Disclaimer: This dissertation has been written by a student and is not an example of our professional work, which you can see examples of here.

Any opinions, findings, conclusions, or recommendations expressed in this dissertation are those of the authors and do not necessarily reflect the views of UKDiss.com.

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

Info: 34973 words (140 pages) Dissertation
Published: 10th Dec 2019

Reference this

Tagged: Biology

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.




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



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 Formulation Components………………………………………73


Page 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





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,


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


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


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


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,


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 Engineering, 14(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 Development, 10(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 Release, 128(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 Research, 103(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 Quality, 39(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 & Technology, 42(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 Chemistry, 26(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 Toxicology, 87(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 C, 133, 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 Polymers, 92(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 Management, 29(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 A, 25(3), 241–258. https://doi.org/10.1080/02652030701744538

Clement, J., & Jarrett, P. (1994). Antibacterial Silver. Metabolism Based Drugs, 1, 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 Research, 11(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 & Technology, 24(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. BioScience, 58(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 Science, 84(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 Nanomedicine, 3(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 Aspects, 372(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 Environment, 568, 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 Research, 52(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 & Technology, 49(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 Chemotherapy, 54(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 & Technology, 43(24), 9216–9222.

Gottschalk, F., Sun, T., & Nowack, B. (2013). Environmental concentrations of engineered nanomaterials: Review of modeling and analytical studies. Environmental Pollution, 181, 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. Ecotoxicology, 17(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 Res, 112, 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 Nano, 5(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. Small, 4(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 Nanotechnology, 6, 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 Pollution, 158(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 Technology, 43(16), 1823–1867. https://doi.org/10.1080/10643389.2012.671750

Kah, M., & Hofmann, T. (2014). Nanopesticide research: Current trends and future priorities. Environment International, 63, 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 Perspect, 119(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 Chemistry, 32, 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 Science, 86(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 & Technology, 46(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 Pharmacology, 38(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 Chemistry, 85(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 & Technology, 34(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 International, 86, 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 Microbiology, 62(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. Science, 311(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 Microencapsulation, 29(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. Phytopathology, 103(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.” HortScience, 48(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 Technology, 45(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 Research, 14(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 Science, 28(2), 290–302.

Rashidi, L., & Khosravi-Darani, K. (2011). The Applications of Nanotechnology in Food Industry. Critical Reviews in Food Science and Nutrition, 51(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. Langmuir, 32(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 Nanotechnology, 10(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 Pollution, 199, 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. Nanomaterials, 5(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 Research, 48, 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 Health, 16, 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 Nanotechnology, 6, 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. Nanoscale, 7(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 Chemistry, 17(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 Toxicology, 35(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 Agriculture, 11(7), 1151–1158.






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). 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. 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).



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 Nanotechnology, 6, 1769–1780.


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 Toxicology, 87(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 Management, 29(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 Research, 12(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 Pharmacology, 38(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 Research, 14(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 Quality, 39(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 & Technology, 42(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 Pollution, 181, 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 Pollution, 158(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 & Technology, 47(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. Ecotoxicology, 17(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 Chemistry, 35(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 Chemistry, 85(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 & Technology, 49(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. Nanotechnology, 24(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 Technology, 45(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. Science, 311(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 Science, 204, 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. Nanomaterials, 5(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. Nanotechnology, 27(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. Nanotoxicology, 9(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 Nano, 8(1), 374–386. https://doi.org/10.1021/nn4044047

27. Clement, J., & Jarrett, P. (1994). Antibacterial Silver. Metabolism Based Drugs, 1, 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 Chemotherapy, 54(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. Nanotechnology, 16(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 aureus. Journal of Biomedical Materials Research, 52(4), 662–668.

31. Fong, J., & Wood, F. (2006). Nanocrystalline silver dressings in wound management: a review. International Journal of Nanomedicine, 1(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. Small, 4(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 Science, 145(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 Nanotechnology, 10(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. Langmuir, 32(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. ChemPhysChem, 13(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 Sciences, 86(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 Toxicology, 29, 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 Polymers, 85, 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 Dynamics, 203(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 Chemistry, 26(10), 2241–2246.

