding and/or compress the charge on the surface of the nanomaterial so that the particle
collisions leads to attachments of particles, and therefore aggregation (Handy et al., 2008).  For
example, surface charge of the nanoparticle can be influenced by the presence of Ca2+, as it can
compete to screen a negatively charged surface (Handy et al., 2008).  Increasing ionic strength
increases the rate and extent of aggregation, and this is evident when moving from freshwater to
saltwater (Lapresta-Fernández et al., 2012; Maurer-Jones et al., 2013)  However, we can
intentionally modify the surface of nanoparticles can prevent or enhance the effect of ionic
strength on aggregation (Maurer-Jones et al., 2013).  For example, ionic strength, pH, and
dissolved organic matter all influence aggregation of nanomaterials in freshwater systems; in
saltwater environments, nanomaterials may aggregate to a greater extent due to increased ionic
strength (Lapresta-Fernández et al., 2012; Maurer-Jones et al., 2013).  And in turn, aggregation
may promote local high concentrations of nanomaterials in sediments, drive surface acting
toxicity to organisms without appreciable bioaccumulation, and influence uptake based on the
exterior chemistry of the organism (Handy et al., 2008).
Natural organic matter, which is the major fraction of dissolved organic matter, can
influence aggregation processes as it can sorb onto nanomaterial surfaces.  As previously
discussed, natural organic matter can displace weakly bound capping agents to form a dynamic,
heterogeneous layer of molecules on the surface of nanomaterials; the chemistry of the capping
agent is crucial in determining aggregation potential (Peijnenburg et al., 2015).  This process can
also be impacted by the degree of charge alteration from the ionic strength of the media.
However, there is significant evidence which indicates that aggregation is limited at realistic
concentrations of natural organic matter (1−30 mg of carbon/L) (Maurer-Jones et al., 2013).
Sorption onto nanoparticle surfaces may additionally aid in transformations beyond aggregation,
such as surface reduction, where natural organic matter can reduce ionic metals at the
nanoparticle surface to increase nanoparticle size (Maurer-Jones et al., 2013).  Natural organic
matter can also discourage homoaggregation, encourage dispersal, reduce possible sedimentation
and enhance stability by steric stabilization (Maurer-Jones et al., 2013; Peijnenburg et al., 2015).
For example, additions of negatively charged humic and fulvic acids to positively charged
mineral nanomaterials encouraged dispersal of nanomaterials in natural freshwater (Handy et al.,
2008; Peijnenburg et al., 2015).
Aggregation of nanomaterials is one of the most important processes, as it can be
impacted by numerous factors (two of which were described above) and can influence other
processes which impact overall fate.  Aggregation occurs because of the deficiency in stabilizing
mechanisms which prevents the natural tendency of nanomaterials to stick together due to van
der Waals forces (Lapresta-Fernández et al., 2012); a disruption of these forces can affect
mobility, size and bioavailability, which have been previously described.  Concentration of
nanomaterials does not affect aggregation rate, but the decrease in nanomaterial concentration
does result in lower collision frequency and the formation of smaller nanomaterial aggregates
(Peijnenburg et al., 2015).  Smaller aggregates can remain in suspension, travel for long
distances, and undergo increased dissolution.
Dissolution and sulfidation of nanomaterials are influenced by both solution chemistry
and intrinsic nanomaterial properties.  These two processes can be illustrated by using silver
nanoparticles as an example.  Silver nanoparticle dissolution in natural waters has been found to
be controlled by their primary particle size, shape, surface coating, concentration, dissolved
oxygen, pH, ionic strength, chloride and ammonia content, temperature, salinity, dissolved
organic carbon and aggregation (Peijnenburg et al., 2015).  In particular, precipitation of AgCl,
Ag2S or complexation with natural organic matter reduces the concentration of free silver ions.
The precipitate can also form a surface coating layer on the original nanoparticles, which results
in surface passivation to prevent further dissolution.  Silver nanoparticles can also retain their
Ag0 nature and continue to dissolve over extended periods of time.  In terms of sulfidation, it
does not occur homogenously throughout the particle surface, but heterogeneously.  Dissolution
also decreases when the ratio of sulfur to silver increases, which influences surface charge, which
in turn enhances the aggregation of silver nanoparticles (Peijnenburg et al., 2015).
This section describes the importance of considering both the intrinsic and extrinsic
factors that can influence nanomaterial fate in addition to outlining their important inherent
properties.  Although nanomaterials provide overwhelming benefits in many sectors due to these
unique properties, these properties add an additional layer of complexity when it comes to
predicting their behavior in the environment.  Although data gaps exist, we can begin to collect
empirical data to support assumed behavior, particularly in environmental media.  Although risk
characterization for nanomaterials is an active area of research, nanomaterials are used in
numerous applications which are described in the next section.
1.3 Applications of Nanotechnology
Due to the beneficial characteristics of materials on the nanoscale, the range of
applications vary widely.  Nanotechnology research has generally focused on understanding the
correlation between the optical, electrical, and magnetic properties of nanomaterials with respect
to size, shape, and surface chemistry (Albanese et al., 2012).  Additionally, nanotechnology
research has taken advantage of the natural scale of biological phenomena to produce solutions
for disease prevention, diagnosis, and treatment (National Nanotechnology Initiative, 2016).
And last, nanotechnology has been utilized in a wide variety of items intended for consumer use
and consumption, as well as remediating contaminated sites and media due to human activity.
Therefore, nanotechnology has been used to improve many sectors such as: information
technology, homeland security, medicine, transportation, energy, food safety, environmental
science, and environmental remediation, among others (Handy et al., 2008; National
Nanotechnology Initiative, 2016).  As there are many sectors in which nanotechnology has had
an impact, discussion of only a handful of sectors will be presented including: foods and food
packaging, medical applications with a focus on drug delivery, consumer products, and
pesticides.  These represent the largest sources of exposure to consumers, as well as the fastest
growing sectors in which nanotechnology is expected to play a role.
1.3.1 Nano-Foods and Food Packaging
One of the fastest growing applications for nanotechnology is within the food sector—
determining how the physicochemical characteristics of nanomaterials can change the structure,
texture, and quality of food, including improvements to packaging.  Origins of utilizing
nanotechnology in the food sector emerged from related sectors, such as pharmaceuticals,
cosmetics and nutraceuticals, and was considered as early as 2003 by the United States Federal
Drug and Food Administration (Chaudhry et al., 2008; Rashidi & Khosravi-Darani, 2011).
Utilizing nanotechnology in the food sector can improve taste, color, flavor, texture, and
consistency of foods, increase absorption and bioavailability of nutrients and health supplements,
create food packaging materials with improved mechanical and antimicrobial barriers, and to
provide nano-sensors for traceability and monitoring the condition of food during transport and
storage (Alfadul & Elneshwy, 2010; Chaudhry et al., 2008; Cushen, Kerry, Morris, Cruz-
Romero, & Cummins, 2012; Ramachandraiah, Han, & Chin, 2015; Rashidi & Khosravi-Darani,
2011).  Currently, nanotechnology-derived food packaging materials are the largest category of
nanotechnology applications for the food sector (Kessler, 2011).  Further innovations are
expected in the realms of production, processing, storage, transportation, traceability, safety, and
security of food, which makes nanotechnology a valuable resource to the food industry
(Chaudhry et al., 2008; Ramachandraiah et al., 2015).  Specifically, innovations from
nanotechnology coupled with the consumer’s acceptance of added-value in terms of quality,
freshness, new tastes, flavors, textures, safety, or reduced cost will dictate the future roles that
nanotechnology will have in the food sector (Cientifica, 2006, Online Report).
With several applications in the food sector for nanotechnology, different types of
nanomaterials are used, including nano-emulsions, nano-encapsulation, surfactant micelles,
emulsion bilayers and reverse micelles (Chaudhry et al., 2008; Cushen et al., 2012).  Nano-
encapsulation of food ingredients and additives can provide protective barriers, enhance flavor
and taste masking, controlled release, and enhanced dispersion for water-insoluble food
ingredients and additives (Chaudhry et al., 2008; Cushen et al., 2012).  (Kessler, 2011) cited the
use of nanoscale oil droplets in salad dressings and spreads that are intended to slow the
separation of ingredients, as well as the presence of nanoscale wax droplets on some fruits and
vegetables.  In terms of food packaging composite plastic bottles which incorporate nanoscale
clays to extend the shelf life of beverages are already on the market (Kessler, 2011).  Also,
nanomaterials which possess antimicrobial or oxygen-scavenging properties are common
additives to food packaging to extend the shelf-life of food items (Chaudhry et al., 2008;
Ramachandraiah et al., 2015). Other nanomaterials that act as sensors to monitor the quality of
the food and biodegradable polymer–nanomaterial composites are additionally used to improve
food packaging (Chaudhry et al., 2008; Cushen et al., 2012; Lee, 2010).
Virtually all known applications of nanotechnology in food and food packaging are in the
United States—however, Australia, New Zealand, South Korea, Taiwan, China, and Israel
participate as well (Helmut Kaiser Consultancy, 2009).  By 2006, the world market for nano-
enhanced food items was valued at $410 million, with food processing valued at $100 million,
food ingredients $100 million, and food packaging $210 million.  Conservative estimates
predicted that the market would grow to $5.8 billion by 2012, with food processing valued at
$1303 million, food ingredients $1475 million, food safety $97 million, and food packaging
$2930 million (Cientifica, 2006, Online Report).  As of 2006, 400 companies were currently
applying nanotechnologies to food, and it was expected to increase in the future, with companies
like Altria, Nestle, Kraft, Heinz and Unilever exploring nanotechnology for their products
(Alfadul & Elneshwy, 2010).
For manufacturers to sell products altered by nanotechnology to the public, regulations
in the United States appear to be lax.  The FDA is responsible for protecting and promoting
public health through the control and supervision of food safety—and although engineered
nanomaterials are recognized by the FDA, foods that contain them are generally considered safe
in comparison to the conventional form of the food item.  A 2010 report by the U.S. Government
Accountability Office confirmed this, as the “FDA’s approach to regulating nanotechnology
allows engineered nanomaterials to enter the food supply as GRAS (generally recognized as
safe) substances without FDA’s knowledge” (Kessler, 2011).  It has also been noted that
manufacturers are not required to label products containing engineered nanomaterials, and there
seems to be a recent trend toward dropping voluntary references to such ingredients from
packaging, websites, and other publications (Kessler, 2011).  However, in 2012, the FDA stated
that nanomaterials would be regulated the same as the bulk material of the same composition
(USEPA, 2014).
In 2014, the FDA published a Guidance Document, which provides non-binding
recommendations about the status of nanomaterials in items regulated by the FDA.  This
document verified that the FDA is aware of nanomaterial use in their regulated products,
however; an item may only be questioned further if it poses a potential risk to human health and
safety (FDA, 2014).  With a lack of information on product labels and non-standardized risk
assessment for these products, the public may not be aware of the risks and/or benefits to
incorporating nanotechnology into foodstuffs.  A review by (Cushen, Kerry, Morris, Cruz-
Romero, & Cummins, 2012) indicated that the public had an optimistic perception about
nanotechnologies in food products, but were found to have little knowledge about this
technology—which is not surprising based on the little knowledge we have on the potential risks
of nanotechnology in foodstuffs.
For organic products, the regulations differ in that there are more protections regarding
the addition of nanomaterials into these foodstuffs.  A policy memorandum published in 2015
addresses the position of the National Organic Board (NOB) of the FDA on including synthetic
nanomaterials during the processing or production of organic food items.  The National Organic
Standards Board (NOSB) within the NOB prohibits the use of nanomaterials unless it is added to
the National List of Allowed and Prohibited Substances which is approved by the Secretary of
Agriculture once proposed by the NOSB (FDA, 2015).  The amendments to the list must also go
through a round of public comment before any changes are made.  Then, if the changes are
approved, the nanomaterial may be used in the processing and production of organic food items.
This system of approval is strict in comparison to conventional foods, and is reminiscent of
European laws, where chemicals (or nanomaterials) are not automatically assumed to be GRAS.
In contrast, the European Legislation has a specific regulation pertaining to novel foods
and novel food ingredients—this regulation establishes a mandatory premarket approval system
for all novel foods (Chaudhry et al., 2008). A ‘‘novel’’ food is defined as a food or food
ingredient not having a significant history of human consumption prior to May 1997 and must fit
into one of these two categories: (1) foods and food ingredients with a new or intentionally
modified primary molecular structure and (2) foods and food ingredients to which has been
applied a production process not currently used, where that process gives rise to significant
changes in the composition or structure of the foods or food ingredients which affect
nutritional value, metabolism or level of undesirable substances (Chaudhry et al., 2008).
Additionally, there is another regulatory control which governs the composition, properties and
use of any material or article intended to come into contact directly or indirectly with food.  The
engineered nanomaterials must be sufficiently inert to preclude substances from being
transferred to the food in quantities large enough to endanger human health, or to bring about an
unacceptable change in the composition of the food or a deterioration in its organoleptic
properties (Chaudhry et al., 2008). The regulation applies to all materials, including plastics,
paper, metals, glass, ceramics, rubber etc.  The issue with this regulation is that it only precludes
the use of substances if they are transferred in quantities large enough to endanger human health,
which is an unquantified amount for most nanomaterials due to the lack of knowledge this area
(Chaudhry et al., 2008).
Risk has been marginally characterized by hypothesizing the behavior of nanomaterials
after exposure, primarily through ingestion, although inhalation and dermal exposure have also
been identified as possible routes (Cushen et al., 2012).  The consumer-safety implications are
intrinsically linked to the physicochemical features of the nanomaterials, which must be fully
characterized as this is what dictates the likelihood and extent of exposure.  The application of
nanotechnology in food has, therefore, led to concerns that ingestion of nano-sized ingredients
and additives through food and drinks may pose certain hazards to consumer health.  With food
packaging, potential migration of nanoparticles from the packaging into food and drinks with
subsequent ingestion is the main route of exposure (Chaudhry et al., 2008; Lee, 2010).  However,
migration data are not currently available, although there are numerous products that contain
nanomaterials in their packaging which are already on the market (Cushen et al., 2012).  For
foods treated with nanomaterials, direct consumption of food and drinks is the main route
exposure.  Nano-sized food ingredients and additives are likely to have a greater ability to cross
the gut wall, leading to enhanced absorption and bioavailability to result in higher plasma
concentrations (Chaudhry et al., 2008).
Part of the reason there is ambiguity in regulating nano-enhanced food products are due
to the complexities nanomaterials pose in evaluating risk, as well as a lack of a single
comprehensive regulatory framework to ensure consumer safety, particularly in the United States
(Corley, Scheufele, & Hu, 2009).  This makes it difficult to evaluate each nanomaterial as there
are no screenings or protections which protects the consumer.  Evaluating risk is impeded by a
general lack of knowledge, insufficient models, and uncertainties with respect to oversight by
government agencies.  A suggestion to utilize the precautionary principle in terms of
incorporating engineered nanomaterials into the food stuffs was indicated by (Kessler, 2011);
however, the United States does not practice this.  For now, food items that incorporate
nanotechnology are identified by the FDA, but the GRAS status prevails unless there is a future
event in which negative risk of using that technology is elevated.
1.3.2 Nanomedicine
Use of nanotechnology in medicine is another major application that is providing avenues
for disease prevention, diagnosis, and treatment.  Advances in nanomedicine have provided the
opportunity for therapies that are more precise by delivering the drug locally to the target area—
and that can be applied earlier in the course of a disease and lead to fewer adverse side-effects
(National Nanotechnology Initiative, 2016).  Due to the specificity, using nanotechnology
increases the therapeutic index of the drug, with a wider margin between the dose needed for
clinical efficacy and the dose-inducing toxicity.  Better imaging and diagnostic tools enabled by
nanotechnology are also paving the way for earlier diagnosis, more individualized treatment
options, and better therapeutic success rates (National Nanotechnology Initiative, 2016; Nel et
al., 2006).
There have already been successful examples of utilizing nanomaterials in the capacities
mentioned above—as well as some exciting possibilities for the future.  Gold nanoparticles have
been used as probes to detect targeted sequences of nucleic acids, as well as in treatments for
cancer and other diseases.  Nanotechnology is also being considered as a replacement for
conventional vaccine delivery, as well as a multiple-strain flu vaccine, which will require fewer
resources to develop (National Nanotechnology Initiative, 2016).  Nanomaterials are currently
being engineered to diagnose and treat atherosclerosis by mimicking HDL (good cholesterol), as
well as nanomaterials which mimic the crystal mineral structure of bone or resin for dental
applications (National Nanotechnology Initiative, 2016).  Perhaps the largest application
mentioned previously is drug delivery—nanomaterials can help deliver the medication directly to
the affected site while minimizing the risk of damage to healthy tissue.  This realm of
nanomedicine has grown significantly, particularly in developing therapies for treating cancer.
Nanomaterials have a large surface area to mass ratio which enhances their ability to
bind, adsorb and carry other compounds such as drugs, probes and proteins (De Jong et al.,2008).
The primary goals in engineering nanomaterials for drug delivery include: creating
drug/nanoparticle complexes that can specifically target areas and successfully deliver the drug
locally; reduce toxicity while maintaining therapeutic effects; aim for greater safety and
biocompatibility; and develop new, safe medicines quickly (De Jong et al., 2008).  However,
there are many challenges that must still be overcome to better understand the
pathophysiological basis of disease, bring more sophisticated diagnostic opportunities, and yield
improved therapies (De Jong et al., 2008).
Historically, there have been three generations of nanomaterials that have been
engineered for biomedical applications.  The first generation consisted of novel nanomaterials
functionalized with basic surface chemistries (non-stealth) to assess biocompatibility and toxicity
(Albanese et al., 2012).  However, many of the studies used serum-free media or did not account
for serum-protein interactions with the nanomaterials, which is not particularly relevant in how
the nanomaterials would interact with the human body.  First generation nanomaterials also did
not use polyethylene glycol (PEG); thus, most in vivo data show the rapid clearance of
nanomaterials (Albanese et al., 2012).  The poor stability of the nanomaterials and rapid
clearance led to the second-generation nanomaterials.
The second generation consisted of nanomaterials with optimized surface chemistries that
improved stability and targeting, particularly for treating cancer.  The surface coating of the
nanoparticle is important for preventing agglomeration and keeping the particles in colloidal
suspension—common examples include: PEG, poly(vinylpyrrolidone) (PVP), dextran, chitosan,
and surfactants like sodium oleate and dodecylamine (De Jong et al., 2008).  Most studies
focused on tumor delivery as a proof of concept, utilizing active targeting and stealth (maximize
blood circulation half-life for continuous delivery of nanoparticles into the tumor via leaky
vasculature) (Albanese et al., 2012).  However, there are several concerns with this generation of
nanomaterials: an overreliance on the enhanced permeation and retention (EPR) effect to deliver
nanoparticles into the tumor; no single nanoparticle size can access all areas of the tumor and
accumulate in significant quantities; and the advantages of active targeting are offset by the
barrier effect (most nanoparticles only travel within the first few layers of cells as they adhere to
their targeted receptors) (Albanese et al., 2012).
The third generation of nanomaterials uses biological, physical, or chemical cues in the
target environment to trigger a change in their properties to maximize drug delivery.  Two
approaches have been used so far: cues inside the tumor environment such as low pH, low
oxygen, or matrix metalloproteinase enzymatic activity; and the second is an artificial cue, such
as the application of near-infrared light inside the target tissue (Albanese et al., 2012).  Using
heat and light can provoke the therapeutic effect (cell death in the case of a tumor)—these
thermosensitive nanoparticles may be used for selective release of the drug after the
nanoparticles reach the desired location in the body (De Jong et al., 2008).  These approaches are
independent of tumor antigens needed previously in the second-generation nanomaterials and do
not rely on the EPR effect.
Nanomaterials used in drug delivery can be of biological origin, like phospholipids,
lipids, lactic acid, dextran, chitosan, or can be engineered from chemicals like polymers, carbon,
silica, and metals (De Jong et al., 2008).    Drugs can be entrapped within nanomaterials of these
compositions to enhance delivery to or improve uptake by target cells.  Additionally, entrapping
the drug within the nanomaterial may reduce the toxicity associated with the free drug to non-
target organs—both scenarios will increase the therapeutic index.  However, one of the problems
with using a nanomaterial to entrap the drug is that the mononuclear phagocytic system of the
liver and spleen can attack the nanomaterials (De Jong et al., 2008).  Liposomes that are
engineered in the nano-size range can also be an effective drug delivery system.  Composed of
phospholipids, they are flexible and biocompatible which allows them to pass along arterioles
and endothelial fenestrations without causing clotting (De Jong et al., 2008).
Although there have been significant strides in improving the efficacy of nanomaterials
engineered for drug delivery, this has come at the expense of fundamental studies which describe
the relationship between exposure to nanomaterials and biological responses.  As nanoparticles
are used for their unique reactive characteristics, these characteristics may also have an impact
on their toxicity. There is not enough information to draw conclusions about how the impact of
size, shape, and particularly surface chemistry-dependent interactions will affect biological
responses (Albanese et al., 2012; De Jong et al., 2008).  These features may lead to changed
body distribution, passage of the blood brain barrier, and triggering of blood coagulation
pathways (De Jong et al., 2008).  As a starting point to build this area of knowledge, full
characterization of the nanomaterial must be determined, and then testing must be completed to
determine whether the nanoparticle carrier’s size, shape etc. impacts toxicity.  An understanding
of fate of the nanoparticle inside of the body is also lacking.  Particles are generally found in
endosomes or lysosomes where they are degraded inside of the cell.  Chemical characteristics
such as surface charge may also determine the fate of nanoparticles in cells due to binding,
uptake and intracellular transport (De Jong et al., 2008).
1.3.3 Consumer Products & Technological Applications
Currently, there are many types of nanoproducts on the market for consumer use, as well
as in various technological applications where consumers directly benefit.  Items deliberately
engineered with nanotechnologies intended for consumer use perhaps pose the largest risk of
exposure—due to prevalence and therefore consistent exposure.  Additionally, there are no
standardized methods for assessing consumer risks or a set of agreed upon metrics for
characterizing nanomaterials to determine environmentally relevant concentrations (Shin et al.,
2015).  Typically, nanomaterials are either added to the bulk material to reinforce the physical
properties of the material or applied on the surface of the product to provide enhanced surface
features such as scratch resistance, water repellency, reflectivity, photoactivity, and for
antimicrobial protection (Bondarenko et al., 2013).  (Vance et al., 2015) echoes the same
beneficial features, but also adds the following: anti-caking properties, miniaturization, for
hardness and strength, for health and cosmetic application, and for environmental treatment.
These enhanced features that nanomaterials provide results in thousands of potential applications
for consumer products, which are expected to become more prevalent in the future.
Based on the Project for Emerging Nanotechnologies (PEN) Nanotechnology Consumer
Products Inventory (CPI), over 1800 registered products are on the market as of 2013.  Health
and fitness items were cited as the category that had the most products containing nanomaterials
in the CPI.  Also, silver was the most prevalent nanomaterial at 24% of all reported products due
to its well-known antimicrobial properties (Vance et al., 2015).  Titanium dioxide, zinc oxide,
and various forms of carbon are also prevalent in consumer products, with 46% of products
containing at least one of these nanomaterials (Royal Commission of Environmental Pollution,
2008; Vance et al., 2015).  However, there are some limitations to the CPI.  Approximately half
of the materials in the database do not provide the composition of the nanomaterials they contain,
and there is a lack of science-based data to support manufacturer claims (Vance et al., 2015).
Additionally, it has been noted that manufacturers are reluctant to provide production amounts of
chemicals, yet this information is important for exposure models and databases (Piccinno et al.,
2012).  Although these limitations exist, the CPI is the most comprehensive database which
documents consumer products on the market that contain nanomaterials.
The CPI describes six broad categories in which consumer products can be grouped:
health and fitness, electronics, home and garden, food and beverage, appliances and automotive.
Health and fitness is the largest category, as 42% of all reported materials in the CPI database
contain nanomaterials.  Most the materials within this category are the personal care items, like
lotions, toothbrushes, hairstyling tools, etc.  Other items in the health and fitness category
include: clothing, cosmetics, sporting goods, filtration, sunscreens and supplements (Vance et al.,
2015).  (Vance et al., 2015) discusses that due to the large proportion of personal care items that
have been reported containing nanotechnologies (34%), there is a particularly high risk of dermal
exposure, despite not knowing the size or the concentrations of the nanomaterials.  This is due to
the way these products are intended to be used: either by directly touching the product as the
nanomaterials are present on the surface of the solid product; or, liquid products that contain
suspended nanomaterials and are meant to be applied to hair or skin.
It is also possible to be exposed to these nanomaterials in the health and fitness category
by inhalation and ingestion.  Although not as likely as dermal exposure, approximately 25% of
items in the CPI pose a risk for inhalation exposure (hairsprays, hairdryers etc.) and 16% of
items pose a risk through ingestion (supplements, throat sprays, etc.) (Vance et al., 2015).  The
identity of nanomaterials in approximately half of these products is unknown, primarily due to
non-mandatory reporting requirements; this dramatically increases the uncertainty of exposure
risk.  However, the products that have known nanomaterial content reveal that metal-based
nanoparticles are the most abundant, and could therefore be a starting point in determining risk to
human and environmental health via these three uptake pathways.  Additionally, it is expected
that in the future, there will be an increasing number of materials that will contain nanomaterials,
so the prevalence and potential exposure risk will increase.
Technological applications in which nanotechnology is utilized span a variety of areas
which include: transportation, energy, electronics and information technology, and
environmental remediation.  The (National Nanotechnology Initiative, 2016) has cited several
advancements in these broad areas, and like consumer products, it is expected that these sectors
will continue to utilize nanotechnology to expand potential applications.  A significant benefit
within the transportation sector includes the use of nanoparticles as catalysts to reduce the
quantity of catalytic materials needed, which saves resources, money and reduces pollutants
released into the environment.  Additionally, nanomaterial use in structural applications within
the transportation sector can improve the longevity of transportation infrastructure as well as
making transportation vehicles more lightweight while also increasing strength of the building
materials (National Nanotechnology Initiative, 2016).
Energy, electronics, and information technology applications may be the most prolific
use of nanotechnology within these broad categories.  The general goal of utilizing
nanotechnology in the energy sector is to enhance alternative energy approaches by increasing
efficiency, to reduce pollution, and to reduce energy consumption (National Nanotechnology
Initiative, 2016).  Improvements have been made in areas of fossil fuel extraction and
production, reduction of pollutants from energy producing facilities by use of carbon nanotubes
in scrubbers, increased efficiency of energy harvesting for renewable energy sources
(particularly solar) and optimization of consumer products like electronics that are quick-
charging, more efficient, lighter weight, have a higher power density, and hold electrical charge
longer.  Generally, electronics have become more portable, lightweight, faster, sleeker designs
and can store more information in comparison to older electronics. These include items like
transistors, smartphones, e-readers, televisions, memory chips, hearing aids, and even in
aerospace applications (National Nanotechnology Initiative, 2016).  As items become more flat,
flexible, lightweight, non-brittle, and highly efficient, there are opportunities for new and
improved items to enter the market.
The last major technological application in which nanotechnology has had a significant
impact is in environmental remediation.  Besides increasing the efficiency of current energy
systems, there is potential for nanotechnologies to assist in detecting and removing contaminants
from affected sites.  Two recent examples include the use of molybdenum disulphide filters
to increase the rate at which the process of desalination occurs by up to five times, and the
use of chemical reactions to remove contaminants from industrial sources which reduces costs as
water does not have to be pumped above ground for treatment (National Nanotechnology
Initiative, 2016).  There are also nanotechnology-enabled sensors that can detect and identify
chemical or biological agents in the air and soil with high sensitivity, and their use is being
investigated in toxic site remediation (National Nanotechnology Initiative, 2016).  Nanomaterials
may also be used as mechanical filters in conjunction with carbon filters to remove particulates
in the air as well as remove odors from enclosed spaces, like airplanes.
The use of nanotechnology in technological applications and in consumer products is vast
as researchers can tailor the structures of engineered nanomaterials.  Materials can be engineered
to be stronger, lighter, more durable, more reactive, among many other traits.  With
approximately 15% of consumer products containing nanomaterials on the market currently, and
an estimated global growth of 18.1% over the next decade, nanotechnology’s role in consumer
products and technological applications will continue to be a large presence in global market.
1.3.4 Nanopesticides
Nanopesticides are an application of nanotechnology, and have become a recent interest
to the research community particularly over the past decade.  A broad definition of nanopesticide
has been provided by (Melanie Kah & Hofmann, 2014): “all plant-protection products that (1)
intentionally include entities in the nanometer size range (up to 1000nm) (2) are assigned with a
‘nano’ prefix and (3) are claimed to exhibit novel properties associated with the small size of
their components.  However, with the contested definition of what a nanomaterial is and the size
at which materials are considered ‘nano’, some nanopesticides may be excluded due to size or
formulation type.  Some of the formulations of nanopesticides exceed the limit of 100nm
proposed in most definitions of nanomaterials, but the prefix nano- is used as the formulation
has novelty or enhanced activity (M. Kah et al., 2013).  As of 2013, the majority of research
concerning nanopesticides were published on insecticides (~55%), then fungicides (30%) and
then herbicides (15%) (Melanie Kah & Hofmann, 2014).
There are several types of nanopesticides, and they are engineered to successfully deliver
the active ingredient to the target organism.  Some examples of nanopesticide formulations
include: emulsions, polymer-based, solid lipids, porous hollow silica, dispersions, and metals or
oxides.  Based on a literature review from (Melanie Kah & Hofmann, 2014), the most common
nanopesticide formulation are the polymer-based formulations (see adapted Figure 2), which
may be based on the following three criteria listed, as polymer-based formulations can achieve
these three goals.  The most common goals of these different formulations include increasing the
solubility of the active ingredient, to alter the release dynamics of the active ingredient (slow or
quick release), and to protect against premature degradation.  Although there are promising
advances in nanopesticide technology, fate and transport of these materials is still largely
misunderstood, as well as how we should regulate these materials.
Many pesticide active ingredients, like pyrethroids for example, have low water
solubilities, which makes them excellent candidates for nanotechnology.  Poorly soluble active
ingredients can be encapsulated, have surfactants added to them, or can be formulated as
emulsifiable concentrates.  Emulsifiable concentrates consist of an active ingredient dissolved in
an organic solvent and a blend of surfactant emulsifiers to ensure spontaneous emulsification
into the water (M. Kah et al., 2013).  Compared to microemulsions, nanoemulsions are typically
between 20-200nm and have approximately 5-10% less surfactant present, so less is needed to
provide the same benefits of using an emulsifier.  Additionally, it has been proposed that uptake
would be enhanced in a nanoemulsion compared to a microemulsion, although literature is scarce
to support this idea.  However, these formulations have poor stability after dilution, and the
solvent may increase the cost, could pose a dermal risk to handlers, and/or may be flammable.
As an alternative, oil/water emulsions have been suggested, as they consist of a mixture of a non-
ionic surfactant, block polymers and polymeric surfactants (M. Kah et al., 2013).  Although this
is a preferable approach to the emulsifiable concentrates, they require a large energy input to
engineer.  Microemulsions tend to form spontaneously when mixed with water, are stable, and
are already available on the market.  As nanoemulsions are difficult to prepare and stabilize, it is
more likely that microemulsions will dominate the industry—perhaps until regulations require
less surfactants to be used in the engineering of these formulations or there is a need for a more
concentrated presence of the active ingredient (M. Kah et al., 2013).
Encapsulation may also provide a mechanism to increase the apparent solubility of the
active ingredient, as well as provide slow/targeted release.  Encapsulations can be polymer-
based, solid lipid-based, or can be porous hollow silica nanoparticles.  Polymer-based
formulations can be composed of polysaccharides, polyesters, or even natural products, like
beeswax or corn oil.  Within the polymer-based formulations, there are several types of
preparations, which include: nanospheres, nanocapsules, nanogels, and electrospun nanofibers.
Nanocapsules have a core-shell structure that can act as a reservoir for active ingredient
dissolved in a polar or non-polar solvent (Anton, Benoit, & Saulnier, 2008).  The distribution of
active ingredient in nanospheres is uncertain.  Nanocapsules may improve the stability of the
spraying solution, increase uptake, increase the praying surface, ad reduce phytotoxicity due to a
more homogenous distribution (M. Kah et al., 2013).  However, a disadvantage to nanocapsules
rather than microcapsules may be that due to size, the concentration of active ingredient in the
nanocapsules relative to the nanocapsule itself may not be sufficient to elicit the effect on the
target organism (M. Kah et al., 2013).  Although not a direct measure of the active ingredient’s
concentration between micro and nano-sized capsules, (Meredith, Harper, & Harper, 2016)
determined that when the capsules were separated by size, there was no significant effect on
toxicity to embryonic zebrafish.  Both size fractions elicited the same responses, which may
indicate the active ingredients’ concentration within both the nano and micro-sized capsules were
the same.
Nanospheres, nanogels and electrospun nanofibers are different preparations within the
polymer-based nanoparticles which are still being evaluated for their efficacy.  Only a handful of
studies have considered these preparations, although potentially promising.  Lansiumamide B,
which is a molecule extracted from an Asian evergreen tree, may hold nematicidial properties
when formulated as a nanospheres (Han, Li, Hao, Tang, & Wan, 2013; Yin, Guo, Han, Wang, &
Wan, 2012).  Nanogels are being explored as a mechanism of pesticide delivery in organic
farming practices with pheromones, essential oils or copper as the active ingredients (Melanie
Kah & Hofmann, 2014).  Nanogels can provide protection against rapid evaporation of the active
ingredient, can resist degradation processes more effectively due to their insolubility in water and
resistance to humidity, as well as improving the loading and release profiles of active ingredients
(Melanie Kah & Hofmann, 2014).  Additionally, (Brunel, El Gueddari, & Moerschbacher, 2013)
utilized chitosan nanogels to deliver copper ions as an antifungal for wood.  The nanogel
provided a medium in which copper was able to be released over a longer period of time, as well
as provided some antifungal activity in addition to the copper.  Electrospun nanofibers are
another interesting polymer-formulation preparation, which can provide continuous release of
the active ingredient without ‘bursts’, or sudden releases, of active ingredient that nanocapsules
or nanospheres may exhibit.  The limited data is inconclusive on whether the use of electrospun
nanofibers provide additional benefits in comparison to other nano-formulated preparations, so
conclusions about their efficacy cannot be drawn.
Solid lipid nanoparticles are typically used in pharmaceutical applications, and have the
advantages of emulsions and liposomes with those of polymer nanoparticles.  However,
researchers are beginning to utilize solid lipid nanoparticles for delivery of pesticides.  For
example, a second-generation solid lipid nanoparticle has been developed, incorporating liquid
lipids into the solid matrix to increase the concentration of active ingredient and decrease
leakage.  However, the use of these particles to protect deltamethrin from photodegradation was
successful albeit significant losses of deltamethrin, which is undesirable (Nguyen, Hwang, Park,
& Park, 2012).  Due to the infancy of this formulation for delivering pesticides, coupled with the
engineering process which is energy intensive, there is a limited amount of publications on solid
lipid nanoparticles.
Porous hollow silica nanoparticles are being investigated as a carrier for controlled
release and for protection from UV degradation.  A handful of studies have determined that the
rate of release was influenced by temperature, pH, and shell thickness (M. Kah et al., 2013).
Additionally, the release of the active ingredient was not consistent possibly due to location of
the active ingredient within the silica nanoparticle.  The active ingredient could be released
externally, internally, or through pore channels.  Based on (Melanie Kah & Hofmann, 2014)’s
review, porous hollow silica nanoparticles are one of the least researched preparations, possibly
due to the biodegradable characteristics of other preparations.
Nanodispersions typically contain nanocrystals that have been suspended in liquid media
for use in the food and pharmaceutical industry.  Materials that are suspended can include
carotenoids, phytosterols, and natural antioxidants.  The goal of nanodispersions are to maximize
the surface area to increase the dissolution velocity and solubility saturation of poorly water
soluble active ingredients (M. Kah et al., 2013).  Although nanodispersions are growing within
the food and pharmaceutical industries, there has only been one study published on a pesticide
active ingredient—where the efficacy of the formulation was the same as conventional active
ingredient application (Elek et al., 2010).  More research is needed in this area to determine if
this type of preparation can effectively deliver active ingredient to the target organism than the
conventional formulation.
The last type of nanopesticide includes metals, metal complexes, and metals as carriers
for organic active ingredients.  There have been a few examples of pesticides like avermectin,
chlorpenafyr, and imidacloprid that are incorporated in polymer-based microcapsules which
contain nanoTiO2 or nanosilver.  Silica and calcium carbonate nanoparticles have also been
considered as options as carriers for organic active ingredients.  These formulations are designed
to promote the photocatalysis of the active ingredient after release, to reduce residues on plants
and in the soil (M. Kah et al., 2013).  Additionally, it is possible to generate nanoparticle-
pesticide complexes, as well as nanosized metal oxides with organic active ingredients.
However, there are still many unanswered questions as to whether application rates may be
compromised due to interactions within the formulation prior to application, as well as increased
production costs and unknown fate and toxicity (M. Kah et al., 2013).  Metals used as carriers for
organic active ingredients are not as prevalent in the literature as metal and metal oxide
nanoparticles alone—the latter has at least double the amount of literature published (Melanie
Kah & Hofmann, 2014).  Perhaps due to the trend toward biodegradable and more
environmentally-friendly pesticide alternatives, metals alone have been a more researched and
popular alternative to metals plus organic active ingredients.
Metal and metal oxide nanoparticles are the most prevalent nanopesticides after polymer-
based nanopesticides that have been researched.  Some examples of metal and metal oxide
nanoparticles include: titanium dioxide, silica, copper, aluminum, and silver nanoparticles, with
silver nanoparticles perhaps being the most well-studied and prevalent in numerous applications.
Copper, silver, and titanium dioxide are known antimicrobials, and silica and aluminum having
insecticidal properties.  Most of the data available with these metal nanoparticles support the
observations that the nano-sized particles are more effective than their bulk nanoparticles in
pesticide activity.  Some of these nanoparticles have other uses outside of their application
in the agricultural realm, such as titanium dioxide—and these uses were discussed in the
consumer products section.  The discussion about silver nanoparticles will be expanded upon
further in the next sub-section of this chapter, as the focus of the second chapter in my thesis
utilizes an alternative silver nanoparticle which aims to reduce the burden of silver released into
the environment.
Silicon has been used previously to reduce both abiotic and biotic stresses to plants, and
silica nanoparticles have therefore been recommended for pest control (Barik, Sahu, & Swain,
2008).  (Debnath et al., 2011) determined that silica nanoparticles (15-30nm) had greater efficacy
in killing insects than the bulk counterpart, even with different surface coatings.  However, the
application rates were the same as diatomaceous earth, which is used for the same purpose as
these silica nanoparticles, so the additional cost of engineering nanoparticles may not be
justified.  Alumna nanoparticles formulated as dust was also tested for insecticidal activity in the
same manner as the silica nanoparticles. The nanoalumna turned out to be as effective as
commercial formulations, and as efficacious as diatomaceous earth when tested on two types of
insects and under three humidity levels (Stadler, Buteler, Weaver, & Sofie, 2012).  However,
with nanoalumna, the mechanism(s) of action are unknown, so optimizing the formulation to
treat a range of insects under a range of environmental conditions is important for the future use
of this formulation (Melanie Kah & Hofmann, 2014).
Titanium dioxide is used for both antibacterial and antifungal purposes.  A study
evaluating the efficacy of nano titanium dioxide with either a zinc or silver coating was used to
combat the bacterial spot disease in tomatoes and roses (M. L. Paret, Vallad, Averett, Jones, &
Olson, 2013; M. Paret, Palmateer, & Knox, 2013).  The treatments ended up significantly
reducing the prevalence of the disease relative to controls, and was as effective as current
treatments.  The use of zinc and silver coatings on the titanium dioxide nanoparticles was
determined to be a more effective alternative to currently used copper coatings in terms of
toxicological and ecological risks (M. L. Paret et al., 2013; M. Paret et al., 2013).
(Mondal & Mani, 2012) reported that nanocopper could suppress the growth of bacterial
blight on pomegranate fruit by a magnitude of four times compared to conventionally used
copper oxychloride.  However, the details of the study were lacking so that these results could
not be compared to any other figure in the literature.  Following the trends of other metal
nanoparticles, it is possible that the use of the nanoformulation is a more effective alternative to
the bulk material.
As previously mentioned, background information about silver nanoparticles will be
discussed in more detail in the next section of this chapter.  Silver nanoparticles are discussed
more thoroughly than others as the second chapter of this thesis focuses on an alternative silver
nanoparticle, which aims to reduce the impact of silver on non-target organisms and the
environment.
1.4 Silver Nanoparticles: State of the Science
1.4.1 Use and Prevalence
Silver has been used as an antimicrobial for centuries in a variety of items.  The first
modern documentation of these properties was in 1869 when scientists discovered that
Aspergillus niger could not grow in silver vessels (Clement & Jarrett, 1994).  Additionally, silver
was used to fabricate cutlery and crockery which helped to prevent growth of bacteria and mold
and were used to prevent and treat infections prior to the advent of antibiotics (Bondarenko et al.,
2013; Bystrzejewska-Piotrowska et al., 2009).  With the knowledge of silver’s antimicrobial
properties, use of silver in modern times has grown significantly, primarily in their use as
nanoparticles.  As of 2015, silver nanoparticles were the most widely commercialized engineered
nanomaterial, and were incorporated into 23.5% of all reported consumer and medical products
(Vance et al., 2015).  Silver nanoparticles may be one of the most well-studied nanomaterials to
date, however there are still questions about the mode of toxic action, determining fate in
complex media, and the implications of their regulatory status.
As previously mentioned, silver nanoparticles are the most widely commercialized
engineered nanomaterial (as of 2015) as they have the widest range of applications and volume
of use.  The Consumer Products Inventory documents products with known uses of
nanomaterials, and those with silver nanoparticles include items such as: cosmetics, clothing,
shoes, detergents, dietary supplements, water filters, phones, and toys, among many others.
Silver nanoparticles are typically applied as a thin layer on the surface of these items, or can be
embedded into the material.  The estimated median global annual production of silver
nanoparticles is 55 tons, and use has risen steadily in the past decade (~52 new products/year);
global production is estimated to be between 12.2-1216 tons by the year 2020 (Piccinno et al.,
2012).  It has already been observed that silver can be released into the environment in both
liquid or solid forms from domestic and/or industrial sources, accidental spillages, and
atmospheric emissions (T. Benn, Cavanagh, Hristovski, Posner, & Westerhoff, 2010; T. M. Benn
& Westerhoff, 2008; Gottschalk et al., 2013; Kaegi et al., 2010; Mackevica, Olsson, & Hansen,
2016); therefore, models estimating predicted environmental concentrations and studies reporting
measured environmental concentrations of silver nanoparticles have been conducted.
1.4.2 Proposed Mechanisms of Action for Silver Nanoparticles
Silver nanoparticles have several suggested mechanisms of action, relating to both
released dissolved silver and the nanoparticle itself.  Toxic responses to silver nanoparticles have
been mainly documented for acute exposures as silver nanoparticles rapidly undergo
transformations in the aquatic environment, which will be discussed in the next sub-section of
this thesis.  When considering particle-specific mechanisms, this is not as well-studied as
mechanisms relating to dissolved silver, as discerning the impact of the particle versus the
dissolved silver is difficult.  In addition, we need to consider the difference between the rate of
dissolution of silver ions from the nanoparticle as a function of surface area as compared to the
rate of dissolution for conventional metal ions (Stone, Harper, Lynch, Dawson, & Harper, 2010).
(Juganson et al., 2015) outlines the three major proposed mechanisms of toxic action which are
applicable to most engineered nanomaterials, including silver nanoparticles: (1) physical
interactions with cells or cellular components, (2) production of reactive oxygen species and
resulting induction of oxidative stress, and (3) release of ions from metal/metal oxide
nanomaterials.
A useful way to visualize and compare toxicity findings of published studies for both
silver nanoparticles and dissolved silver is to construct a species sensitivity distribution (SSD).
SSDs display cumulative probability distributions of toxicity values on a logistic scale for
multiple species.  (Bondarenko et al., 2013) compiled LC50 values from several studies and
displayed the SSD for silver nanoparticles and for dissolved silver.  The SSDs can be viewed in
Figure 3, and show that for crustaceans, fish and algae, dissolved silver is an order of magnitude
more toxic than silver nanoparticles (Bondarenko et al., 2013).  For the data collected, some
median LC50 values varied significantly from the average LC50, which can be partially explained
by the composition of the test medium which can affect dissolution of silver from the
nanoparticle and the speciation of the silver ions (Bondarenko et al., 2013).  Additional
considerations for the difference includes modifications of surface of the particle, perhaps by
dissolved organic matter, or agglomeration events which may have led to reduction of dissolved
silver due to precipitation and sedimentation of the particle.
As silver nanoparticles can undergo dissolution to release silver ions in solution,
discussion of the mechanisms of toxic action for silver ions are pertinent.  In the literature, there
are several mechanisms of action that have been proposed, which will be presented here.  The
three possible mechanisms include (1) disruption of the ion-efflux system in cellular membranes
and a resulting increase in membrane permeability (2) interactions with thiol groups that can
inactivate important enzymes in the electron transport chain in cellular oxidation and (3)
denaturing of DNA and RNA which leads to DNA condensation and subsequent disruptions in
DNA replication and RNA translation.
When discussing the effects of silver exposure to aquatic organisms (particularly fish),
the disruption of the Na+/K+ ATPase pump by binding to sulfhydryl groups in the cell
membranes of the gills is a commonly cited mechanism of action.  This reduces the plasma
concentrations of Na+ and Cl, which can stress the organism and can lead to a disturbance in the
cardiovascular fluid volume which induces cardiovascular collapse and death.  In both (Brauner
& Wood, 2002; Wood, Hogstrand, Galvez, & Munger, 1996), they observed up to a 35%
reduction in whole-body Na+ when exposing developing rainbow trout to silver nitrate.  At the
time, it was speculated that this reduction would have an impact on the cardiovascular system,
but evidence of this was not provided until (Webb & Wood, 1998)’s study, where loss of ionic
regulatory function and therefore a measured net loss of ions in the plasma led to death of the
fish.  Decrease in the thickness of the gill filaments in zebrafish has also been observed
(Lapresta-Fernández et al., 2012).
Another proposed mechanism of action for silver is the interaction of silver with thiol
groups of proteins involved with the electron-transport chain in cellular oxidation as well as
those within the phospholipid bilayer.  In terms of cellular oxidation, the enzymes that were
inactivated by silver include succinate dehydrogenase and aconitase which are bound in the cell
membrane (Gordon et al., 2010).  Silver specifically bonded to the sulfhydryl (thiol) groups in
amino acids to promote the release of iron.  At the same time, it was observed that hydroxyl
radical formation occurred by an indirect mechanism likely mediated by reactive oxygen species
(Gordon et al., 2010).  These observations were made in Staphylococcus epidermidis, which is a
common bacterium found on the human body, but is also responsible for infections particularly
when invasive surgeries are performed.  Silver can also oxidize thiol groups in cell-wall proteins,
which results in destabilization.  This destabilization of the membrane can disrupt cell
homeostasis by decreasing the electrochemical gradient of the cell, which can deactivate energy-
dependent reactions (Lapresta-Fernández et al., 2012).
Silver can also impact DNA replication and RNA translation, which is critical for protein
synthesis.  (Feng et al., 2000) utilized two species of bacteria, one Gram-positive and one Gram-
negative.  When exposed to silver nitrate, similar observations were made for both bacteria: the
cell membrane shrank and detached from the cell wall, an electron-light region appeared in the
center of the cells, with condensed DNA molecules positioned in the center of it; and silver ions
were detected inside the cells (Feng et al., 2000).  The electron-light region near the nucleus of
the cell was thought to be a protective mechanism, as small molecular weight proteins are made
and surround the genetic material.  However, if there is enough silver, the formation of electron-
dense granules will occur and overwhelm the electron-light region—these electron-dense
granules contain the silver and will subsequently harm the genetic material (Feng et al., 2000).
When the DNA is in a condensed form, the DNA loses its replicating ability.  Additionally, the
DNA is not available for transcription in the protein synthesis process, which is severely
detrimental to the organism.  Furthermore, silver can directly bind to RNA polymerase, which
interferes with the transcription process of protein synthesis (Wang, Xia, & Liu, 2015).
In addition to toxicity elicited by ionic silver, the nanoparticle itself can also lead to toxic
responses.  Silver nanoparticles can induce oxidative damage at the cell membrane, as well as
bind to proteins within the cell membrane (ionic channels, receptors, and porins) and inactivate
enzymes (perioxdase, catalase, superoxide dismutase, and NADH dehydrogenase II in the
respiratory system) (Hwang et al., 2008; Lapresta-Fernández et al., 2012).  This can interfere
with the proton pool in the intermembrane space or the electron flow in the respiratory process,
which can then generate reactive oxygen species (ROS) such as superoxide (Hwang et al., 2008;
Lapresta-Fernández et al., 2012; Peijnenburg et al., 2015).  ROS can also reduce silver (I) ions;
this provides a pathway for continual generation of ROS and regeneration of silver nanoparticles
(Peijnenburg et al., 2015).  Another way silver nanoparticles can generate ROS is by surface
plasmon enhancement; free electrons within silver nanoparticles oscillate at the same frequency
as incident light photons which results in localized surface plasmon resonance, which forms
superoxide radicals which can be converted to ROS (Massarsky, Trudeau, & Moon, 2014).
1.4.3 Transformations of Silver Nanoparticles
Silver nanoparticles are subject to dynamic environmental conditions which impact their
fate in the environment.  Generally, the presence of humic and fulvic acids, pH, temperature, and
the ionic strength can all impact the bioavailability of silver nanoparticles to biological
organisms.  In particular, increasing levels of dissolved organic matter and increasing pH reduce
the rate of silver ion dissolution from silver nanoparticles whereas the rate of dissolution
increases with increasing temperature.  Additionally, aggregation and deposition of silver
nanoparticles will occur in acidic and/or high ionic strength environments, especially those with
high concentrations of divalent cations. Complexation is also an important factor, where
chloride, organic sulfur and nitrogen ligands present high binding affinity to silver ions.
Upon release into the environment, it has been suggested that leached silver nanoparticles
will first pass through sewage treatment plants, where most will precipitate in the sludge and a
minority will be present in the effluent to reach aquatic environments (Peijnenburg et al., 2015).
(Li, Hartmann, Döblinger, & Schuster, 2013) has estimated that after both mechanical and
biological treatment of wastewater, approximately 95% of silver nanoparticles have been
removed, and only 5% remain in the effluent (for a plant that processes 520,000 t/d, only 4.4 g/d
would be released into the effluent).  (Furtado et al., 2015) simulated a boreal lake ecosystem
and investigated the fate of both citrate and PVP-coated silver nanoparticles.  Although there was
not an effect on the persistence of the nanoparticles based on the different coatings, other studies
have shown that when the surface chemistry is modified, silver burden on the environment can
be significantly reduced (Ellis, Valsami-Jones, Lead, & Baalousha, 2016; Richter et al., 2015,
2016).  Their analysis did show that periphyton on the mesocosm wall and sediments were
important sinks for silver, and that the particles were stable in the water column perhaps due to
the low ionic strength and high dissolved organic carbon in the water.  This study illustrates that
water quality characteristics can impact silver nanoparticle behavior.
Salinity, dissolved organic matter, temperature, pH and ionic concentration (in particular)
can all transform silver nanoparticles.  Accumulation in compartments, dissolution, degradation,
and aggregation/agglomeration are all possible events that can occur—agglomeration and
dissolution are cited as the most likely and important transformations (Furtado et al., 2016).
Adapted Figure 4 from Furtado et al., 2016 illustrates the different transformations silver
nanoparticles can undergo and their effect on fate and toxicity in natural waters.  As the surface
of a silver nanoparticle is susceptible to reactions, adsorption of natural organic matter,
macromolecules, and reactions with oxygen and sulfur are common occurrences (Maurer-Jones
et al., 2013).  When the surface of a silver nanoparticle is oxidized, Ag2O forms, and will
dissolve to release Ag+ (Li et al., 2013).  When humic and fulvic acid concentrations are high,
aggregation of silver nanoparticles is reduced which enhances their mobility in water and
reduces dissolution (Settimio et al., 2015).  Additionally, the rate of silver ion release from silver
nanoparticles is dependent on temperature and pH; ions are released at a faster rate in increasing
water temperatures (0-37°C) and with decreasing pH (Lapresta-Fernández et al., 2012).  With the
multitude of factors that can impact silver bioavailability, it is unlikely that dissolved silver
would be present in large concentrations due to these processes.
The most important factors that contribute to the reduction of bioavailable dissolved silver
in natural waters is complexation and cation competition.  An early model of the silver biotic
ligand model (BLM) considered the key toxic sites on the gill of the rainbow trout and the
possible effects of cation competition (McGeer, Playle, Wood, & Galvez, 2000).  Their study
built in log K values (affinity constants) for calcium and sodium to better predict the impact of
cation competition.  Also, chloride, sulfide/sulfhydryl groups, and dissolved organic matter can
bind to silver and reduce the amount of dissolved silver available to interact with the gill surface.
For example, sulfidation can affect surface charge, dissolution rate and can reduce toxicity as
silver sulfide is less soluble (Levard, Hotze, Lowry, & Brown, 2012).  Chloride can complex
with silver and generate a precipitate of silver chloride, which removes silver ions from solution.
Since concentrations of complexable and competitive ions in natural waters vary, we can utilize
the BLM for silver to help understand the complexation reactions that may occur, to predict the
toxicity of metals for water with defined chemistries, and to set water quality criteria.
(Bielmyer et al., 2007) validated the predictive capabilities of the silver acute BLM
by testing eight different natural freshwaters with two species, an invertebrate and a fish
(Ceriodaphnia dubia and Pimephales promelas).  The waters collected differed significantly in
pH, ionic concentration, and water hardness to determine whether the BLM for silver accurately
predicted the LC50 for both organisms.  For the aquatic invertebrate C. dubia, the BLM made
reasonable predictions of silver toxicity (within a factor of 2) except in cases where the ionic
strength and water hardness were <35mg CaCO3/L.  For these cases, the model overpredicted
silver toxicity, but this is preferable as the model is conservative in nature.  However, this was
not the same case for P. promelas, as the model underestimated toxicity, particularly in waters
with low ionic strength.  As acclimation of the organism was accounted for and was found to not
influence silver toxicity to P. promelas, it is possible that physiological differences, such as
silver binding affinity to the gill might be the reason for the incorrect BLM prediction.  Also, the
authors suggest that the decreasing silver toxicity observed as the organisms’ size increased
could have been a result of the rate of ionic turnover.  This study demonstrated that the BLM for
silver predicted toxicity rather well, however, it can still be refined to include possible
physiological differences.
1.4.4 Regulation of Silver Nanoparticles
Through validation studies like these, we can add to the existing pool of data for a
metal and set water quality criteria.  Currently for silver, the Ambient Water Quality Criteria
(AWQC) in use was set in 1980 by the USEPA.  The primary factor considered in setting the
acute AWQC was water hardness and is represented by this equation:
e(1.72 ln⁡hardness-6.72)
As the water hardness is the only parameter we need to consider, determining the AWQC for
silver is relatively easy.  However, this model is too simplistic.  Only considering water hardness
when ionic composition and concentration of dissolved organic matter are known to have a
substantial impact on silver bioavailability is not adequate to estimate the toxicity of silver to
susceptible organisms.  These factors were alluded to in the (USEPA, 1980) document, but there
may have not been enough convincing evidence for these factors to be built into the AWQC at
that time.
Thirty-seven years later, the AWQC still has not changed, although use of silver in the
form of silver nanoparticles has grown exponentially.  As previously discussed, dissolution is a
major factor in the fate of silver nanoparticles in natural waters, so it is pertinent to include the
contribution of silver nanoparticles in the AQWC.  At the time of publishing, the EPA mentioned
that the major use of silver was in the photographic and dental industry, and was not considered a
significant pollutant (USEPA, 1980).  However, with the increasing use of silver nanoparticles in
a range of applications, we may want to revisit the acute AWQC for silver and to determine if
generating a chronic AWQC is appropriate.  Additionally, we will want to incorporate the
validation studies of the silver BLM and the relevant water quality parameters that affect silver
bioavailability.  However, the EPA deems the risk of silver nanoparticles adversely affecting
human health as low.  This is apparent in the lack of strict regulations on silver nanoparticles, no
mandatory reporting requirements by producers, and by evaluating potential risks on a case by
case basis.
As of 2014, nanomaterials are regulated without specific provisions in existing United
States legislation.  They are lumped with their bulk material counterparts, and can be regulated
as hazardous chemical substances or pesticides, under the Toxic Substances Control Act (TSCA)
or the Federal Insecticide, Fungicide, Rodenticide Act (FIFRA) (USEPA, 2014).  When used in
or as food additives, drugs, or cosmetics, nanomaterials are regulated under the Federal Food,
Drug, and Cosmetic Act (FFDCA) (USEPA, 2014; Vance et al., 2015).
For example, as silver nanoparticles are commonly used as antimicrobials, there is the
potential for them to be regulated under FIFRA and the FFDCA.  This was petitioned for by the
Institute for Agriculture and Trade Policy with the International Center for Technology
Assessment in 2008, but there were some issues with declaring silver nanoparticles as protected
under these acts.  The first is that protections would be grouped with bulk silver as silver
nanoparticles are not recognized as a distinct chemical.  The impact of the actual particle itself
would not be considered, which can be problematic, as the particle may have separate
toxicological impacts.  Although there have been calls to oversee research concerning the
commercial application of nanotechnologies and comprehensive labeling for products that
contain nanoparticles, this has not resulted in legislative action.  Second, if the recognized silver
nanoparticles as a pesticide, they would have to set a Maximum Residue Level (MRL) and be
forced to withdraw all silver nanoparticles from commerce as well as any products found to have
MRLs above the established limit (Suppan, 2015), which would be costly and logistically
difficult.  Instead, the EPA decided to consider each product on a case by case basis (USEPA,
2016).
To determine whether a nanomaterial (including silver) poses a risk to human health and
the environment, there are two new considerations that were included in the recent reform of
TSCA, which occurred on June 22nd, 2016.  The first consideration is a pre-manufacture
notification for new nanomaterials and chemicals.  This requires manufacturers of new chemical
substances to provide specific information to the EPA for review prior to manufacturing
chemicals or introducing them into commerce (USEPA, 2016).  That way, the EPA can take
action to ensure that chemicals that may or will pose an unreasonable risk to human health or the
environment are effectively controlled (USEPA, 2016).  Specific actions that the EPA has taken
under the pre-manufacture notification consideration include: limiting the use of nanomaterials,
requiring the use of personal protective equipment and engineering controls, limiting
environmental releases, and requiring testing to generate health and environmental effects data
(USEPA, 2016).
The information gathering rule is the second consideration that the EPA has adopted in
the recent reform of TSCA.  This rule applies to both new and existing nanomaterials, and
requires one-time reporting and recordkeeping of existing exposure and health and safety
information on nanoscale chemical substances in commerce (USEPA, 2016).  Companies that
manufacture or process nanoscale materials already in commerce must notify the EPA of the
chemical identity, production volume, manufacture methods, processing, use, exposure and
release information, and available health and safety data (USEPA, 2016).  This information will
be used to determine if further action is needed for a nanomaterial under TSCA, including
additional information collection, if needed (USEPA, 2016).
Besides FFDCA, TSCA, and FIFRA, silver nanoparticles (and others) may also be subject
to other regulatory laws depending on their location.  Nanomaterials can be regulated under the
Clean Water Act (CWA), Clean Air Act (CAA), and the Safe Water Drinking Act if there are
discharges, run-off or general contamination of air and (drinking) water.  Additionally, at waste
sites, nanomaterial risk can be evaluated and addressed by both the Comprehensive
Environmental Response, Compensation, and Liability Act (CERCLA) and the Resource
Conservation and Recovery Act (RCRA).  Lastly, Occupational Health and Safety
Administration (OSHA) standards may apply to nanomaterials, and in 25 states OSHA has
approved federal safety standards for nanomaterials in private industries (USEPA, 2014).  The
National Institute for Occupational Health and Safety (NIOSH) has also set recommended
exposure limits for a handful of nanomaterials, in addition to developing interim guidelines for
the health implications and applications of nanomaterials, recommended personal protection
guidelines and work practices (USEPA, 2014).
In contrast to the United States, the European Union takes a more precautionary approach
to regulating nanomaterials by requiring more stringent testing based on production volumes
instead of determining risk on a case-by-case basis.  Generally, nanomaterials are regulated
under the Registration, Evaluation, Authorization, and Restriction of Chemicals (REACH), and
the Classification, Labeling, and Packaging (CLP) legislation (ECHA, 2016).  Although there are
no explicit requirements for nanomaterials under REACH or CLP, they meet the regulations’
substance definition and therefore the provisions apply (ECHA, 2016).  Additionally, if biocidal
materials consist of nanoparticles, aggregates, or agglomerates in which at least 50% of primary
particles have at least one dimension between 1-100nm, the Biocidal Products Regulation (BPR)
provides further protections (ECHA, 2016; Vance et al., 2015).  Cosmetics that contain
nanomaterials are regulated by the European Commission.
According to REACH, the potential ecotoxicological effects of all chemicals that are
produced in a volume of more than one ton per year and sold in the EU must be evaluated by the
producer/importer (ECHA, 2016; Juganson et al., 2015).  These tests are required for each
nanomaterial before it enters the market is more extensive compared to US requirements.  All
tests are performed according to OECD standards.  Acute tests for aquatic invertebrates are
performed over a 48-hour period, acute algal tests assess growth inhibition over a 72-hour period
and acute testing for fish occurs over a 96-hour period.  Table 2 (following this chapter)
describes the tests required for production volumes under REACH.  In addition to these tests
which assess the ecotoxicological impacts of both chemicals and nanomaterials, there are
additional requirements which are only applicable to nanomaterials.  These include assessing the
dispersion conditions and thorough characterization of the particle in the test environment
(Bondarenko et al., 2013).  Based on these outcomes, nanomaterials will be classified with
respect to their toxicity according to the response of the most sensitive of the three organisms
tested, which is analogous to non-nano materials.
Overall, the regulation of nanomaterials (including silver) for the United States may fall
under several existing laws, however; there are no specific provisions written to address nano-
related impacts.  With the new revisions to TSCA, there may be more opportunity to review
nanomaterials as they are subject to the pre-manufacture notices and the information gathering
rule.  Under REACH in the European Union, nanomaterials are required to be regulated due to
the language of the law.  Additionally, there are some specific guidelines for nanomaterial
testing, which is important if we are to assess the risk of nanomaterials in the environment.
1.4.5 Applicability of my Research
To close this chapter, I will briefly discuss how my research described in Chapter 2 adds
to the body of knowledge on silver nanoparticles.  My research focuses on an alternative silver
nanoparticle which has been modified to be more ‘environmentally-friendly’ as compared to
conventional silver nanoparticles.  There are three elements to the modified nanoparticle that I
tested to discern their toxicity using the embryonic zebrafish model.  Instead of a solid silver
core which is typical of conventional silver nanoparticles, the modified particle has a lignin core.
The lignin core is recycled from a natural source, and is biodegradable.  Silver is bound to the
lignin at a concentration that has been optimized to limit extraneous silver release into the
surrounding environment, while still maintaining antimicrobial efficacy.  Surrounding the lignin-
core nanoparticle is a positively-charged surface stabilizer, which is abbreviated as PDAC.  In
addition to stabilizing the surface of the nanoparticle, PDAC also has antimicrobial properties
and acts as an attractant to the negatively charged cell membrane of bacteria.
As the use of silver nanoparticles in consumer products is growing at consistent rate, it
is expected that release of silver nanoparticles into the environment after use or disposal will
increase significantly.  After release into the environment, dissolution of silver from the
nanoparticle is possible and can potentially impact sensitive non-target aquatic organisms.  Not
only can silver ions adversely impact organisms, but the nanoparticle itself can also elicit harm.
This alternative nanoparticle may reduce the overall burden of released silver in the environment
as well as potential particle-related toxicity.  If the elements of the alternative particle do not
cause excessive toxicity to our model organism, it is possible that these particles can be utilized
in consumer products to provide the same level of antimicrobial action, while reducing unwanted
environmental effects.
The practice of utilizing green chemistry during nanomaterial design and synthesis is
becoming a common trend as we learn more of nanomaterial fate and toxicity.  The goal is to
reduce harmful effects to the environment and non-target organisms while still maintaining the
desired antimicrobial activity (in this case) during application.  The study I performed not only
evaluates the toxicity of the formulation components, but also makes the case for utilizing green
chemistry practices in nanomaterial design.  Our study indicates that replacing the silver core and
reducing the amount of available silver can potentially reduce the toxic burden that is associated
with conventional silver nanoparticles.  Additionally, our study indicates that the surface
stabilizer, which also has antimicrobial properties, was the most toxic component of the
formulation.  Due to this finding, our collaborators are investigating the use of an alternative
nanoparticle coating which is biologically-derived that may have the potential to be less toxic.
These findings add to the existing knowledge of how green chemistry practices can reduce the
impact associated with nanomaterial toxicity and fate, in addition to encouraging further research
in this area of nanomaterial design and synthesis.
 
 
 
 
 
 
 
 
 
 
 
 
 
1.6 Figures and Tables

 
Table 1. List of Naturally-Occurring Nanoparticles. This table describes where naturally occurring nanoparticles are found, the particle types and ecotoxicological potential of each nanoparticle. Figure is borrowed from (Handy et al., 2008).

Figure 1. Display of Important External Factors Which Impact the Fate of Nanomaterials.  This figure displays the numerous transformations and interactions that can occur in aquatic environments.  Figure is borrowed from (Peijnenburg et al., 2015).

Figure 2. Types of Nanopesticides Researched in the Literature up to October 2013.  This figure represents the growth in researched nanopesticides.  Typically, the knowledge of these different formulations correlates with the number of papers published on them.  Figure borrowed from (Melanie Kah & Hofmann, 2014).

Figure 3.  Species Sensitivity Distributions (SSDs) of both Silver Salt and Silver Nanoparticles.  These SSDs illustrate that there is a shift in the toxicity (an order of magnitude) from the silver salt (more toxic) to the silver nanoparticle (less toxic).  Aquatic invertebrates are the most sensitive, followed by algae, fish, nematodes, bacteria, fish and protozoa.  Figure borrowed from (Bondarenko et al., 2013).

Figure 4. Transformations and Interactions of Silver Nanoparticles in Aquatic Media.  This figure illustrates more specific interactions that can occur in the environment pertaining to silver nanoparticles.  Agglomeration, dissolution, sedimentation and complexation are all important processes that can occur based on the characteristics of natural waters.  Figure borrowed from (Furtado et al., 2016).

Volume (tons) Tests Required
1-10 Acute tests with aquatic invertebrates (Daphnia magna preferred) and plants (algae preferred)
10-100 The above tests plus acute tests with fish and activated sludge respiration
100-1000 All aforementioned studies including completing chronic studies with the same organisms; early life stage toxicity tests on fish (embryo and sac-fry), juvenile growth tests on fish; acute terrestrial tests for invertebrates and plants; determine effects on soil microorganisms
≥1000 Chronic terrestrial toxicity tests performed with invertebrates, plants, sediment organisms and birds in addition to all of the above tests

Table 2. Tests Required for Production Volumes of Nanomaterials Under REACH. 
 
1.5 References
 
Accuray Research LLP. (2016). Global Nanotechnology Market Analysis & Trends – Industry Forecast to 2025. Retrieved from http://www.researchandmarkets.com/research/3m2wkr/global
Albanese, A., Tang, P. S., & Chan, W. C. W. (2012). The Effect of Nanoparticle Size, Shape, and Surface Chemistry on Biological Systems. Annual Review of Biomedical Engineering14(1), 1–16. https://doi.org/10.1146/annurev-bioeng-071811-150124
Alfadul, S. M., & Elneshwy, A. A. (2010). Use of nanotechnology in food processing, packaging and safety–review. African Journal of Food, Agriculture, Nutrition and Development10(6). Retrieved from http://www.ajol.info/index.php/ajfand/article/view/58068
Anton, N., Benoit, J.-P., & Saulnier, P. (2008). Design and production of nanoparticles formulated from nano-emulsion templates—A review. Journal of Controlled Release128(3), 185–199. https://doi.org/10.1016/j.jconrel.2008.02.007
ASTM. (2012). Standard Terminology Relating to Nanotechnologies.
Barik, T. K., Sahu, B., & Swain, V. (2008). Nanosilica—from medicine to pest control. Parasitology Research103(2), 253–258. https://doi.org/10.1007/s00436-008-0975-7
Benn, T., Cavanagh, B., Hristovski, K., Posner, J. D., & Westerhoff, P. (2010). The Release of Nanosilver from Consumer Products Used in the Home. Journal of Environment Quality39(6), 1875. https://doi.org/10.2134/jeq2009.0363
Benn, T. M., & Westerhoff, P. (2008). Nanoparticle Silver Released into Water from Commercially Available Sock Fabrics. Environmental Science & Technology42(11), 4133–4139. https://doi.org/10.1021/es7032718
Bielmyer, G. K., Grosell, M., Paquin, P. R., Mathews, R., Wu, K. B., Santore, R. C., & Brix, K. V. (2007). Validation study of the acute biotic ligand model for silver. Environmental Toxicology and Chemistry26(10), 2241–2246.
Bondarenko, O., Juganson, K., Ivask, A., Kasemets, K., Mortimer, M., & Kahru, A. (2013). Toxicity of Ag, CuO and ZnO nanoparticles to selected environmentally relevant test organisms and mammalian cells in vitro: a critical review. Archives of Toxicology87(7), 1181–1200. https://doi.org/10.1007/s00204-013-1079-4
Brauner, C., & Wood, C. (2002). Effect of long-term silver exposure on survival and ionoregulatory development in rainbow trout (Oncorhynchus mykiss) embryos and larvae, in the presence and absence of added dissolved organic matter. Comparative Biochemistry and Physiology Part C133, 161–173.
Brunel, F., El Gueddari, N. E., & Moerschbacher, B. M. (2013). Complexation of copper(II) with chitosan nanogels: Toward control of microbial growth. Carbohydrate Polymers92(2), 1348–1356. https://doi.org/10.1016/j.carbpol.2012.10.025
BSI Standards Publication. (2011). Nanoparticles: Vocabulary.
Bystrzejewska-Piotrowska, G., Golimowski, J., & Urban, P. L. (2009). Nanoparticles: Their potential toxicity, waste and environmental management. Waste Management29(9), 2587–2595. https://doi.org/10.1016/j.wasman.2009.04.001
Chaudhry, Q., Scotter, M., Blackburn, J., Ross, B., Boxall, A., Castle, L., … Watkins, R. (2008). Applications and implications of nanotechnologies for the food sector. Food Additives & Contaminants: Part A25(3), 241–258. https://doi.org/10.1080/02652030701744538
Clement, J., & Jarrett, P. (1994). Antibacterial Silver. Metabolism Based Drugs1, 467–482.
Corley, E. A., Scheufele, D. A., & Hu, Q. (2009). Of risks and regulations: how leading U.S. nanoscientists form policy stances about nanotechnology. Journal of Nanoparticle Research11(7), 1573–1585. https://doi.org/10.1007/s11051-009-9671-5
Cushen, M., Kerry, J., Morris, M., Cruz-Romero, M., & Cummins, E. (2012). Nanotechnologies in the food industry – Recent developments, risks and regulation. Trends in Food Science & Technology24(1), 30–46. https://doi.org/10.1016/j.tifs.2011.10.006
Dawson, N. G. (2008). Sweating the Small Stuff: Environmental Risk and Nanotechnology. BioScience58(8), 690. https://doi.org/10.1641/B580805
Debnath, N., Das, S., Seth, D., Chandra, R., Bhattacharya, S. C., & Goswami, A. (2011). Entomotoxic effect of silica nanoparticles against Sitophilus oryzae (L.). Journal of Pest Science84(1), 99–105. https://doi.org/10.1007/s10340-010-0332-3
De Jong, W. H., Borm, P. J., & others. (2008). Drug delivery and nanoparticles: applications and hazards. International Journal of Nanomedicine3(2), 133.
ECHA. (2016). Regulations for Nanomaterials Under REACH. Retrieved from https://echa.europa.eu/regulations/nanomaterials
Elek, N., Hoffman, R., Raviv, U., Resh, R., Ishaaya, I., & Magdassi, S. (2010). Novaluron nanoparticles: Formation and potential use in controlling agricultural insect pests. Colloids and Surfaces A: Physicochemical and Engineering Aspects372(1-3), 66–72. https://doi.org/10.1016/j.colsurfa.2010.09.034
Ellis, L.-J. A., Valsami-Jones, E., Lead, J. R., & Baalousha, M. (2016). Impact of surface coating and environmental conditions on the fate and transport of silver nanoparticles in the aquatic environment. Science of The Total Environment568, 95–106. https://doi.org/10.1016/j.scitotenv.2016.05.199
FDA. (2014). Considering Whether an FDA-Regulated Product Involves the Application of Nanotechnology.
FDA. (2015). FDA’s Approach to Regulation of Nanotechnology Products. Retrieved from http://www.fda.gov/ScienceResearch/SpecialTopics/Nanotechnology/ucm301114.htm
Feng, Q. L., Wu, J., Chen, G. Q., Cui, F. Z., Kim, T. N., Kim, J. O., & others. (2000). A mechanistic study of the antibacterial effect of silver ions on Escherichia coli and Staphylococcus aureus. Journal of Biomedical Materials Research52(4), 662–668.
Furtado, L. M., Bundschuh, M., & Metcalfe, C. D. (2016). Monitoring the Fate and Transformation of Silver Nanoparticles in Natural Waters. Bulletin of Environmental Contamination and Toxicology. https://doi.org/10.1007/s00128-016-1888-2
Furtado, L. M., Norman, B. C., Xenopoulos, M. A., Frost, P. C., Metcalfe, C. D., & Hintelmann, H. (2015). Environmental Fate of Silver Nanoparticles in Boreal Lake Ecosystems. Environmental Science & Technology49(14), 8441–8450. https://doi.org/10.1021/acs.est.5b01116
Gordon, O., Vig Slenters, T., Brunetto, P. S., Villaruz, A. E., Sturdevant, D. E., Otto, M., … Fromm, K. M. (2010). Silver Coordination Polymers for Prevention of Implant Infection: Thiol Interaction, Impact on Respiratory Chain Enzymes, and Hydroxyl Radical Induction. Antimicrobial Agents and Chemotherapy54(10), 4208–4218. https://doi.org/10.1128/AAC.01830-09
Gottschalk, F., Sonderer, T., Scholz, R. W., & Nowack, B. (2009). Modeled environmental concentrations of engineered nanomaterials (TiO2, ZnO, Ag, CNT, fullerenes) for different regions. Environmental Science & Technology43(24), 9216–9222.
Gottschalk, F., Sun, T., & Nowack, B. (2013). Environmental concentrations of engineered nanomaterials: Review of modeling and analytical studies. Environmental Pollution181, 287–300. https://doi.org/10.1016/j.envpol.2013.06.003
Handy, R. D., Owen, R., & Valsami-Jones, E. (2008). The ecotoxicology of nanoparticles and nanomaterials: current status, knowledge gaps, challenges, and future needs. Ecotoxicology17(5), 315–325. https://doi.org/10.1007/s10646-008-0206-0
Han, Y., Li, L., Hao, W., Tang, M., & Wan, S. (2013). Larvicidal activity of lansiumamide b from the seeds of Clausena lansium against Aedes albopictus (Diptera: Culcidae). Parasitol Res112, 511–516.
Harper, S. L., Carriere, J. L., Miller, J. M., Hutchison, J. E., Maddux, B. L. S., & Tanguay, R. L. (2011). Systematic Evaluation of Nanomaterial Toxicity: Utility of Standardized Materials and Rapid Assays. ACS Nano5(6), 4688–4697. https://doi.org/10.1021/nn200546k
Helmut Kaiser Consultancy. (2009). Summary about the State of Nanotechnology Industry Worldwide 2006-2015. Retrieved from http://www.hkc22.com/nanomarkets.html
Hwang, E. T., Lee, J. H., Chae, Y. J., Kim, Y. S., Kim, B. C., Sang, B.-I., & Gu, M. B. (2008). Analysis of the Toxic Mode of Action of Silver Nanoparticles Using Stress-Specific Bioluminescent Bacteria. Small4(6), 746–750. https://doi.org/10.1002/smll.200700954
Juganson, K., Ivask, A., Blinova, I., Mortimer, M., & Kahru, A. (2015). NanoE-Tox: New and in-depth database concerning ecotoxicity of nanomaterials. Beilstein Journal of Nanotechnology6, 1788–1804. https://doi.org/10.3762/bjnano.6.183
Kaegi, R., Sinnet, B., Zuleeg, S., Hagendorfer, H., Mueller, E., Vonbank, R., … Burkhardt, M. (2010). Release of silver nanoparticles from outdoor facades. Environmental Pollution158(9), 2900–2905. https://doi.org/10.1016/j.envpol.2010.06.009
Kah, M., Beulke, S., Tiede, K., & Hofmann, T. (2013). Nanopesticides: State of Knowledge, Environmental Fate, and Exposure Modeling. Critical Reviews in Environmental Science and Technology43(16), 1823–1867. https://doi.org/10.1080/10643389.2012.671750
Kah, M., & Hofmann, T. (2014). Nanopesticide research: Current trends and future priorities. Environment International63, 224–235. https://doi.org/10.1016/j.envint.2013.11.015
Keiper, A. (2003). The nanotechnology revolution. The New Atlantis, (2), 17–34.
Kessler, R. (2011). Engineered nanoparticles in consumer products: understanding a new ingredient. Environ Health Perspect119(3), A120–A125.
Lapresta-Fernández, A., Fernández, A., & Blasco, J. (2012). Nanoecotoxicity effects of engineered silver and gold nanoparticles in aquatic organisms. TrAC Trends in Analytical Chemistry32, 40–59. https://doi.org/10.1016/j.trac.2011.09.007
Lee, K. T. (2010). Quality and safety aspects of meat products as affected by various physical manipulations of packaging materials. Meat Science86(1), 138–150. https://doi.org/10.1016/j.meatsci.2010.04.035
Levard, C., Hotze, E. M., Lowry, G. V., & Brown, G. E. (2012). Environmental Transformations of Silver Nanoparticles: Impact on Stability and Toxicity. Environmental Science & Technology46(13), 6900–6914. https://doi.org/10.1021/es2037405
Li, L., Hartmann, G., Döblinger, M., & Schuster, M. (2013). Quantification of Nanoscale Silver Particles Removal and Release from Municipal Wastewater Treatment Plants in Germany. Environmental Science & Technology, 130620163339004. https://doi.org/10.1021/es3041658
Liu, X., Tang, Harper, Steevens, J., Xu, R., & Harper. (2013). Predictive modeling of nanomaterial exposure effects in biological systems. International Journal of Nanomedicine, 31. https://doi.org/10.2147/IJN.S40742
Mackevica, A., Olsson, M. E., & Hansen, S. F. (2016). The release of silver nanoparticles from commercial toothbrushes. Journal of Hazardous Materials. https://doi.org/10.1016/j.jhazmat.2016.03.067
Massarsky, A., Trudeau, V. L., & Moon, T. W. (2014). Predicting the environmental impact of nanosilver. Environmental Toxicology and Pharmacology38(3), 861–873. https://doi.org/10.1016/j.etap.2014.10.006
Maurer-Jones, M. A., Gunsolus, I. L., Murphy, C. J., & Haynes, C. L. (2013). Toxicity of Engineered Nanoparticles in the Environment. Analytical Chemistry85(6), 3036–3049. https://doi.org/10.1021/ac303636s
Maynard, A. (2006). Nanotechnology: A Research Strategy for Addressing Risk.
McEvoy, M. (2015). Policy Memorandum for Nanotechnologies in Organic Foods.
McGeer, J. C., Playle, R. C., Wood, C. M., & Galvez, F. (2000). A Physiologically Based Biotic Ligand Model for Predicting the Acute Toxicity of Waterborne Silver to Rainbow Trout in Freshwaters. Environmental Science & Technology34(19), 4199–4207. https://doi.org/10.1021/es9912168
Meredith, A. N., Harper, B., & Harper, S. L. (2016). The influence of size on the toxicity of an encapsulated pesticide: a comparison of micron- and nano-sized capsules. Environment International86, 68–74. https://doi.org/10.1016/j.envint.2015.10.012
Mondal, K. K., & Mani, C. (2012). Investigation of the antibacterial properties of nanocopper against Xanthomonas axonopodis pv. punicae, the incitant of pomegranate bacterial blight. Annals of Microbiology62(2), 889–893. https://doi.org/10.1007/s13213-011-0382-7
National Nanotechnology Initiative. (2016). Nanoparticle and Nanotechnology General Descriptions. Retrieved from http://www.nano.gov/
Nel, A., Xia, T., Mädler, L., & Li, N. (2006). Toxic potential of materials at the nanolevel. Science311(5761), 622–627.
Nguyen, H. M., Hwang, I. C., Park, J. W., & Park, H. J. (2012). Enhanced payload and photo-protection for pesticides using nanostructured lipid carriers with corn oil as liquid lipid. Journal of Microencapsulation29(6), 596–604. https://doi.org/10.3109/02652048.2012.668960
Paret, M. L., Vallad, G. E., Averett, D. R., Jones, J. B., & Olson, S. M. (2013). Photocatalysis: effect of light-activated nanoscale formulations of TiO2 on Xanthomonas perforans and control of bacterial spot of tomato. Phytopathology103(3), 228–236.
Paret, M., Palmateer, A., & Knox, G. (2013). Evaluation of a Light-activated Nanoparticle Formulation of Titanium Dioxide with Zinc for Management of Bacterial Leaf Spot on Rosa “Noare.” HortScience48(2), 189–192.
Peijnenburg, W. J. G. M., Baalousha, M., Chen, J., Chaudry, Q., Von der kammer, F., Kuhlbusch, T. A. J., … Koelmans, A. A. (2015). A Review of the Properties and Processes Determining the Fate of Engineered Nanomaterials in the Aquatic Environment. Critical Reviews in Environmental Science and Technology45(19), 2084–2134. https://doi.org/10.1080/10643389.2015.1010430
Piccinno, F., Gottschalk, F., Seeger, S., & Nowack, B. (2012). Industrial production quantities and uses of ten engineered nanomaterials in Europe and the world. Journal of Nanoparticle Research14(9). https://doi.org/10.1007/s11051-012-1109-9
Ramachandraiah, K., Han, S. G., & Chin, K. B. (2015). Nanotechnology in Meat Processing and Packaging: Potential Applications–A Review. Asian Austrlasian Journal of Animal Science28(2), 290–302.
Rashidi, L., & Khosravi-Darani, K. (2011). The Applications of Nanotechnology in Food Industry. Critical Reviews in Food Science and Nutrition51(8), 723–730. https://doi.org/10.1080/10408391003785417
Richter, A. P., Bharti, B., Armstrong, H. B., Brown, J. S., Plemmons, D., Paunov, V. N., … Velev, O. D. (2016). Synthesis and Characterization of Biodegradable Lignin Nanoparticles with Tunable Surface Properties. Langmuir32(25), 6468–6477. https://doi.org/10.1021/acs.langmuir.6b01088
Richter, A. P., Brown, J. S., Bharti, B., Wang, A., Gangwal, S., Houck, K., … Velev, O. D. (2015). An environmentally benign antimicrobial nanoparticle based on a silver-infused lignin core. Nature Nanotechnology10(9), 817–823. https://doi.org/10.1038/nnano.2015.141
Royal Commission of Environmental Pollution. (2008). Novel Materials in the Environment: The Case of Nanotechnology.
Sargent, J. (2016, September 15). Nanotechnology: A Policy Primer.
Settimio, L., McLaughlin, M. J., Kirby, J. K., Langdon, K. A., Janik, L., & Smith, S. (2015). Complexation of silver and dissolved organic matter in soil water extracts. Environmental Pollution199, 174–184. https://doi.org/10.1016/j.envpol.2015.01.027
Shin, S., Song, I., & Um, S. (2015). Role of Physicochemical Properties in Nanoparticle Toxicity. Nanomaterials5(3), 1351–1365. https://doi.org/10.3390/nano5031351
Stadler, T., Buteler, M., Weaver, D. K., & Sofie, S. (2012). Comparative toxicity of nanostructured alumina and a commercial inert dust for Sitophilus oryzae (L.) and Rhyzopertha dominica (F.) at varying ambient humidity levels. Journal of Stored Products Research48, 81–90. https://doi.org/10.1016/j.jspr.2011.09.004
Stone, D., Harper, B., Lynch, I., Dawson, K., & Harper, S. (2010). Exposure Assessment: Recommendations for Nanotechnology-Based Pesticides. International Journal of Occupational and Environmental Health16, 467–474.
Suppan, S. (2015). Suing EPA for failure to regulate nano-pesticides. Retrieved from http://www.iatp.org/blog/201504/suing-epa-for-failure-to-regulate-nano-pesticides-0
USEPA. (1980). Ambient Water Quality Criteria for Silver.
USEPA. (2014). Technical Fact Sheet–Nanomaterials.
USEPA. (2016). Control of Nanoscale Materials under the Toxic Substances Control Act. Retrieved from https://www.epa.gov/reviewing-new-chemicals-under-toxic-substances-control-act-tsca/control-nanoscale-materials-under
Vance, M. E., Kuiken, T., Vejerano, E. P., McGinnis, S. P., Hochella, M. F., Rejeski, D., & Hull, M. S. (2015). Nanotechnology in the real world: Redeveloping the nanomaterial consumer products inventory. Beilstein Journal of Nanotechnology6, 1769–1780. https://doi.org/10.3762/bjnano.6.181
Wang, Z., Xia, T., & Liu, S. (2015). Mechanisms of nanosilver-induced toxicological effects: more attention should be paid to its sublethal effects. Nanoscale7(17), 7470–7481. https://doi.org/10.1039/C5NR01133G
Webb, N. A., & Wood, C. M. (1998). Physiological analysis of the stress response associated with acute silver nitrate exposure in freshwater rainbow trout (Oncorhynchus mykiss). Environmental Toxicology and Chemistry17(4), 579–588.
Wood, C. M., Hogstrand, C., Galvez, F., & Munger, R. S. (1996). The physiology of waterborne silver toxicity in freshwater rainbow trout (Oncorhynchus mykiss) 1. The effects of ionic Ag+. Aquatic Toxicology35(2), 93–109.
Yin, Y., Guo, Q., Han, Y., Wang, L., & Wan, S. (2012). Preparation, characterization of nematicidal actuvity of lansiumamide b nano-capsules. Journal of Integrative Agriculture11(7), 1151–1158.
 
 
CHAPTER 2: TOXICOLOGICAL ASSESSMENT OF A LIGNIN CORE
NANOPARTICLE DOPED WITH SILVER AS AN ALTERNATIVE TO
CONVENTIONAL SILVER NANOPARTICLES
Cassandra E. Nix, Bryan J. Harper, Cathryn G. Conner, Alexander P. Richter, Orlin D. Velev& Stacey L. Harper
2.1 Abstract
Elevated levels of silver in the environment have been detected due to an increase in silver nanoparticle (AgNP) use in consumer products.  To potentially reduce the burden of continuous silver ion release typical of conventional AgNPs, a lignin-core particle doped with silver ions and surface stabilized with a polycationic electrolyte layer was engineered.  Our objective was to determine whether any of the formulation components elicit toxicological responses using embryonic zebrafish.  Ionic silver and free surface stabilizer were the most toxic constituents, although when associated separately or together with the lignin core, toxicity of the formulations decreased significantly.  Formulations containing silver had a significantly higher prevalence of uninflated swim bladder and yolk sac edema.  Comparative analysis of dialyzed samples, which intended to simulate post-consumer use, showed a significant increase in mortality as the samples aged, in addition to eliciting significant increases in types of sub-lethal responses relative to the non-dialyzed samples.  ICP-OES/MS analysis indicated that silver ion release from the particle into solution was continuous and the rate of release was component-specific.  Overall, our study indicates that the lignin core is an effective alternative to conventional AgNPs for potentially reducing the burden of silver released into the environment.
2.2 Keywords
Nanotechnology, environmentally-friendly, pesticide, antimicrobial, zebrafish
 
 
 
2.3 Introduction
Silver nanoparticles (AgNPs) are an effective antimicrobial agent and the most widely commercialized engineered nanomaterial, incorporated into half of all reported consumer and medical products in the Nanotechnology Consumer Products Inventory.1  Prominent examples include cosmetics, clothing, shoes, detergents, water filters, phones, laptops and toys.2-4  AgNP use has risen steadily in the past decade (~52 new products per year), and global production is estimated to be between 12.6-1216 tons per year by 2020.5-6  With the increasing production and use of AgNPs, the fate and the subsequent release of silver in nanomaterial and ionic form into the environment are of concern.  However, by applying the principles of green chemistry during nanomaterial design and synthesis, harmful effects to the environment can be limited while maintaining the desired antimicrobial activity during application.7
Research indicates that AgNPs can enter aqueous environments from discharges at the point of production, by erosion from household products, and from disposal of silver-containing products.8-12  These studies have prompted the investigation of AgNP interactions in the environment,13 particularly aquatic systems, to determine which general intrinsic and extrinsic properties are important in determining fate.10,14-16  Extrinsic properties include environmental factors and processes that can impact the fate of the particles in aquatic systems such as pH, temperature and composition of the water, as well as processes like sedimentation, deposition, dissolution, agglomeration, and/or particle sulfidation.17-20  Intrinsic factors address inherent particle characteristics such as size, shape, chemical composition, surface structure and surface charge.13, 21-24  Extrinsic factors can interact with intrinsic features of nanoparticles to alter particle behavior with concomitant effects on properties such as the bioavailability of AgNPs to living organisms; thus, a more comprehensive understanding is needed.14,25
AgNPs are known to be toxic to many aquatic organisms including algae, bacteria, invertebrates, and fish.2  Several mechanisms of action have been proposed, mainly attributing the toxicity of AgNPs to silver ions released from the nanoparticle.  However, nanoparticle-specific mechanisms are also being investigated, with data suggesting that mechanistic differences exist compared to dissolved silver.5,26  Silver ion specific mechanisms include: interactions with thiols and electron donor groups which can impact enzymes and DNA which makes them unavailable for cellular processes,27-29 denaturing of DNA and RNA which ultimately affects protein synthesis,30-31 and production of superoxide radicals and other reactive oxygen species via reactions with oxygen.30  Particle-specific mechanisms have been suggested that focus on the ability of AgNPs to cause cell membrane damage, leading to disruption in the ion efflux system in cells.32-33  Since multiple aquatic organisms may be at risk due to the increasing prevalence of silver in the environment, it is important to consider ways to reduce the environmental silver burden related to AgNPs.
To reduce potential silver release into the environment and impacts to non-target organisms while maintaining the same antimicrobial efficacy, a silver-doped lignin nanoparticle was engineered.34  The core material is composed of lignin, which was chosen as it is a natural biodegradable biopolymer.35  Similar synthesized lignin nanoparticles have been shown to have no impact on algae and yeast survival, suggesting they have a high level of biocompatibility.36 The lignin is easily precipitated into nano-sized particles using environmentally-friendly solvents, and the resulting nanoparticles can be infused with the minimum amount of silver ions needed for antimicrobial efficacy.  The particles are then surface-functionalized with a polycationic electrolyte layer to stabilize the particle, as well as to provide additional antimicrobial impact.  The lignin nanoparticles exhibit both high and low affinity binding regions for silver ions, and these differing affinities as well as the electrostatic barrier provided by the surface stabilizer impact the rate of release of silver ions into the surrounding solution.34  It is expected that the low affinity binding sites will primarily release the majority of the weakly bound silver in the first 24 hours;34 however, we also wanted to investigate the long-term release from the high affinity binding sites, so two of the formulated samples were dialyzed to remove the weakly bound silver.  When compared to their non-dialyzed counterparts, this allowed us to determine whether there are any differences in toxicological responses after the conclusion of the consumer stage of their life cycle and to quantify the potential environmental burden of silver from these particles.
Our aim was to elucidate which aspects of the formulation contribute most to the toxicity of the formulation and to discover whether these nanoparticles exhibit any toxicity post-consumer use.  We hypothesized that silver ions and the surface stabilizer are the main contributors to the aquatic toxicity of these nanoparticles and that once the particles have been diluted in solution and released their silver, there would be a reduction in toxicity.  To test these hypotheses, we utilized the embryonic zebrafish assay which is a widely-used model for toxicity testing as it provides a suite of developmental endpoints that are critical to the survival of the organism.37-38  Zebrafish also develop quickly and are optically transparent which allows for easy observations of phenotypic responses.37  Additionally, they share similar homology to humans, so observed effects of chemical stressors from this assay can potentially be extrapolated to human physiological responses.39
2.4 Materials and Methods
2.4.1 Materials and Characterization
Reference component solutions of silver nitrate (AgNO3) salt (CAS# 7761-88-8, Fisher Scientific) at 50 mg/L of Ag+ dissolved in ultrapure water and polydiallyldimethylammonium chloride (PDAC) (MW 100,000 – 200,000, CAS# 26062-79-3, Sigma Aldrich) at 200 mg/L in untrapure water were prepared and refrigerated at 4 °C until use.  The lignin (Indulin AT) for the nanoparticle core was extracted from biomass as a by-product of Kraft pulping processes.35,40  The Indulin AT lignin powder (lot MB05) and supporting documentation were obtained from MeadWestVaco, SC.  The size range of the particles after synthesis was 84 ± 5 nm.34  Stock nanomaterial suspensions of the lignin nanoparticle (NP), the silver functionalized lignin nanoparticle (NP+Ag), the silver functionalized particle with the cationic PDAC surface (NP+Ag+PDAC), and the lignin nanoparticle with PDAC alone (NP+PDAC) were prepared as previously described.34-35  Seven-fold dilutions of stock nanomaterial suspensions were performed with fishwater to prepare the varied exposure solutions.  Fishwater was prepared by dissolving 260 mg/L Instant Ocean salts (Aquatic Ecosystems, Apopka, FL) in reverse osmosis water and adjusting pH to 7.2 ± 0.2 using ~0.1 g sodium bicarbonate (conductivity 480-600 S/cm).38  Experimental materials were stored under the same conditions as the reference materials.  The NP+Ag and NP+PDAC formulations were solely used for comparative purposes, whereas the NP+Ag+PDAC is the proposed complete product formulation.
The samples to be dialyzed (NP+Ag and NP+Ag+PDAC) were placed in deionized water for 24 hours which included a Slide-A-Lyzer MINI Dialysis Device (Thermo Scientific, Waltham, MA) with a 10K molecular weight cutoff membrane to remove dissolved silver from solution prior to dilution and testing.  A second sample of NP+Ag was also dialyzed and allowed to age for 5 months prior to testing.  Thus, the dialyzed samples included NP+Ag Aged, NP+Ag Fresh, and NP+Ag+PDAC Fresh with the “Fresh” and “Aged” designations referring to when the sample was tested relative to when it was dialyzed.
The hydrodynamic diameter (HDD) and the zeta potential of each formulation that contained particles were measured in triplicate using a Zetasizer Nano ZS (Malvern Instruments Ltd., Worcestershire, UK) at 26.8 °C after dilution with fishwater to 50 mg/L.  Aliquots (1 mL) were stored in an incubator under the same conditions as the embryonic zebrafish until ready for analysis.  Measurements were made over a five-day period which also included an initial measurement (Day 0) which correlates with the exposure time of the experiment.
2.4.2 Embryonic Zebrafish Assay
Exposure solutions of reference and nanomaterial suspensions were dispensed into 96-well plates with each row having 12 wells of a given concentration of test material.  Each well was filled with 200 L of test solution and one of the eight rows on the plate was reserved for fishwater alone as a control.  Adult zebrafish (Danio rerio) were maintained at the Sinnhuber Aquatic Research Laboratory (SARL) at Oregon State University.  Embryos received from SARL were approximately 6 to 8 hours post-fertilization (hpf) and were inspected under a dissecting microscope to ensure viability and developmental stage, then placed individually into wells of a 96-well plate.  The chorionic membrane surrounding the zebrafish was preserved.  Two replicate exposures were conducted over two weeks for each material, which allowed us to have a total sample size of 24 fish per concentration, per material.  After plating, the exposure wells were sealed with Parafilm to reduce evaporation, and embryos were incubated at 26.8 °C under a 14:10 light:dark photoperiod.
2.4.3 Toxicological Evaluations of Embryonic Zebrafish
Fish were observed at 24 hpf and 120 hpf for mortality as well as a suite of developmental, morphological and physiological abnormalities.  At 24 hpf, embryos were evaluated for mortality, spontaneous movement, delayed developmental progression, and notochord malformations.  At 120 hpf, mortality was evaluated in conjunction with malformations of the snout, brain, pectoral and caudal fin, eye, jaw, otic structures, axis, trunk, somites, swim bladder and body pigmentation.  In addition, physiological and behavioral endpoints evaluated at 120 hpf include the presence of pericardial or yolk sac edema, impaired circulation and active touch response.38  Hatching success was measured between 48 and 120 hpf, with embryos that hatched between 48 and 72 hpf being considered normal, and any individuals hatching after 72 hpf were considered delayed.41  All endpoints were reported as either absent or present.  Representative images of control fish and any individuals that displayed developmental abnormalities at 24 and 120 hpf were taken with an Olympus SZX10 microscope (Tokyo, Japan) fitted with an Olympus SC100 high resolution digital color camera (Olympus Corporation, Center Valley, PA) and are included in the Supplemental Information (SI, Figure S1).  All experiments were performed in compliance with national care and use guidelines and approved by the Institutional Animal Care and Use Committee (IACUC) at Oregon State University (ACUP #4764).
2.4.4 Measurement of Dissolved Silver and Particle-Associated Silver
Both the concentration of dissolved silver released from the nanoparticles and the silver associated with the particle itself were quantified by Inductively Coupled Plasma-Optical Emission Spectroscopy or Mass Spectrometry (ICP-OES or ICP-MS).  To quantify silver content in solution, acid digestion of particles was performed using established methods with known recoveries for silver nanoparticles.19,42  Triplicate 0.5 mL samples of stock suspensions were centrifuged at 13,000 ×g for 10 minutes in a 3 kDa centrifugal filter (VWR, Radnor, PA) with a polyethersulfone (PES) membrane to separate the lignin particles from the filtrate.  A total of 0.45 mL of filtrate sample was collected, diluted 10-fold with ultrapure water, and adjusted with 70% trace-metal grade HNO3 to a final concentration of 3% HNO3.  For the lignin particle digestion, 0.1 mL of stock solution was digested in the same manner as the filtrate samples, without the centrifugation step.  All samples were digested in Teflon tubes at 200 °C with 3 mL 70% trace-metal grade HNO3.  The acid was allowed to completely evaporate and the process was repeated 3 times.  Final digested samples were dissolved in 5 mL of 3% HNO3 prior to ICP-OES/MS analysis.  The silver ICP standard was purchased from RICCA Chemical Company (Ricca Chemical Company, Arlington, TX) and diluted to six concentrations spanning the expected concentrations.  All samples, including standards, were analyzed in triplicate with ICP-OES (Teledyne Leeman Labs, Hudson, NH) for silver content except the filtrate from the NP+Ag+PDAC sample which was analyzed by ICP-MS (Thermo-Fisher, Waltham, MA) to provide a lower level of detection (≥5 µg/L).
2.4.5 Statistical Analyses
All differences were considered significant at p ≤ 0.05, and statistical analyses were conducted with SigmaPlot 13.0 (Systat Software, San Jose, CA) unless otherwise noted.  For measurements of zeta potential and HDD, significant differences were determined with one-way analysis of variance (ANOVA) and a Tukey’s post-hoc analysis.  The post-hoc analysis was used to detect significant differences between the initial measurements and each of the measurements made over the five days.  Two-way ANOVA was conducted to ensure that there was no significant difference in mortality between replicate exposure plates prior to pooling of the data.  The concentration of the material and a grouping variable were designated as the explanatory variables, and mortality was the response variable.   Concentration-response curves were generated with the Environmental Protection Agency’s Toxicity Relationship Analysis Program (EPA TRAP v. 1.30, March 2015) and compared using two-way ANOVA.  EPA TRAP was also used to calculate the concentration at which fifty percent of exposed zebrafish perished (LC50) for each material, and the Litchfield/Wilcoxon formula was utilized for LC50 comparisons between treatments.43   Significant sub-lethal endpoints were determined by Fisher’s Exact Test by comparing the control (fishwater alone) response to each concentration response tested.  To determine whether there were significant differences in the concentration of silver associated with the particle versus the filtrate between the comparable formulations in the ICP analysis, one-way ANOVA was performed.
2.5 Results and Discussion
2.5.1 Particle Characterization
Average zeta potential and HDD measurements for the formulation components and dialyzed formulations were plotted over a five-day period and are illustrated in the SI, Figure S2.  No significant changes were observed in the zeta potential measurements of NP+Ag+PDAC or NP+Ag+PDAC Fresh formulations.  However, for NP, NP+PDAC, NP+Ag Aged and the NP+Ag Fresh formulations, there were significant fluctuations relative to Day 0 in zeta potential over the last three days, although the values were within the range of moderate stability, edging on incipient aggregation.  For HDD measurements, no significant changes were observed for the NP+Ag+PDAC Fresh sample.  For the other dialyzed samples, significant differences were measured at the first and second day, as well as the fourth and fifth day in the NP+Ag Aged sample.  For the formulation components, significant differences were detected in all samples after the first 24 hours.  However, for all samples analyzed, the sizes fluctuated between 70 and 100 nm, but did not increase or decrease consistently.  The only sample that increased consistently in size was the NP+Ag sample, which ranged from approximately 80 to 120 nm over 5 days in fishwater.
2.5.2 Analysis of Dissolved Silver and Particle-Associated Silver
The concentration of silver in solution and the silver associated with the particles was quantified in the five nanoparticle samples that included silver.  Figure 1 shows the concentration of silver present in the particle and in solution which, when combined, matches the nominal concentration provided for each material.  In all cases, the silver associated with the particle was greater than the silver present in solution (1.62 to 132 times greater), which was expected as the particle is designed to leach the silver ions.  The full formulation (NP+Ag+PDAC) contains approximately 11 times more silver associated with the particle than the dialyzed full formulation (NP+Ag+PDAC Fresh).  Additionally, the age of the NP+Ag sample played a role in silver distribution, as the older formulation contained approximately 10 times more silver in the filtrate than the particle.
Previous analyses of similar particles by Richter and colleagues34 found that the concentration of silver associated with the particle after dialysis was approximately 18%; however, in this study we found much higher concentrations associated with the particle (61.7 to 99.2%).  This may have been due to variance between batches of the lignin nanoparticle stock solutions, as well as the differences in digestion techniques.  Although the dialysis process is effective at releasing the loosely bound silver from the particles, the electrostatic barrier in the samples that contained PDAC may have impacted the rate at which silver was released.  The full formulation had the lowest release of dissolved silver, although there was significantly more silver released from the freshly dialyzed full formulation sample, suggesting that PDAC may retard the release of ionic silver by repulsive electrostatic interactions.  Additionally, through previous characterization of the lignin particle functional groups by Richter and colleagues,34-35 there is a higher proportion of organically bound sulfur compared to other lignin types (9 times that of high-purity lignin), which would likely provide strong binding sites for dissolved silver.44
2.5.3 Comparative Analysis of Formulation Toxicity
No significant differences were found between replicate exposure plates; therefore, replicates were pooled to increase the sample size to 24 fish per concentration, per material.  To encompass all possible comparisons but for clarity in interpreting the data, two groupings were made which parallel our hypotheses.  These two groupings are formulation comparisons and dialyzed sample comparisons.  Concentration-response curves for the two groupings are illustrated in the SI (Figure S3).  Additionally, a modeled concentration-response curve was generated for the reference material silver nitrate, which is included in the SI (Figure S4).
2.5.3.1 Formulation Components
As represented in Figure 2a, the lignin core nanoparticle (NP) itself was the least toxic component (LC50 = 323 mg/L), and when the NP was combined with the other aspects of the formulation, LC50 values decreased significantly in all cases (NP+Ag LC50 = 200 mg/L, NP+PDAC LC50 = 33 mg/L, NP+Ag+PDAC LC50 = 32 mg/L).  The presence of silver in the full formulation (NP+Ag+PDAC) did not change the overall toxicity relative to NP+PDAC.  Additionally, when PDAC was present in the formulation (NP+PDAC or NP+Ag+PDAC), a significant increase in mortality events occurred.  PDAC and Ag+ alone were the two most toxic constituents, with LC50’s of 5.39 mg/L and 1.53 mg/L, respectively.
This LC50 for Ag+ is greater than many published literature values for zebrafish,46-50 however; exposure time and conditions differ in these studies which may explain the observed differences in toxicity.  Our zebrafish embryos were exposed at 8 hpf with the chorion intact, whereas some of the referenced studies did not expose the fish until after hatching, or even as adults.  The chorionic membrane can modulate silver toxicity by sequestering ions to prevent them from entering the perivitelline fluid;51 and removing the chorion has been shown to increase toxic responses.39,51  It is likely that the presence of the chorion may have played a large role in modulating silver toxicity, but the exposure media may have also played a role as well.
The hardness of our prepared fishwater may have altered in the toxicity of silver nitrate as well as nanoparticle-containing formulations to the embryonic zebrafish.  Based on dissolved magnesium and calcium concentrations, our fishwater is categorized as soft water (<60 mg/L CaCO3), whereas many of the above studies utilize moderately hard to hard water when exposing their zebrafish (up to 148 mg/L CaCO3).  It has been reported that LC50’s tend to be higher in the presence of dissolved organic matter, which has the greatest effect on silver toxicity, followed by Cl, Na+ and Ca2+.52  This is based on the coalescence effect, which leads to complexation and/or formation of nanoparticle agglomerates and/or aggregates, which can increase apparent toxicity due to difficult uptake.53  As our fishwater was categorized as soft (low concentrations of Na+ and Ca2+), we then determined the concentration of chloride ions present, as silver ion bioavailability can be impacted due to complexation and subsequent precipitation of silver chloride.25,47
In the Instant Ocean salt formulation used to make the fishwater, the majority of the cations are paired with chloride and we determined the chloride concentration in our fishwater to be 142 mg/L, which is approximately 55% of the dissolved ion content.  It is possible that when the fishwater was used to dilute the silver-containing treatments, the silver complexed with the chloride and precipitated out of solution, which may have led to a greater LC50 value.  To determine whether precipitation was a significant factor, we utilized Visual MINTEQ (v.3.1) for each silver-containing formulation (SI, Figure S5).  We used the filtrate concentrations from the ICP data for each of the formulations that contained silver, and determined that nearly all of the dissolved silver (98.2 to 99.8%) would complex with the chloride present in the fishwater to form a precipitate of silver chloride, except the NP+Ag+PDAC sample, as the concentration of Ag+ in solution was very low.  However, as the exposures occurred over a five-day period, and the movement of silver from the particle to the surrounding solution is a dynamic process, dissolved silver could have still been bioavailable to the zebrafish.
Considering our nanoparticle and silver combination formulations, comparisons to published LC50 values for conventional AgNPs differ significantly.  Reported AgNP LC50’s for fish generally range between 0.05 and 20 mg/L.2,54  Variations in reported LC50 values may relate to differences in the type of exposure, exposure time, age of the fish, the presence of the chorionic membrane, the use of bare or coated nanoparticles, and/or differences in exposure media.  A study completed by Bar-Ilan and colleagues55 matched our exposure conditions most consistently and had reported LC50’s within the range above, although the LC50’s differed significantly from our findings.  They exposed embryonic zebrafish to different sizes of colloidal AgNPs (3-100 nm) and found a range of LC50’s of 10.1 to 14.7 mg/L.  Although the sizes of the particles, length and timing of the exposure, and retention of the chorionic membrane is consistent with our experiment, our LC50’s were an order of magnitude less toxic.  For example, the measured LC50 for the NP+Ag formulation was 200 mg/L.  The difference may be due to the use of the lignin particle, which was shown via ICP-OES/MS analysis to retain bound silver rather than it all being released as silver ions into solution (Figure 1).  Therefore, potential exposure to silver ions would be reduced in the presence of the lignin particles, leading to an apparent increase in the LC50.  With replacing the silver core typical of conventional AgNPs, the concentration of silver released to the environment may be reduced, which was one of the goals of formulating the nanoparticle with a lignin core.
PDAC alone was the second most toxic component tested, and formulations that contained PDAC were significantly more toxic than formulations that did not contain PDAC (Figure 2a).  Although PDAC is a high charge density cationic polymer commonly used as a flocculant/coagulant in wastewater treatment, it is also cited as a cytotoxin that interacts with cell membranes to elicit cell damage, and eventually necrosis.56-57  Our results correspond with this literature finding, as embryos that were exposed to PDAC alone progressively blackened and disintegrated, starting at 5.75 mg/L.  Other formulations that contained PDAC did not elicit this response, perhaps because the PDAC is electrostatically associated with the lignin particle, or was “complexed”.  Free, or “non-complexed” PDAC, can interact with blood components such as erythrocytes and plasma proteins, cell membranes, and extracellular matrix proteins leading to undesired side effects not seen with complexed polycations.58  Our experimental observations support this, as we see that the un-complexed PDAC sample is indeed more toxic than any formulation that contains a nanoparticle-PDAC complex (Figure 2a).  Additionally, research suggests polycationic polymers like PDAC can disrupt the lipid bilayer, with larger polymers leading to the formation of holes in the lipid membrane that increased membrane permeability.59-60  Given that information, it is possible that PDAC made the fish more susceptible to dissolved silver in the formulation as a result of changes in membrane permeability.  The positively charged PDAC-coated particles (SI, Figure S2) may have also been attracted to the negatively-charged membranes of the zebrafish, which could have increased the exposure to silver associated with the particle.
2.5.3.2 Dialyzed Formulations
The purpose of dialyzing the samples was to simulate post-consumer use of the nanoparticle by purging the surrounding solution of excess silver ions.  In the LC50 comparisons of dialyzed materials, two results can be observed (Figure 2b).  First, there was no difference in LC50 between the dialyzed complete formulation (NP+Ag+PDAC Fresh) and its non-dialyzed counterpart (NP+Ag+PDAC).  This may be due to similar nominal concentrations of silver, as calculated by adding the silver associated with the particle and silver present in the filtrate (Figure 1).  Second, the NP+Ag samples showed significant differences in toxicity post-dialysis, with LC50 values increasing immediately after dialysis (NP+Ag Fresh, LC50 = 244mg/L) then decreasing over time (Figure 2b), indicating that there was an increase in toxicity with increasing age of the dialyzed sample (NP+Ag LC50 = 200 mg/L and NP+Ag Aged LC50 = 139 mg/L).  This increase in toxicity correlates with the increased amount of silver released from the aged sample (Figure 1).  Perhaps over time, the higher affinity binding sites release more silver into solution compared to the freshly dialyzed sample, which led to a decrease in the calculated LC50.
2.5.4 Analysis of Sub-Lethal Endpoints
There were several endpoints that were significant (p ≤ 0.05) which included uninflated swim bladder and snout malformations (morphological), delay in hatching and developmental progression (developmental), and impaired circulation, yolk sac edema and pericardial edema (physiological).  Concentrations at which sub-lethal endpoints significantly increased relative to control fish within each material tested are designated with an asterisk in Figure 3a-g.  Overall, there is a significant increase in the types of sub-lethal responses observed in the dialyzed samples compared to the non-dialyzed samples.  The dialyzed samples, particularly the aged sample, has proportionally less silver associated with the particle than the non-dialyzed samples, which could explain the increase in sub-lethal impacts, as well as the lack of sub-lethal endpoints for NP+Ag+PDAC (Figure 1).

b
a

Swim bladder malformations occurred when silver or PDAC were included in the formulation; however, this did not occur in the complete formulation (NP+Ag+PDAC) except when freshly dialyzed (Figure 3a).  Exposure to silver ions during embryonic zebrafish development has been described as impacting cholinergic signaling, which is important in swim bladder formation.60  Swim bladder malformations were not significant in the NP+Ag+PDAC sample, likely due to high ratio of silver associated with the particle as compared to the filtrate (Figure 1).
Malformations of the snout were only significant in the freshly dialyzed full formulation. Although others have reported silver nanoparticles causing snout malformations,54 we do not see this malformation in silver nitrate or any other formulation containing silver, suggesting some other mechanisms of malformation may be involved.   Perhaps the presence of all three formulation components may have contributed to the prevalence of snout malformations, in addition to the increased concentration of silver in the filtrate as it exceeded the other treatments by a factor of six (Figure 1), except for the aged dialyzed particle (NP+Ag Aged).
Yolk sac edema was present in all fish exposed to formulations containing silver, and impaired circulation was significant in both freshly dialyzed formulations (Figure 3b). Significant pericardial edema responses were only noted in the NP+Ag and NP+Ag+PDAC Fresh formulations.  Similar to our findings, several other studies have reported that as a result of silver nanoparticle exposure, pericardial edema, yolk sac edema and impaired circulation are prevalent in the early developmental stages of zebrafish.62-65  The development of the circulatory system and the formation of the heart typically occur between 21 and 24 hpf in zebrafish embryos,41,64 thus it is likely that the release of dissolved silver from the nanoparticles resulted in these endpoints being prevalent, however particle specific responses cannot be ruled out.  The chorion has been reported to modulate metal toxicity,51 but there has been recent evidence that nanoparticles (30 to 72 nm) can move through the chorionic membrane canals, then distribute to numerous parts of the fish, including the brain, heart, yolk, and blood.66  Due to the distribution of silver in the yolk and heart, edemas commonly occur due to disturbances in osmoregulation.55,62-63,67  Once these developmental pathways are disturbed during early development, normal embryogenesis can be impacted, resulting in numerous defects.68  In addition, silver nanoparticles may agglomerate, which can clog chorion pore canals, and affect osmotic balance, which could then result in edemas.66
Pericardial edema coupled with impaired circulation following AgNP exposure has been shown to be concentration dependent,63 with an increase in prevalence from 10-100 mg/L.  This was the trend we observed in our study, although our study saw increases up to 125 mg/L with a maximum of 17% responding.  The most reasonable explanation for the slight difference in observations may be due to differences in the nanoparticle surface coatings on the tested particles and the binding sites on the particle core.  Polyvinyl alcohol was used as the surface coating in Asharani and colleagues,63 which may have altered the dynamics of silver ion release relative to our samples.  Our samples contained PDAC, which provides an electrostatic barrier that could impede silver ion release, which we did see in our samples (Figure 1); however, when freshly dialyzed, PDAC had limited impact on silver ion release, suggesting solution equilibrium may be controlled by PDAC.  Additionally, although there was silver present in the filtrate of the freshly dialyzed samples, the lignin core bound the majority of the silver, probably due to the higher binding affinity sites.
Significant delay in hatching and delayed developmental progression were the two developmental abnormalities observed in our study (Figure 3c).  As delayed developmental progression was found following exposure to the bare lignin nanoparticle and no other samples, the bare particle may interact with necessary ions in solution making them limited for supporting embryo development.  This is further supported by the finding that this process does not occur when the particle is functionalized with silver, PDAC or both, as the binding sites on the lignin particle are already occupied.  Silver nitrate was the only material tested that led to a delay in hatching.  A delay in hatching is primarily caused by deactivation of the ZHE1 enzyme, which prevents chorionic degradation.69  Lin and colleagues69 have shown that dissolved metals can interfere with ZHE1, and although silver was included in their assay, it did not lead to a significant decrease in ZHE1 activity.  Asharani and colleagues62 exposed zebrafish to Ag+ at 2.14 mg/L, and observed a delay in hatching at 4% compared to controls, but this was not significant.  Approximately 11% of our fish exhibited a delay in hatching compared to controls, however; Kimmel and colleagues41 suggested that delay in hatching does not impact the rate of development compared to individuals that hatched within 72 hpf.
The results of this study provide several insights into a nanoparticle engineered to be an environmentally friendly alternative to conventional silver nanoparticles.  Our data shows that the use of lignin as the nanoparticle core could be a viable alternative as it did not pose a significant toxicological hazard to our test organism.  Ionic silver and PDAC alone were the most toxic components of the formulation, which may be attributed to their higher diffusivity and propensity to interact with cell membranes rather than silver and/or PDAC associated with the particle.  The inclusion of PDAC not only adds antimicrobial activity to the particle, but also seems to delay the release of silver ions, so in situations where time-release of antimicrobial agents is desired, stabilizing the particles with PDAC may be warranted.  This data also encourages further development of similar nanomaterials to minimize their impact on the environment.  One way of reducing the environmental impact of these engineered nanomaterials is to design them in a way to minimize the release of soluble components or to replace these components with less toxic ingredients.  We are presently investigating the use of an alternative nanoparticle coating which is biologically-derived that may have the potential to be less toxic in comparison to PDAC.
 
 
 
 
 
 
 
 
 
 
 
 
2.6 Figures

 
 
 
 
 
 
Figure 1. Concentration of silver associated with the filtrate and particle.  ICP analysis determined the actual concentration of silver in the silver-containing formulations.  All samples were analyzed via ICP-OES except for the filtrate in the NP+Ag+PDAC sample, which was analyzed with ICP-MS.  Each sample had three experimental replicates, with three machine replicates for each sample.  There was significance in both the [Ag+] associated with the particle and the [Ag+] in the filtrate between all samples in comparable groups (two and three component formulations).

Figure 2. LC50’s for formulation components (a) and dialyzed samples (b).  Significant differences between LC50 values are indicated with a change in letter above the bar.  Standard error bars are represented for each material (n = 24).



 
 
 
Figure 3. Percent of zebrafish exhibiting significant sub-lethal responses.  Endpoints were deemed significant at p ≤ 0.05.  Each concentration within each tested formulation had an n = 24.  There was a significant increase in the types of sub-lethal responses observed in the freshly dialyzed samples (snout and pericardial edema), and exposure to silver-containing formulations increased prevalence of swim bladder and yolk sac edema.
 
 
 
 
 
 
 
 
 
 
 
 
2.7 Supplemental Information

Figure S1. Representative images of zebrafish with and without significant developmental impacts.


 
 
Figure S2. Average zeta potential and hydrodynamic diameter (HDD) measurements for particle-containing formulations over a five-day period.  Significant differences were determined by one-way ANOVA with a Tukey’s post-hoc analysis and are designated with an asterisk.  Day 1-5 measurements were compared to the initial measurement (Day 0).  Figures S1a-b are average zeta potential measurements for the formulation components (a) and the dialyzed formulations (b), and figures S1c-d are the average HDD measurements for the formulation components (c) and the dialyzed formulations (d).  Although some measurements were significantly different from the initial measurements, there was not a consistent trend of increasing or decreasing zeta potential or HDD, except for the NP+Ag formulation in Figure S1c.
Table S1. Metadata Associated with Zeta Potential Measurements. 

Model used to compute the zeta potential Smoluchowski equation
pH 7.2
Ionic strength 260 mg/L
Ionic composition Made per reference 38
Temperature 26.8 °C
Viscosity 0.8508 mPa
Macromolecules/NOM None
Duration of Measurement 1 minute
Applied voltage 148 V
Number of instrument measurements made and averaged to determine ZP 12
Total number of replicate measurements 3

Figure S3. Concentration-response comparisons for formulation components (a) and dialyzed materials (b) based on zebrafish mortality at 120 hpf (a)Significant differences (p ≤ 0.05) existed between materials in the formulation component treatments.  The lignin particle exhibited the lowest toxicity, followed by silver, and PDAC.  (b) No significant differences (p > 0.05) existed in the dialyzed sample treatments.  Comparisons included the two full formulations (NP+Ag+PDAC) and the three NP+Ag formulations.

 
 
 
 
Figures S4. Modeled concentration-response curve for the reference material silver nitrate based on zebrafish mortality at 120 hpf.  Silver nitrate is the source of the silver ions for the particle.  Curve was modeled in EPA TRAP (v.1.30, March 2015).

 
 
 
 
 
 
 
 
 
 
 
 
 
 
 
Figure S5. Visual MINTEQ output for all silver-containing formulations.  Highlighted cells represent the dissolved silver and the silver that may precipitate out due to the composition of the fishwater.

  1. NP+Ag Aged

 

 
 
 
 
 
 
 
 
 
 
 
 
 
 
 

  1. NP+Ag Fresh
  1. NP+Ag+PDAC


  1. NP+Ag+PDAC Fresh
  1. NP+Ag

 

 
2.8 References
1. Vance, M. E., Kuiken, T., Vejerano, E. P., McGinnis, S. P., Hochella, M. F., Rejeski, D., & Hull, M. S. (2015). Nanotechnology in the real world: Redeveloping the nanomaterial consumer products inventory. Beilstein Journal of Nanotechnology6, 1769–1780.
https://doi.org/10.3762/bjnano.6.181
2. Bondarenko, O., Juganson, K., Ivask, A., Kasemets, K., Mortimer, M., & Kahru, A. (2013). Toxicity of Ag, CuO and ZnO nanoparticles to selected environmentally relevant test organisms and mammalian cells in vitro: a critical review. Archives of Toxicology87(7), 1181–1200. https://doi.org/10.1007/s00204-013-1079-4
3. Bystrzejewska-Piotrowska, G., Golimowski, J., & Urban, P. L. (2009). Nanoparticles: Their potential toxicity, waste and environmental management. Waste Management29(9), 2587–2595. https://doi.org/10.1016/j.wasman.2009.04.001
4. Marambio-Jones, C., & Hoek, E. M. V. (2010). A review of the antibacterial effects of silver nanomaterials and potential implications for human health and the environment. Journal of Nanoparticle Research12(5), 1531–1551. https://doi.org/10.1007/s11051-010-9900-y
5. Massarsky, A., Trudeau, V. L., & Moon, T. W. (2014). Predicting the environmental impact of nanosilver. Environmental Toxicology and Pharmacology38(3), 861–873. https://doi.org/10.1016/j.etap.2014.10.006
6. Piccinno, F., Gottschalk, F., Seeger, S., & Nowack, B. (2012). Industrial production quantities and uses of ten engineered nanomaterials in Europe and the world. Journal of Nanoparticle Research14(9). https://doi.org/10.1007/s11051-012-1109-9
7. Anastas, P., & Eghbali, N. (2010). Green Chemistry: Principles and Practice. Chem. Soc. Rev.39(1), 301–312. https://doi.org/10.1039/B918763B
8. Benn, T., Cavanagh, B., Hristovski, K., Posner, J. D., & Westerhoff, P. (2010). The Release of Nanosilver from Consumer Products Used in the Home. Journal of Environment Quality39(6), 1875. https://doi.org/10.2134/jeq2009.0363
9. Benn, T. M., & Westerhoff, P. (2008). Nanoparticle Silver Released into Water from Commercially Available Sock Fabrics. Environmental Science & Technology42(11), 4133–4139. https://doi.org/10.1021/es7032718
10. Gottschalk, F., Sun, T., & Nowack, B. (2013). Environmental concentrations of engineered nanomaterials: Review of modeling and analytical studies. Environmental Pollution181, 287–300. https://doi.org/10.1016/j.envpol.2013.06.003
11. Kaegi, R., Sinnet, B., Zuleeg, S., Hagendorfer, H., Mueller, E., Vonbank, R., … Burkhardt, M. (2010). Release of silver nanoparticles from outdoor facades. Environmental Pollution158(9), 2900–2905. https://doi.org/10.1016/j.envpol.2010.06.009
12. Mackevica, A., Olsson, M. E., & Hansen, S. F. (2016). The release of silver nanoparticles from commercial toothbrushes. Journal of Hazardous Materials. https://doi.org/10.1016/j.jhazmat.2016.03.067
13. Dobias, J., & Bernier-Latmani, R. (2013). Silver Release from Silver Nanoparticles in Natural Waters. Environmental Science & Technology47(9), 4140–4146. https://doi.org/10.1021/es304023p 43.
14. Handy, R. D., Owen, R., & Valsami-Jones, E. (2008). The ecotoxicology of nanoparticles and nanomaterials: current status, knowledge gaps, challenges, and future needs. Ecotoxicology17(5), 315–325. https://doi.org/10.1007/s10646-008-0206-0
15. Selck, H., Handy, R. D., Fernandes, T. F., Klaine, S. J., & Petersen, E. J. (2016). Nanomaterials in the aquatic environment: A European Union-United States perspective on the status of ecotoxicity testing, research priorities, and challenges ahead: Nanomaterials in the aquatic environment. Environmental Toxicology and Chemistry35(5), 1055–1067. https://doi.org/10.1002/etc.3385
16. Maurer-Jones, M. A., Gunsolus, I. L., Murphy, C. J., & Haynes, C. L. (2013). Toxicity of                             Engineered Nanoparticles in the Environment. Analytical Chemistry85(6), 3036–3049.               https://doi.org/10.1021/ac303636s
17. Furtado, L. M., Norman, B. C., Xenopoulos, M. A., Frost, P. C., Metcalfe, C. D., & Hintelmann, H. (2015). Environmental Fate of Silver Nanoparticles in Boreal Lake Ecosystems. Environmental Science & Technology49(14), 8441–8450. https://doi.org/10.1021/acs.est.5b01116
18. Furtado, L. M., Bundschuh, M., & Metcalfe, C. D. (2016). Monitoring the Fate and Transformation of Silver Nanoparticles in Natural Waters. Bulletin of Environmental Contamination and Toxicology. https://doi.org/10.1007/s00128-016-1888-2
19. Kim, K.-T., Truong, L., Wehmas, L., & Tanguay, R. L. (2013). Silver nanoparticle toxicity in the embryonic zebrafish is governed by particle dispersion and ionic environment. Nanotechnology24(11), 115101. https://doi.org/10.1088/0957-4484/24/11/115101
20. Peijnenburg, W. J. G. M., Baalousha, M., Chen, J., Chaudry, Q., Von der kammer, F., Kuhlbusch, T. A. J., … Koelmans, A. A. (2015). A Review of the Properties and Processes Determining the Fate of Engineered Nanomaterials in the Aquatic Environment. Critical Reviews in Environmental Science and Technology45(19), 2084–2134. https://doi.org/10.1080/10643389.2015.1010430
21. Nel, A., Xia, T., Mädler, L., & Li, N. (2006). Toxic potential of materials at the nanolevel. Science311(5761), 622–627.
22. Sharma, V. K., Siskova, K. M., Zboril, R., & Gardea-Torresdey, J. L. (2014). Organic-coated silver nanoparticles in biological and environmental conditions: Fate, stability and toxicity. Advances in Colloid and Interface Science204, 15–34. https://doi.org/10.1016/j.cis.2013.12.002
23. Shin, S., Song, I., & Um, S. (2015). Role of Physicochemical Properties in Nanoparticle Toxicity. Nanomaterials5(3), 1351–1365. https://doi.org/10.3390/nano5031351
24. Lacave, J. M., Retuerto, A., Vicario-Parés, U., Gilliland, D., Oron, M., Cajaraville, M. P., &               Orbea, A. (2016). Effects of metal-bearing nanoparticles (Ag, Au, CdS, ZnO, SiO2) on               developing zebrafish embryos. Nanotechnology27(32), 325102.               https://doi.org/10.1088/0957-4484/27/32/325102
25. Groh, K. J., Dalkvist, T., Piccapietra, F., Behra, R., Suter, M. J.-F., & Schirmer, K. (2015). Critical influence of chloride ions on silver ion-mediated acute toxicity of silver nanoparticles to zebrafish embryos. Nanotoxicology9(1), 81–91. https://doi.org/10.3109/17435390.2014.893379
26. Ivask, A., ElBadawy, A., Kaweeteerawat, C., Boren, D., Fischer, H., Ji, Z., … Godwin, H. A. (2014). Toxicity Mechanisms in Escherichia coli Vary for Silver Nanoparticles and Differ from Ionic Silver. ACS Nano8(1), 374–386. https://doi.org/10.1021/nn4044047
27. Clement, J., & Jarrett, P. (1994). Antibacterial Silver. Metabolism Based Drugs1, 467–482.
28. Gordon, O., Vig Slenters, T., Brunetto, P. S., Villaruz, A. E., Sturdevant, D. E., Otto, M., … Fromm, K. M. (2010). Silver Coordination Polymers for Prevention of Implant Infection: Thiol Interaction, Impact on Respiratory Chain Enzymes, and Hydroxyl Radical Induction. Antimicrobial Agents and Chemotherapy54(10), 4208–4218. https://doi.org/10.1128/AAC.01830-09
29. Morones, J. R., Elechiguerra, J. L., Camacho, A., Holt, K., Kouri, J. B., Ramírez, J. T., & Yacaman, M. J. (2005). The bactericidal effect of silver nanoparticles. Nanotechnology16(10), 2346–2353. https://doi.org/10.1088/0957-4484/16/10/059
30. Feng, Q. L., Wu, J., Chen, G. Q., Cui, F. Z., Kim, T. N., Kim, J. O., & others. (2000). A mechanistic study of the antibacterial effect of silver ions on Escherichia coli and Staphylococcus aureusJournal of Biomedical Materials Research52(4), 662–668.
31. Fong, J., & Wood, F. (2006). Nanocrystalline silver dressings in wound management: a review. International Journal of Nanomedicine1(4), 441–449. https://doi.org/10.2147/nano.2006.1.4.441
32. Hwang, E. T., Lee, J. H., Chae, Y. J., Kim, Y. S., Kim, B. C., Sang, B.-I., & Gu, M. B. (2008). Analysis of the Toxic Mode of Action of Silver Nanoparticles Using Stress-Specific Bioluminescent Bacteria. Small4(6), 746–750. https://doi.org/10.1002/smll.200700954
33. Sharma, V. K., Yngard, R. A., & Lin, Y. (2009). Silver nanoparticles: Green synthesis and their antimicrobial activities. Advances in Colloid and Interface Science145(1-2), 83–96. https://doi.org/10.1016/j.cis.2008.09.002
34. Richter, A. P., Brown, J. S., Bharti, B., Wang, A., Gangwal, S., Houck, K., … Velev, O. D. (2015). An environmentally benign antimicrobial nanoparticle based on a silver-infused lignin core. Nature Nanotechnology10(9), 817–823. https://doi.org/10.1038/nnano.2015.141
35. Richter, A. P., Bharti, B., Armstrong, H. B., Brown, J. S., Plemmons, D., Paunov, V. N., … Velev, O. D. (2016). Synthesis and Characterization of Biodegradable Lignin Nanoparticles with Tunable Surface Properties. Langmuir32(25), 6468–6477. https://doi.org/10.1021/acs.langmuir.6b01088
36. Frangville, C., Rutkevičius, M., Richter, A. P., Velev, O. D., Stoyanov, S. D., & Paunov, V. N. (2012). Fabrication of Environmentally Biodegradable Lignin Nanoparticles. ChemPhysChem13(18), 4235–4243. https://doi.org/10.1002/cphc.201200537
37. Hill, A. J. (2005). Zebrafish as a Model Vertebrate for Investigating Chemical Toxicity. Toxicological Sciences86(1), 6–19. https://doi.org/10.1093/toxsci/kfi110
38. Truong, L., Harper, S., & Tanguay, R. (2011). Evaluation of Embryotoxicity Using the Zebrafish Model. In Drug safety evaluation: methods and protocols (pp. 271–279). New York, NY: Humana Press.
39. Kim, K.-T., & Tanguay, R. L. (2014). The role of chorion on toxicity of silver nanoparticles in the embryonic zebrafish assay. Environmental Health and Toxicology29, e2014021. https://doi.org/10.5620/eht.e2014021
40. Duval, A., & Lawoko, M. (2014). A review on lignin-based polymeric, micro- and nano-structured materials. Reactive and Functional Polymers85, 78–96. https://doi.org/10.1016/j.reactfunctpolym.2014.09.017
41. Kimmel, C. B., Ballard, W. W., Kimmel, S. R., Ullmann, B., & Schilling, T. F. (1995). Stages of embryonic development of the zebrafish. Developmental Dynamics203(3), 253–310. https://doi.org/10.1002/aja.1002030302
42. Wu, F., Harper, B. J., & Harper, S. L. (2017). Differential dissolution and toxicity of surface               functionalized silver nanoparticles in small-scale microcosms: impacts of community               complexity. Environ. Sci.: Nano. https://doi.org/10.1039/C6EN00324A
43. Sprague, J., & Fogels, A. (1977). Watch the Y in Bioassay (Environmental Protection Service Technical Report No. EPS-5-AR-77-1) (pp. 107–118). Procedural 3rd Aquatic Toxicology Workshop, Halifax, Nova Scotia, Canada.
44. Bielmyer, G. K., Grosell, M., Paquin, P. R., Mathews, R., Wu, K. B., Santore, R. C., & Brix, K. V. (2007). Validation study of the acute biotic ligand model for silver. Environmental Toxicology and Chemistry26(10), 2241–2246.
45. Lahnsteiner, F. (2008). The effect of internal and external cryoprotectants on zebrafish               (Danio rerio) embryos. Theriogenology69(3), 384–396.               https://doi.org/10.1016/j.theriogenology.2007.10.007
46. Alsop, D., & Wood, C. M. (2011). Metal uptake and acute toxicity in zebrafish: Common mechanisms across multiple metals. Aquatic Toxicology105(3-4), 385–393. https://doi.org/10.1016/j.aquatox.2011.07.010
47. Bielmyer, G., Brix, K., & Grosell, M. (2008). Is Cl protection against silver toxicity due to chemical speciation?



Recommendation
EssayHub’s Community of Professional Tutors & Editors
Tutoring Service, EssayHub
Professional Essay Writers for Hire
Essay Writing Service, EssayPro
Professional Custom
Professional Custom Essay Writing Services
In need of qualified essay help online or professional assistance with your research paper?
Browsing the web for a reliable custom writing service to give you a hand with college assignment?
Out of time and require quick and moreover effective support with your term paper or dissertation?