45. Lahnsteiner, F. (2008). The effect of internal and external cryoprotectants on zebrafish               (Danio rerio) embryos. Theriogenology, 69(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 Toxicology, 105(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? Aquatic Toxicology, 87(2), 81–87. https://doi.org/10.1016/j.aquatox.2008.01.004

48. Bilberg, K., Hovgaard, M. B., Besenbacher, F., & Baatrup, E. (2012). In Vivo Toxicity of Silver Nanoparticles and Silver Ions in Zebrafish (Danio rerio). Journal of Toxicology, 2012, 1–9. https://doi.org/10.1155/2012/293784

49. Powers, C. M., Slotkin, T. A., Seidler, F. J., Badireddy, A. R., & Padilla, S. (2011). Silver nanoparticles alter zebrafish development and larval behavior: Distinct roles for particle size, coating and composition. Neurotoxicology and Teratology, 33(6), 708–714. https://doi.org/10.1016/j.ntt.2011.02.002

50. Powers, C. M., Yen, J., Linney, E. A., Seidler, F. J., & Slotkin, T. A. (2010). Silver exposure in developing zebrafish (Danio rerio): Persistent effects on larval behavior and survival. Neurotoxicology and Teratology, 32(3), 391–397. https://doi.org/10.1016/j.ntt.2010.01.009

51. Rombough, P. (1985). The influence of zona radiata on the toxicities of zinc, lead, mercury, copper and silver ions to embryos of steelhead trout Salmo gairdneri. Comparative Biochemistry and Physiology, 82C(1), 115–117.

52. 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 & Technology, 34(19), 4199–4207. https://doi.org/10.1021/es9912168

53. Lapresta-Fernández, A., Fernández, A., & Blasco, J. (2012). Nanoecotoxicity effects of engineered silver and gold nanoparticles in aquatic organisms. TrAC Trends in Analytical Chemistry, 32, 40–59. https://doi.org/10.1016/j.trac.2011.09.007

54. Reidy, B., Haase, A., Luch, A., Dawson, K., & Lynch, I. (2013). Mechanisms of Silver Nanoparticle Release, Transformation and Toxicity: A Critical Review of Current Knowledge and Recommendations for Future Studies and Applications. Materials, 6(6), 2295–2350. https://doi.org/10.3390/ma6062295

55. Bar-Ilan, O., Albrecht, R. M., Fako, V. E., & Furgeson, D. Y. (2009). Toxicity Assessments of Multisized Gold and Silver Nanoparticles in Zebrafish Embryos. Small, 5(16), 1897–1910. https://doi.org/10.1002/smll.200801716

56. Fischer, D., Li, Y., Ahlemeyer, B., Krieglstein, J., & Kissel, T. (2003). In vitro cytotoxicity testing of polycations: influence of polymer structure on cell viability and hemolysis. Biomaterials, 24(7), 1121–1131.

57. Wandrey, C., Hernandez-Barajas, J., & Hunkeler, D. (1999). Diallyldimethylammonium Chloride and its Polymers. Advances in Polymer Sciences, 145, 123–177.

58. Kircheis, R., Wightman, L., & Wagner, E. (2001). Design and gene delivery activity of modified polyethylenimines. Advanced Drug Delivery Reviews, 53, 341–358.

59. Hong, S., Leroueil, P. R., Janus, E. K., Peters, J. L., Kober, M.-M., Islam, M. T., … Banaszak Holl, M. M. (2006). Interaction of Polycationic Polymers with Supported Lipid Bilayers and Cells: Nanoscale Hole Formation and Enhanced Membrane Permeability. Bioconjugate Chemistry, 17(3), 728–734. https://doi.org/10.1021/bc060077y

60. Mecke, A., Majoros, I. J., Patri, A. K., Baker, J. R., Banaszak Holl, M. M., & Orr, B. G. (2005). Lipid Bilayer Disruption by Polycationic Polymers: The Roles of Size and Chemical Functional Group. Langmuir, 21(23), 10348–10354. https://doi.org/10.1021/la050629l

61. Robertson, G. N., McGee, C. A. S., Dumbarton, T. C., Croll, R. P., & Smith, F. M. (2007). Development of the swim bladder and its innervation in the zebrafish, Danio rerio. Journal of Morphology, 268(11), 967–985. https://doi.org/10.1002/jmor.10558

62. Asharani, P. V., Lian Wu, Y., Gong, Z., & Valiyaveettil, S. (2008). Toxicity of silver               nanoparticles in zebrafish models. Nanotechnology, 19(25), 255102.               https://doi.org/10.1088/09574484/19/25/255102

63. Asharani, P. V., Lian Wu, Y., Gong, Z., & Valiyaveettil, S. (2011). Comparison of the toxicity of silver, gold and platinum nanoparticles in developing zebrafish embryos. Nanotoxicology, 5(1), 43–54. https://doi.org/10.3109/17435390.2010.489207

64. Lee, K. J., Browning, L. M., Nallathamby, P. D., Osgood, C. J., & Xu, X.-H. N. (2013). Silver nanoparticles induce developmental stage-specific embryonic phenotypes in zebrafish. Nanoscale, 5(23), 11625. https://doi.org/10.1039/c3nr03210h

65. Osborne, O. J., Johnston, B. D., Moger, J., Balousha, M., Lead, J. R., Kudoh, T., & Tyler, C. R. (2013). Effects of particle size and coating on nanoscale Ag and TiO2 exposure in zebrafish (Danio rerio) embryos. Nanotoxicology, 7(8), 1315–1324. https://doi.org/10.3109/17435390.2012.737484

66. Liu, W., Long, Y., Yin, N., Zhao, X., Sun, C., Zhou, Q., & Jiang, G. (2016). Toxicity of engineered nanoparticles to fish. Engineered Nanoparticles and the Environment: Biophysicochemical Processes and Toxicity, 4. Retrieved from http://books.google.com/books?hl=en&lr=&id=SmHtDAAAQBAJ&oi=fnd&pg=PA347&dq=%22Reactive+Oxygen%22+%22scenario.+Many+NPs+are+not+immediately+water%22+%22example,+Ag+NPs+that+were+released+into+lakes%22+%22other+vertebrates+including+human+beings+(Dorea%22+%22with+the+uncertainties+due+to+NPs%E2%80%99+special%22+&ots=4oBPjEUcY1&sig=F9ydNMo2BnPyK0vCHglqdRqPKyc

67. Lee, K. J., Nallathamby, P. D., Browning, L. M., Osgood, C. J., & Xu, X.-H. N. (2007). In Vivo Imaging of Transport and Biocompatibility of Single Silver Nanoparticles in Early Development of Zebrafish Embryos. ACS Nano, 1(2), 133–143. https://doi.org/10.1021/nn700048y

68. Kiener, T. K., Selptsova-Friedrich, I., & Hunziker, W. (2008). Tjp3/zo-3 is critical for epidermal barrier function in zebrafish embryos. Developmental Biology, 316(1), 36–49. https://doi.org/10.1016/j.ydbio.2007.12.047

69.  Lin, S., Zhao, Y., Ji, Z., Ear, J., Chang, C. H., Zhang, H., … Nel, A. E. (2013). Zebrafish               High-Throughput Screening to Study the Impact of Dissolvable Metal Oxide               Nanoparticles on the Hatching Enzyme, ZHE1. Small, 9(9-10), 1776–1785.               https://doi.org/10.1002/smll.201202128


Over the past few decades, nanotechnology has significantly impacted various sectors by

improving technologies and expanding conventional applications.  To illustrate the importance

of nanotechnology, I have provided a thorough background which includes discussion on the

impact and origins of nanotechnology and engineered nanomaterials, the important inherent

properties of nanomaterials with a discussion on risk characterization, and discussed the various

applications of nanotechnology.  The opportunities provided by nanotechnology continue to

increase as the global market is expected to grow by 18.1% over the next decade to reach

$173.95 billion by 2025 (Accuray Research LLP, 2016).  With this growth, we must continue to

fill gaps in the literature to understand the risk associated with increased use of nanomaterials.

Some of the gaps in the literature can be addressed by providing complete

characterizations to understand particle behavior to allow for comparisons to other

nanomaterials.  We can also improve predictive models for estimating transformations of

nanomaterials in the environment, particularly by testing nanomaterials in natural waters.

Additionally, we should address the outdated AQWC and determine whether we need specific

regulations for nanomaterials, as there is no mention of particle-specific considerations in United

States law.  If these gaps are addressed, more complete risk assessments can be performed to

determine whether the use of a nanoparticle poses unreasonable risk to human or environmental


Although data gaps exist for many nanomaterials, silver nanoparticles are one of the most

well-studied and most utilized nanomaterials.  There are many mechanisms of action that are

proposed to elicit toxic responses in organisms, with the majority relating to released silver ions.

With silver ions consistently released from silver nanoparticles, it has been noted that silver can

enter water systems and have unintended impacts on aquatic organisms.  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.  By utilizing a lignin core, this is an application of green chemistry which is becoming

more prevalent as we learn more of nanomaterial fate and toxicity.  The goal of green chemistry

is to reduce harmful effects to the environment and non-target organisms while still maintaining

the desired antimicrobial activity (in the case of silver nanoparticles) during application.

In our study, we found that ionic silver and the free surface stabilizer were the most toxic constituents, although when associated separately or together with the lignin core, toxicity of the formulations decreased significantly—this indicated that the lignin core is an effective alternative to conventional silver nanoparticles for potentially reducing the burden of silver released into the environment.  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.

Cite This Work

To export a reference to this article please select a referencing stye below:

Reference Copied to Clipboard.
Reference Copied to Clipboard.
Reference Copied to Clipboard.
Reference Copied to Clipboard.
Reference Copied to Clipboard.
Reference Copied to Clipboard.
Reference Copied to Clipboard.

Related Services

View all

Related Content

All Tags

Content relating to: "Biology"

Biology is the scientific study of the natural processes of living organisms or life in all its forms. including origin, growth, reproduction, structure, and behaviour and encompasses numerous fields such as botany, zoology, mycology, and microbiology.

Related Articles

DMCA / Removal Request

If you are the original writer of this dissertation and no longer wish to have your work published on the UKDiss.com website then please: