POTENTIAL ENVIRONMENTAL IMPLICATIONS OF …ufdcimages.uflib.ufl.edu/UF/E0/02/21/29/00001/gao_j.pdf · 2013-05-31 · 1 NANOTECHNOLOGY AND THE ENVIRONMENT: APPLICATIONS AND IMPLICATIONS

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POTENTIAL ENVIRONMENTAL IMPLICATIONS OF MANUFACTURED NANOMATERIALS: TOXICITY, MOBILITY, AND NANOWASTES IN AQUATIC AND

SOIL SYSTEMS

By

JIE GAO

A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT

OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY

UNIVERSITY OF FLORIDA

2008

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© 2008 Jie Gao

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This dissertation is dedicated to my family, my parents and my husband, for their tremendous love and support

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ACKNOWLEDGMENTS

I would like to sincerely thank my advisor, Dr. Jean Claude J. Bonzongo, for his support,

guidance, encouragement, and most importantly, his friendship during my graduate study at the

University of Florida. I was really blessed to be under his mentorship over the past four years. I

am also very grateful to all my committee members, Dr. Gabriel Bitton, Dr. Joseph J. Delfino,

Dr. Dmitry Kopelevich, Dr. Lena Q. Ma, and Dr. Rongling Wu, for their generous help and

support during my journey as a graduate student at the University of Florida.

I would like to thank Dr. Kirk J. Ziegler and Dr. Bin Gao for their valuable assistance and

guidance on specific aspects of my research. Thanks are also extended to Peter Meyers and Craig

Watts from Hydrosphere Research for their time and patience during my training in toxicity

testing and for providing the pure cultures of Ceriodaphnia dubia and Pseudokirchneriella

subcapitata used in this research. Many thanks to Mr. Gill Brubaker and Mr. Gary Scheiffele

from Particle Engineering Research Center for letting me use the Ion Chromatography. Thanks

are also extende to Drs. Joseph Griffitt and David Barber from the Center for Environmental and

Human Toxicology for collaboration and help. Finally, I extend my gratitude to Dr. Nan Feng

and Mr. Randy Wang for their assistance with laboratory experiments, and to my fellow students

in our research group and in the Department of Environmental Engineering Sciences for their

camaraderie, kindness and support.

I would like to express very special thanks to my parents for their tremendous and

inconditional love and support. Only with their unselfish love and encouragement could I gain

self-confidence and ability to take on challenges and overcome difficulties in my life. Finally,

my thanks go to my husband, Yu Wang, for his support, patience, and unwavering love

throughout the past seven years of my college undergraduate and graduate studies.

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TABLE OF CONTENTS page

ACKNOWLEDGMENTS ...............................................................................................................4

LIST OF TABLES...........................................................................................................................8

LIST OF FIGURES .......................................................................................................................10

ABSTRACT...................................................................................................................................12

CHAPTER

1 NANOTECHNOLOGY AND THE ENVIRONMENT: APPLICATIONS AND IMPLICATIONS ....................................................................................................................15

1.1 Problem Statement............................................................................................................15 1.2 Production and Use of Nanomaterials ..............................................................................18 1.3 Potential Toxicity of Manufactured Nanomaterials .........................................................19 1.4 Potential Effects of MNs on Ecosystem Functions ..........................................................20 1.5 Environmental Fate and Transport of MNs ......................................................................22 1.6 Research Objectives..........................................................................................................23

2 POTENTIAL TOXICITY OF CARBON AND METAL BASED NANOMATERIALS.....28

2.1 Introduction.......................................................................................................................28 2.2 Materials and Methods .....................................................................................................31

2.2.1 Chemicals ...............................................................................................................31 2.2.2 Preparation of Nanomaterial Suspensions..............................................................31 2.2.3 96-hour Algal Chronic Toxicity Assay Using Pseudokirchneriella

subcapitata (Selenatastrum capricornutum)................................................................33 2.2.4 48-hour Acute Toxicity Assay Using Ceriodaphnia dubia as Test Model

Organism......................................................................................................................34 2.2.5 MetPLATE Test .....................................................................................................35

2.3 Results and Discussion .....................................................................................................36 2.3.1 Characterization of Nano-metal Particles...............................................................36 2.3.2 Toxicity of Solvents and Surfactants......................................................................37 2.3.3 Toxicity of Tested Nanomaterials ..........................................................................38

2.3.3.1. Fullerene (C60).............................................................................................38 2.3.3.2. Single-Walled Nanotubes (SWNTs) ...........................................................39 2.3.3.3. Metallic nanomaterials ................................................................................40

2.4 Conclusions.......................................................................................................................43

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3 TOXICITY OF SELECTED MANUFACTURED NANOMATERIALS DISPERSED IN NATURAL WATERS WITH GRADIENTS IN IONIC STRENGTH AND DISSOLVED ORGANIC MATTER CONTENT..................................................................55


3.2.1 Collection of Water Samples..................................................................................56 3.2.2 Preparation of Nanomaterial Suspensions in Collected Water Samples ................57 3.2.3 Determination of MNs Concentrations in Prepared Suspensions ..........................57 3.2.4 Toxicity of MNs Suspended in Natural Waters .....................................................58

3.2.4.1 The Ceriodaphnia dubia assay.....................................................................58 3.2.4.2 MetPLATE test ............................................................................................59

3.3 Results and Discussion .....................................................................................................60 3.3.1 Characterization of Water Samples ........................................................................60 3.3.2 Total Concentration of Dispersed Nanomaterials ..................................................60 3.3.3 Evaluation of Acute Toxicity of Nanomaterials Suspended in Natural Waters

to Ceriodaphnia dubia .................................................................................................63 3.3.4 Acute Toxicity with MetPLATE............................................................................65

3.4 Conclusions.......................................................................................................................66

4 MOBILITY OF SINGLE-WALLED CARBON NANOTUBES (SWNTS) IN SATURATED HETEROGENEOUS POROUS MEDIA ......................................................74


4.2.1 Single-Walled Carbon Nanotube Sample Preparation ...........................................76 4.2.2 Soil Sample Collection and Characterization.........................................................77 4.2.3 Column Experiments ..............................................................................................77 4.2.4 Modeling.................................................................................................................78

4.3 Results and Discussion .....................................................................................................79 4.3.1 Bromide Transport and Breakthrough Curves in Sandy and Clay soils.................79 4.3.2 SWNTs Transport in Sandy Soils...........................................................................79 4.3.3 SWNTs Transport in Clay Soils .............................................................................81

4.4 Conclusions.......................................................................................................................82

5 POTENTIAL IMPACTS OF MANUFACTURED NANOMATERIALS ON BIOGEOCHEMICAL PROCESSES IN SEDIMENTS.........................................................89


5.2.1 Preparation of Nanomaterial Suspensions..............................................................91 5.2.2 Sediment Collection ...............................................................................................91 5.2.3 Dominant Terminal Electron Accepting Processes (TEAPs) in Sediments and

Sediment Manipulation in this Study...........................................................................91 5.2.4 Analytical Techniques ............................................................................................95 5.2.5 Data Analysis..........................................................................................................95

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5.3 Results and Discussion .....................................................................................................95 5.4 Conclusions.......................................................................................................................98

6 NANOWASTES IN THE ENVIRONMENT: THE “TROJAN HORSE EFFECT” OF NANOMATERIALS............................................................................................................103

6.1 Introduction.....................................................................................................................103 6.2 Materials and Methods ...................................................................................................104 6.3 Results and Discussion ...................................................................................................105 6.4 Conclusions.....................................................................................................................107

7 CONCLUSIONS AND RECOMMENDATIONS...............................................................111

7.1 Conclusions.....................................................................................................................111 7.2 Recommendations...........................................................................................................112

APPENDIX

A TESTED CONCENTRATIONS OF CARBON- AND METAL-BASED NANOMATERIALS IN THREE DIFFERENT TOXICITY ASSAYS..............................114

B TESTED CONCENTRATIONS OF MANUFACTURED NANOMATERIAL SUSPENSIONS IN TOXICITY TESTS USING THE C. DAPHNIA 48-H ACUTE TOXICITY ASSAY AND METPLATE TEST ...................................................................115

LIST OF REFERENCES.............................................................................................................116

BIOGRAPHICAL SKETCH .......................................................................................................135

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LIST OF TABLES

Table page 1-1 Examples of materials and applications of nanotechnology (Karn 2004) .........................26

1-2 Examples of proposed environmental applications of manufactured nanomaterials.........27

2-1 Tested surfactants and solvents and their chemical compositions.....................................45

2-2 Characteristics of metallic nanoparticles used in toxicity experiments in this study ........46

2-3 Chemical composition of the preliminary algal assay procedure (PAAP) culture medium ..............................................................................................................................47

2-4 Concentrations of tested surfactants resulting in lethal effect on 50% of the population (LC50) based on the 48-h Ceriodaphnia dubia assay.......................................48

2-5 Surfactant concentrations resulting in the inhibition of 50% of growth (IC50) in a 96-h P. subcapitata chronic toxicity assay..............................................................................49

2-6 Percent mortality of C. dubia exposed to solutions with increasing THF concentrations in 48-h accute toxicity assay......................................................................50

2-7 Concentrations of tested metal- and carbon- based nanoparticles resulting in lethal effect on 50% of the population (LC50) based on the 48-h Ceriodaphnia dubia assay. ....51

2-8 Examples of published EC50 values for fullerenes (C60), single-walled carbon nanotubes (SWNTs), and nano-copper (nano-Cu) on daphnia and zebrafish ...................52

2-9 Concentrations of tested metal- and carbon- based nanoparticles resulting in growth inhibition of 50% of the population (IC50) based on the 96-h P. subcapitata chronic toxicity assay......................................................................................................................53

2-10 Concentrations of tested metal- and carbon- based nanoparticles resulting in 50% inhibition of color development in MetPLATE test. .........................................................54

3-1 Characteristics of water samples prior to contact with C60, Ag and Cu nanoparticles ......73

4-1 Physicochemical characteristics of the sandy (Gainesville, Florida) and clayey (Atlanta, Georgia) soils used in column experiments (Feng et al. 2007). .........................87

4-2 Transport parameters estimated by CXTFIT for bromide and SWNT in GA and SDS in sandy soils......................................................................................................................88

A-1 Tested concentrations of carbon- and metal-based nanomaterials in three different toxicity assays ..................................................................................................................114

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B-1 Tested concentrations of manufactured nanomaterial suspensions in toxicity tests using the C. daphnia 48-h acute toxicity assay and MetPLATE test ..............................115

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LIST OF FIGURES

Figure page 1-1 Conceptual diagram of the life cycle and potential pathways and fate of

manufactured nanomaterials ..............................................................................................25

2-1 Size distribution of selected metallic MN samples as obtained from commercial sources................................................................................................................................44

3-1 Map of the Suwannee River watershed and tributaries showing the three sampling locations .............................................................................................................................67

3-2 Concentrations of silver, copper, and C60 in different water samples spiked with individual nanomaterial and then filtered ..........................................................................68

3-3 Linear correlation of Cu concentrations and dissolved organic matter (DOC) in water samples SR1, SR2 and DI water (R2=0.9864). ..................................................................69

3-4 48-h LC50 values of silver- and copper-spiked water samples to Ceriodaphnia dubia .....70

3-5 Relationship between the 48-h LC50 values of nanocopper suspensions to Ceriodaphnia dubia and dissolved organic matter (DOC) concentrations in SR1, SR2, and DI water..............................................................................................................71

3-6 IC50 values of nanosilver- and nanocopper suspensions using MetPLATE test................72

4-1 Schematic diagram of the experimental setup for SWNTs transport in packed heterogeneous sandy or clay soils......................................................................................84

4-2 Breakthrough curves of experimental and simulated data of bromide (Br-) in sandy and clay soils......................................................................................................................85

4-3 Breakthrough curves of experimental and simulated data of SWNT-GA and SWNT-SDS suspensions in sandy soil columns ............................................................................86

5-1 Kinetics of acetate degradation, and nitrate, nitrite and sulfate concentrations in sediment slurries without (controls) or spiked with tested nanomaterials (C60, nanosilver, and CdSe quantum dots) .................................................................................99

5-2 Kinetics of acetate degradation, and nitrate, nitrite and sulfate concentrations in sediment slurries spiked with excess nitrate and without (controls) or with tested nanomaterial additions (C60, nanosilver, and CdSe quantum dots) .................................100

5-3 Kinetics of acetate degradation and nitrate, nitrite and sulfate concentrations in sediment slurries spiked with excess sulfate and without (controls) or with tested nanomaterial additions (C60, nanosilver, and CdSe quantum dots) .................................101

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5-4 Pseudo-first order kinetics of acetate disappearance from sediment-slurries treated with either silver nanoparticles or CdSe quantum dots as compared to the non-treated controls.............................................................................................................................102

6-1 Percent THg converted to methyl-Hg in sediment slurries spiked with SiO2-TiO2-Hg complexes and incubated at different pH.........................................................................108

6-2 Kinetics of Hg methylation in sediment slurries spiked with SiO2-TiO2-Hg complexes at pH 4 (triangles; native pH), 5 (circles), and 6 (squares)............................109

6-3 Toxicity effect of Synthetic Precipitation Leaching Procedure (SPLP) solutions obtained from leaching of virgin and Hg-loaded SiO2-TiO2 nanocomposites in a 1:60 ratio (ml SPLP/mg nanomaterials)...................................................................................110

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Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy

POTENTIAL ENVIRONMENTAL IMPLICATIONS OF MANUFACTURED

NANOMATERIALS: TOXICITY, MOBILITY, AND NANOWASTES IN AQUATIC AND SOIL SYSTEMS

By

Jie Gao

August 2008

Chair: Jean-Claude Bonzongo Major: Environmental Engineering Sciences

Nanotechnology has been singled out by industry and governments to become the world’s

largest industrial revolution, and it carries the potential to substantially benefit environmental

quality through pollution prevention, treatment, and remediation. However, nanotechnology

could also lead to serious environmental problems since the environmental behavior and fate of

manufactured nanomaterials (MNs) are not predictable from that of chemically similar but larger

compounds. The goal of this study was to develop an understanding of the potentially complex

interplay between MNs and the health of organisms and ecosystems. The potential effects of

MNs were evaluated by testing the hypothesis that: “chemical elements used in the production of

MNs could lead to environmental dysfunctions due to: (1) the potential toxicity of these elements

and their derivatives, (2) the small size driven mobility of MNs through heterogeneous porous

media and ultimate contamination of aquifers, (3) their toxicity to microorganisms and the

resulting negative impacts on key environmental microbial catalyzed reactions, and (4) the large

surface area which would allow MNs to act as carriers/delivers of pollutants adsorbed onto

them”.

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To address this broad hypothesis, three well-established small-scale toxicity tests (i.e. the

Ceriodaphnia dubia acute toxicity test, the Pseudokirchneriella subcapitata chronic toxicity test,

and MetPLATE™), were used. In addition, studies at the system level were conducted using a

combination of column and batch experiments to investigate the transport behavior of MNs in

heterogeneous porous media and the interactions of MNs with microbial-catalyzed oxidation of

organic matter in sediments.

Carbon (i.e. fullerenes (C60), single-walled carbon nanotubes (SWNTs)) and metal (i.e.

CdSe quantum dots, and powders of the following nanometals—Ag, Cu, Co, Ni, and Al) based

MNs, were used in different laboratory experiments. All tested MNs showed some degree of

toxicity response to either one or more of the above three microbiotests, with nano-Cu and nano-

Ag being the most toxic. The use of experimental conditions that mimic likely scenarios of MNs

introduction to aquatic systems showed that the toxicity response of test model organisms to

MNs under such conditions would be affected by key water quality parameters such as organic

matter content and solution chemistry. Column studies of SWNTs transport in heterogeneous

porous soils showed that soils characteristics and the chemical composition of MN suspensions

affect transport behaviors, and that the latter can be quantitatively predicted by use of

mathematical models such as the convection-dispersion equation. Finally, the use of sediment

slurries spiked with either each type of MNs or pollutant (i.e. mercury) bound to MNs allowed

the assessment of: (1) the impact of MNs on microbially-catalyzed oxidation of organic matter,

and (2) the potential for Hg-bound to SiO2-TiO2 nanocomposites obtained from flue gas

remediation studies to become available in sedimentary environments as a function of pH.

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Overall, these findings help shed light in the poorly studied environmental implications of

MNs. However, several questions remain unanswered as these short-term laboratory

investigations may not be able to predict the fate and transport of MNs on a long-term basis.

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CHAPTER 1 NANOTECHNOLOGY AND THE ENVIRONMENT: APPLICATIONS AND

IMPLICATIONS

1.1 Problem Statement

Nanoscience and nanotechnologies are currently generating an extraordinary interest as

they carry expectations that are believed to bring about changes as profound as the industrial

revolution, antibiotics, and nuclear weapons all in one (Ajayan et al. 1999). Manufactured

nanoparticles (MNs) find use in a wide variety of human activities, from medical, electronics, to

environmental research and applications. But, while the advantages of nanoscience and

nanotechnologies are multiple with many more still to be discovered, the potential implications

of this “new” technology on the environment and living organisms remain largely unknown

(Fischer and Chan 2007; Oberdorster 2004; Seetharam and Sridhar 2007; Warheit 2008). In fact,

in current high-throughput societies, one would anticipate the peak in MNs production and use to

be followed by either intentional (landfills) and/or non-intentional (diffuse) introduction of these

materials into different environmental compartments. The conceptual diagram presented in

Figure 1-1 summarizes the potential pathways and fate of MNs from “cradle to grave”. From this

simplified diagram, it is obvious that MNs could enter the environment and come in contact with

living organisms from different stages of their life cycle. Accordingly, a significant effort is

needed for “upstream determination” of the potential impacts of this emerging technology on the

environment and human health.

The literature is now quite abundant with papers dealing with both the preparation and use

of MNs in several industrial, environmental, and medical applications (Borderieux et al. 2004;

Davis 1997; Eng 2004; Florence et al. 1995; Jensen et al. 1996; Li et al. 2006; Pitoniak et al.

2003; Rutherglen and Burke 2007; Tungittiplakorn et al. 2004; Wang and Zhang 1997; Wang et

al. 2008; Yogeswaran and Chen 2008; Zajtchuk 1999). Unfortunately, and despite the current

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ongoing discussion on MNs, the study on the environmental fate and impacts of nanoparticles

remains a frontier science. Although growing rather quickly, the number of published papers

with experimental data on this subject is still very limited.

At the end of the 20th and the beginning of the 21st century, the UK Government

commissioned the Royal Society and the Royal Academy of Engineering to carry out an

independent study into current and future developments in nanoscience and nanotechnologies

and their potential negative implications. Their findings were published in a report entitled

“Nanoscience and nanotechnologies: opportunities and uncertainties”, released in July 2004

(www.royalsoc.ac.uk/policy). Amongst several key points highlighted by the study, the report

noted the lack of published data on negative impacts of MNs and recommended research into the

hazards and exposure pathways of nanoparticles to reduce the many uncertainties related to their

potential impacts on health, safety, and the environment. In parallel with their effort, and mostly

after their final report, experimental data along the lines of the above-mentioned

recommendations have been increasing in the literature. For instance, research on toxicity of

MNs on living organisms is increasing very rapidly, starting from some initial laboratory studies

exposing fish to carbonaceous MNs (Oberdorster 2004) to the widespread use of several

mammal models (Borm et al. 2004; Davoren et al. 2007; Koyama et al. 2006; Oberdorster 2000;

Wick et al. 2007), and several aquatic test model organisms (Cheng et al. 2007; Roberts et al.

2007; Smith et al. 2007; Templeton et al. 2006; Zhu et al. 2006). With regard to living

organisms, a laboratory study exposing fish to fullerenes (C60) showed that C60 could cause brain

damage, resulting in a 17-fold increase in fish brain damage as compared to non-treated controls

when exposed to 0.5 ppm of fullerene aqueous suspension (Oberdorster 2004). The observed

damage, the lipid peroxidation is known to impair the normal functioning of cell membranes and

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has been linked to illnesses such as Alzheimer’s disease in humans. Besides the carbon based

MNs, metal nanoparticles have also shown a tendency for both bioaccumulation and toxicity.

Using optical and transmission electron microscopy, Xu et al. (2004) showed that non-

functionalized Ag nanoparticles as large as 80 nm in diameter could cross the cell membrane of

Pseudomonas aeruginosa. However, the tested Ag-nanoparticles did not damage the cell

membrane. Similar tests on E. coli and red blood cells using functionalized gold (Au)

nanoparticles showed moderate toxicity caused by lysis of cell membranes mediated by

nanoparciles cross-linked with cationic side chains (Goodman et al. 2004). Overall, these studies

showed that NMs could accumulate in living cells. Therefore, the potential exist for MNs to be

transferred through different trophic levels. Unfortunately, the extent and potential effects of

such transfer, if any, remain unknown.

In addition to toxicity studies, and although still very limited, both the fate and transport of

MNs in natural systems have been investigated (Espinasse et al. 2007; Lecoanet et al. 2004;

Lecoanet and Wiesner 2004). Research by Lecoanet and Wiesner (2004) and Lecoanet et al.

(2004) have shown that MNs could exhibit different transport behaviors in porous media. These

findings are important with regard to the assessment of both the efficacy (e.g., when used for

remediation purposes) and environmental impacts of MNs.

The field of nanotechnology is so new that there has been little time to develop substantive

data on exposure and hazards. To date, a few studies have investigated the toxicological effects

of both direct and indirect exposure to nanomaterials, but unfortunately, no clear guidelines for

these effects exist (Colvin 2003). On the other hand, the environmental toxicity of nanomaterials

is not predictable from that of similar but common larger compounds. Finally, despite the fact

that there has been a recent increasing effort to understand ecosystems at the system level, both

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theoretical and experimental investigations of the impacts of MNs on ecosystem functions are

still lacking.

1.2 Production and Use of Nanomaterials

Nanotechnology has been pegged by industry and governments to become the world’s

largest and fastest industrial revolution (Roco et al. 1999), due to its vast potential for use in

different fields such as medicine, electronics, chemistry, and engineering (Kung and Kung 2004;

Navalakhe and Nandedkar 2007).

According to the U.S. National Institute for Occupational Safety and Health (NIOSH),

global investment in nanotechnology by government alone rose from $432 million in 1997 to

about $3 billion in 2003 and the predicted value of nanotech-related products will exceed $1

trillion worldwide by 2015 (Toensmeier 2004). MNs, defined as particles with 100 nm or less in

diameter, are used in semiconductor manufacture and biomedical applications, as well as

consumer products ranging from anti-aging cream to sunblocks (ETCGroupGenotype 2004;

Hund-Rinke and Simon 2006; Wiesner 2003). The production and synthesis of MNs have been

accomplished by a wide variety of techniques including sol-gel technique and arc processes, to

name a few (Farhat and Scott 2006; Niederberger et al. 2006). Several products containing

nanomaterials (e.g., Smith and Nephew antimicrobial dressings, Babolat Nanotube Power tennis

rackets, and NuCelle SunSense sunscreen) are already found on the market and many more are

anticipated, particularly with the application of nanotechnology to electronics. At least 44

elements are commercially available in nanoscale form and another 20 elements are expected to

be on the market in the near future (ETCGroup 2003). The most common nanomaterials and

their applications are listed in Table 1-1 (Karn 2004). With regard to environmental applications,

nanotechnology may offer novel materials and processes for pollution prevention and treatment

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(Zhang and Karn 2005). Table 1-2 summarizes the environmental applications of some

nanomaterials.

1.3 Potential Toxicity of Manufactured Nanomaterials

Aside from interests in the potential applications of nanotechnology, there are also

concerns that nanoparticles may be more toxic than their bulk particles because of larger surface

area, enhanced chemical reactivity and potential for cell penetration (Monteiro-Riviere and

Orsière 2007). Studies have demonstrated that even the considerably inert TiO2 can in nano size

exert a “strong oxidizing power that attacks organic molecules” or produce highly reactive free

radicals (Adams et al. 2006; Borm 2002; ETCGroup 2003). While used in cosmetics, the

photocatalytic activity of TiO2 nanoparticles can lead to the degradation of organic additives and

to the generation of active species and further induce the transformation of biological molecules

on the skin, which initiate harmful reactions or even photo-induced mutagenicity for the skin

(Picatonotto et al. 2001).

Nanotechnology has promises in many biological and pharmaceutical applications, but

knowledge of their toxicological effects on living organisms is still very limited. Due to their

small size and the ability to escape macrophages, nanoparticles can penetrate the human body via

various routes and persist in the system (Karakoti et al. 2006). They can enter human tissues via

the lungs after inhalation, through the digestive system, and through the skin by dermal contact.

Once in the body, they are able to penetrate even very small capillaries and distribute throughout

the system (Braydich-Stolle et al. 2005). A number of studies have investigated the toxicological

effects of nanomaterials both in vivo and in vitro (Jeng and Swanson 2006; Kirchner et al. 2005;

Limbach et al. 2007; Soto et al. 2007; Usenko et al. 2007). It is observed that carbon nanotubes

and fullerenes could induce the formation of reactive oxygen species (ROS) and associated

oxidative stress, and therefore cytotoxicity (Oberdorster 2004; Pulskamp et al. 2007; Shvedova et

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al. 2005). Nano-copper particles can distribute throughout the body into blood, brain, lung, heart,

kidney, spleen, liver, intestine and stomach, and provoke dysfunctions of the organs (Chen et al.

2006). Nano-silver particles, quantum dots and some other metal oxide nanoparticles are able to

produce ROS as well (Adams et al. 2006; Cho et al. 2007; Jain and Pradeep 2005; Kim et al.

2007; Stoimenov et al. 2002; Wang et al. 2007).

Although still in its infancy, the emerging field of nanotoxicology is fast growing. For

instance, standardized test methods are lacking in order to assess MNs safety, interpret extra-

laboratory data, and generate an online databank that is accessible to all users and manufacturers

(Kovochich et al. 2007). Therefore, knowledge of exposure to and hazard of nanomaterials are in

great need to fully understand their risks and impacts.

1.4 Potential Effects of MNs on Ecosystem Functions

Ecosystems around the world accomplish numerous “natural services” and most if not all

of them seem to have common main characteristics, including the flow of energy, the flow of

material, flow of information, and participation of biota and water. Therefore, the ability to

qualitatively and/or quantitatively characterize any of the above listed natural services can be

used to assess the impact of pollutants on ecosystem functions. Such abilities are provided by

thermodynamics, which has been successfully applied in describing the basic properties of

ecosystems (e.g. flow of matter and energy). By combining the flow of material and energy to

the participation of biota (microorganisms in this case), a series of reactions involved in

sedimentary cycling of organic carbon can be used as proxy to detect the potential impacts of

MNs on specific basic ecosystem functions. For instance, in pristine soils and sediments, the

composition and distribution of different microbial populations are usually well-established,

although changes associated with shift in seasons and other major parameters are also common.

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However, the introduction of a pollutant to such natural systems can result in a significant impact

on the composition of microbial communities and/or their activities. In such cases, the potential

consequences could range from a simple delay in the biodegradation of organic matter to major

environmental impacts such as the production of more toxic derivatives, with bioaccumulation

potential and negative effects on ecosystem functions.

In addition, MNs, which could occur in the environment in initially non-toxic levels would

likely bioaccumulate as an acute toxic response may not be expressed. Although the ability of

MNs to behave like some lipophylic pollutants such as polychlorinated biphenyls (PCBs) and

mercury remains largely unknown, it is probable that most carbonaceous MNs tend to behave

like the two chemicals. If so, they would then result in severe ecological consequences due to the

combination of bioaccumulation/biomagnification phenomena and the rather well-established

toxicity of some of these nanoparticles. Therefore, as MNs become widely used, aquatic and

terrestrial ecosystems would tend to behave as terminal sinks (Lyon et al. 2007). The reported

acute toxicity of several MNs based on laboratory experiments that use high MN-concentrations

in comparison to what might be expected in natural systems, tend to suggest that at trace

concentrations (ppt to ppb range), MNs could actually accumulate in living organisms without an

immediate toxicity response. Accordingly, such bioaccumulated MNs could make their way up

the food chain as low trophic organisms are consumed by those higher in the energy pyramid.

For instance, MNs accumulation by daphnids (Zhu et al. 2006) and earthworms (Brumfiel 2003)

could constitute a point of entry to food chains.

Unfortunately, this aspect of the potential implications of MNs has been overlooked,

despite the fact that the above observations should raise concerns about MN-impacts at a system

level.

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1.5 Environmental Fate and Transport of MNs

Nanomaterials exhibit novel and significantly different physical, chemical, and biological

properties than their larger counterparts, and this is due primarily to their small size and unique

structure (Masciangioli and Zhang 2003). If MN applications develop as projected, nanoparticle

introduction to ground waters and soils could present significant challenges with regard to

exposure risk for both aquatic/soil organisms and human health (Colvin 2003). The aquatic

environment may become contaminated through discharges of domestic wastewater effluents

(e.g., MNs leaching from washing machines, clothes and cosmetics), and through accidental

spillage by both manufacturing and transportation industries. In addition, their use in newly

developed environmental remediation techniques (e.g. groundwater and soil remediation) would

lead to intentional introduction of MNs to natural systems. Based on current knowledge, the

small size of MNs increases the potential for their dispersion (e.g. enhanced mobility) and

exposure (e.g. crossing of cell membranes); while their large specific surface area increases

chemical reactivity—and could facilitate adsorption and transport of other toxic pollutants in the

environment. Also, when released to the environment, MNs might undergo transformations due

to aggregation, sorption/desorption, deposition and bio-uptake—and identifying factors

controlling these processes is the key to predicting their environmental fate and impact (Arnall

2003). This is because changes in MN surface chemistry in combination with shifts in key

environmental parameters such as levels and type of organic matter, pH, and ionic strength could

affect their mobility, and therefore, their environmental fate and transport (Espinasse et al. 2007;

Kanel et al. 2007; Schrick et al. 2004).

A recent study by Lecoanet et al. (2004) showed that the mobility of different MNs in

porous media could differ substantially. For instance, fullerol (C60 hydroxide, C60(OH)m, m = 22-

26) and single wall carbon nanotubes (SWNTs) show a high mobility than silica, anatase,

23

ferroxane, alumoxane, and clusters of C60 under similar experimental conditions. These

observations point to the need to investigate the transport of MNs in porous media on a case-by-

case basis. Such studies should also take into account the physicochemical characteristics of the

matrix used as well as solution chemistry.

1.6 Research Objectives

The overall objective of this study is to assess the potential impacts of MNs on biota and

ecosystem functions. To meet this objective, a research approach was developed as screening

tools for toxicity effect identification. This first step was then followed by further laboratory

studies using selected toxic MNs to investigate (i) their toxicity under conditions representative

of natural systems, (ii) their mobility in porous media, (iii) their potential impact on basic

ecological functions (i.e., organic matter oxidation in sediment), and (iv) the fate of pollutants

adsorbed onto MNs, once released to the environment. Therefore, following this introductory

chapter on background information on nanotechnology and its implications, this dissertation is

structured as follows:

• Chapter 2 focuses on toxicological effects of several carbon and metal based nanomaterials determined by use of three small-scale bioassays: (1) an aquatic invertebrate based test—the 48-h Ceriodaphnia dubia short-term assay; (2) a freshwater algal-based test—the 96-h Pseudokirchneriella subcapitata or S. capricornutum chronic assay; and (3) a bacterial produced enzyme based test—the MetPLATE test. These tests were selected to cover different biological responses and adapted from previously published procedures for use in MNs toxicity tests

• Chapter 3 emphasizes the transformations and toxicity of MNs in natural waters. In this chapter a few selected MNs identified as toxic in Chapter 2 were tested for toxicity after being suspended into natural waters of different chemical compositions. The effects of organic matter content and ionic strength are investigated.

• Chapter 4 deals with the mobility aspects of single-walled carbon nanotubes (SWNTs) in heterogeneous porous media and the use of a modeling tool to predict the transport behavior of SWNTs versus soil characteristics and solution composition.

• Chapter 5 evaluates the potential effects of selected toxic MNs on biogeochemical processes in sediment. The implication is that any toxicity on microorganisms responsible

24

for degradation of organic matter would result in ecological implications such as reduced rates of organic matter oxidation.

• Chapter 6 focuses on the environmental fate of pollutants adsorbed onto MNs. In this study, mercury (Hg) is used as an example pollutant bound to SiO2-TiO2 nanocomposites, and Hg methylation is used as proxy for bioavailability of adsorbed Hg to sediment microorganisms.

• Chapter 7 provides a general conclusion and emphasizes the major findings of this study, as well as further research avenues.

25

Figure 1-1. Conceptual diagram of the life cycle and potential pathways and fate of manufactured nanomaterials (adapted from Helland et al. 2007).

Synthesis methods

NANO -PRODUCTS

AIR

SOIL

WATER

End of Life

Property Change- post -synthesis- intermediate

production

Environmental Health End Points

Living Organisms

Environment:Property Change

Exposure

Release

Release

Release

Release

End of Life

- post -synthesis- intermediate

production



Exposure

Release

Release

Release

Release

Nano-Products

Synthesis methods

NANO -PRODUCTS

AIR

SOIL

WATER

End of Life

Property Change- post -synthesis- intermediate

production


Living Organisms


Exposure

Release

Release

Release

Release

End of Life

- post -synthesis- intermediate

production



Exposure

Release

Release

Release

Release

Nano-Products

26

Table 1-1. Examples of materials and applications of nanotechnology (Karn 2004)

Nanostructures Size Example Material or Application

Clusters,nanocrystals, quantum dots Radius: 1-10 nm Insulators, semiconductors,

metals, magnetic materials

Nanowires Diameter: 1-100 nm Metals, semiconductors, oxides, sulfides, nitrides

Nanotubes Diameter: 1-100 nm Carbon, including fullerenes, layered chalcogenides

Other nanoparticles Radius: 1-100 nm Ceramic oxides, Buckyballs

27

Table 1-2. Examples of proposed environmental applications of manufactured nanomaterials Nanomaterials Applications Sources

Cadmium and mercury removal and reduction

Skubal et al. 2002

Reduction of Cr (VI) in aqueous solution Jiang et al. 2006

Nanosized TiO2

Efficient antimicrobial agent Mitoraj et al. 2007 Poly(ethylene) glycol modified urethane acrylate (PMUA) nanoparticles

An effective means to enhance the in-situ biodegradation rate in remediation through natural attenuation of contaminants

Tungittiplakorn et al. 2005

Nanoscale iron particles

Very effective for the transformation and detoxification of a wide variety of common environmental contaminants, such as chlorinated organic solvents, organochlorine pesticides, PCBs, and heavy metals

Zhang 2003 Nurmi et al. 2005 Kanel, et al. 2006 Liu and Zhao 2007 Kanel et al. 2006

An efficient sorbent for removal and determination of organic compounds

Long and Yang 2001 Cai et al. 2005 Jin et al. 2007 Lu and Liu 2006

Carbon nanotubes

Can be employed as biosensors for the detection of a number of biomolecules

Balasubramanian and Burghard 2006

Silica-titania nanocomposites

Shows high mercury vapor removal efficiency

Pitoniak et al. 2005

28

CHAPTER 2 POTENTIAL TOXICITY OF CARBON AND METAL BASED NANOMATERIALS

2.1 Introduction

Nanomaterials and nanotechnologies have become the world’s largest and fastest industrial

revolution, with an anticipated capacity to affect many industrial activities and lead to discovery

and implementation of unique materials and devices from electronics to engineered tissues (Roco

et al. 1999). Despite these anticipated applications, the products of nanoscience and

nanotechnologies also raise concerns about their potential health and environmental impacts

(Isobe et al. 2006). Thus, while much research effort is currently directed toward exploring the

properties and applications of nanomaterials in medicinal, industrial, agricultural, and

environmental fields, the body of experimental work on the potential implications of

nanomaterials on living organisms and ecosystem functions is now fast growing in response to

the introduction of nanomaterial-based products to the market place (Biswas and Wu 2005;

Griffitt et al. 2007; Nyberg et al. 2008; Oberdorster et al. 2006; Oberdörster et al. 2005).

However, this research remains in its infancy, as new nanosized materials are being produced

and incorporated into commercial products.

In both application and implication studies, homogenous and well-dispersed suspensions of

nanoparticles are desired for best performance. For example, when individually dispersed the

single-walled carbon nanotubes (SWNTs) have potential applications in biological sensing and

drug delivery systems (Zhang et al. 2005). However, current preparation methods lead to the

production of SWNTs aggregates, and without further treatment, the degree of aggregation

becomes a limiting factor in using these nanotubes (Zhang et al. 2005). On the other hand, since

the toxicity of such nanoparticles is size-dependent, their uncontrollable aggregation behavior

would likely make the interpretation of experimental results rather difficult. Therefore, the

29

determination of different physical characteristics of prepared suspensions is necessary prior, and

if possible, during toxicity studies (Oberdörster et al. 2005).

Fullerenes (C60 or buckyballs) and SWNTs have been among the most widely studied and

used carbon-based nanometerials due to their unique structural and electronic properties that

enable numerous industrial, medical and environmental applications (Ke and Qiao 2007; Lu and

Liu 2006; Rutherglen and Burke 2007; Wang et al. 2007). However, the low solubility of these

carbon-based nanomaterials in aqueous solutions has delayed their use, while stimulating

research on dispersion in aqueous solutions (Lee and Kim 2008; Matarredona et al. 2003;

Mitchell and Krishnamoorti 2007; Priya and Byrne 2008). For instance, methods have been

proposed for preparation of fullerene suspension in water using a step-wise approach that

includes dissolution into organic solvent followed by a back extraction into water (Degushi et al.

2001; Lyon 2006). Also, with the aid of poly(vinylpyrrolidone) or PVP, fullerenes (C60) could be

suspended in water at concentrations as high as 400 μg/mL (Yamakoshi et al. 1994).

Additionally, it has been reported that SWNTs can be dispersed in a number of aqueous

surfactant solutions such as sodium dodecyl sulfate (SDS), sodium dodecylbenzene sulfonate

(NaDDBS), and Triton X-100 (Wang et al. 2004; Zhang et al. 2005). Unfortunately, although the

surfactant induced functionalization of SWNTs enhances their suspension/dispersion, it can also

affect their inherent properties (Garg and Sinnott 1998), and likely, modify their interactions

with and impacts on living cells.

It is estimated that the worldwide market for nano-products would reach $1 trillion by

2015 (Roco 2005). This anticipated large production of nanomaterials and their likely

widespread use could lead to new environmental problems, such as new classes of toxins and

related environmental hazards (Masciangioli and Zhang 2003). With regard to human health, the

30

nano size enables nanoparticles to penetrate the human body via respiratory routes, skin or

digestive system exposures and persist without being phagocytosed (Karakoti et al. 2006).

Nanomaterials could even distribute throughout out the entire body, cross the blood-brain barrier,

and reach the olfactory bulb and the cerebellum (Braydich-Stolle et al. 2005). At the cellular

level, nanoparticles could interact with the cell lipid-bilayer membrane or other membrane

receptors, or enter the cells passively or actively (Tetley 2007). Realizing the toxicological

effects of nanoparticles, scientists have tested and demonstrated the toxicity of various

nanoparticles through both in vitro and in vivo assays (Jeng and Swanson 2006; Limbach et al.

2007; Soto et al. 2005; Usenko et al. 2007). Although inconclusive in most instances, the toxicity

of nanomaterials is hypothesized to result from a variety of effects including (i) the chemical

composition and the ability of a given nanoparticle to release free radicals, (ii) the particle size

and geometry, (iii) surface area and reactivity, (iv) surface treatments or modifications, (v) the

degree of aggregation/agglomeration, and (vi) particle electrostatic attraction potential (Gwinn

and Vallyathan 2006; Lanone and Boczkowski 2006; Long et al. 2007; Nel et al. 2006; Smith et

al. 2007; Tang et al. 2007).

As more and more nano-products are manufactured and used, chances of them being

released to the environment at different stages of their life cycle would also increase (see Figure

1-1). Recently, concerns over the potential environmental impacts of carbonaceous and metallic

nanomaterials have led to extensive toxicological studies (Braydich-Stolle et al. 2005; Ji et al.

2007; Lyon et al. 2006; Manna et al. 2005; Oberdorster 2004; Papis et al. 2007). However,

results obtained from most of these toxicity experiments are not easily used to predict the actual

effects of nanomaterials in natural systems (Fischer and Chan 2007).

31

In this study, a wide variety of carbon- and metal-based nanomaterials is screened for their

potential toxicity using three different small-scale micro-biotests (i.e. MetPLATE, Ceriodaphnia

dubia and Pseudokirchneriella subcapitata tests). The ultimate objective is to identify

nanomaterials with toxicity effects and use such toxic nanoparticles in further investigations

focusing on their potential impacts in natural systems.

2.2 Materials and Methods

2.2.1 Chemicals

Besides water, different solvents and surfactants used in the preparation of MNs

suspensions could lead to erroneous toxicity results. This is because certain solvents/surfactants

can be highly toxic to some organisms. To eliminate this uncertainty, the toxicity of 6 surfactants

and 2 organic solvents commonly used to obtain highly dispersed carbon-based nanomaterial

suspensions was determined first (see Table 2-1). Besides the poly(vinylpyrrolidone) (PVP) and

tetrahydrofuran (THF) which were obtained from Fisher Scientific (Atlanta, GA, USA), the rest

of the tested surfactants were purchased from Sigma-Aldrich (St. Louis, MO, USA). For toxicity

tests, Nanopure® water was used to prepare a concentrated solution of 50 g/L for PVP, and 10

g/L solutions for each of the other surfactants. These stock solutions were then used to prepare

necessary dilutions for toxicity tests. THF solutions were prepared by direct dilution of aliquot

volumes into Nanopure® water to produce samples with increasing concentration gradients and

then tested for toxicity.

2.2.2 Preparation of Nanomaterial Suspensions

Single walled carbon nanotubes (SWNTs) suspensions were prepared using an initial mass

of ~ 40 mg of raw SWNTs (Rice HPR 145.1, Rice University, Houston TX). The SWNTs were

mixed with 200 mL of an aqueous Gum Arabic (GA) surfactant solution (1 wt. %) to produce an

initial concentration of ~200 μg/mL. To obtain a highly dispersed suspension from this initial

32

solution, the mixture was homogenized using a high-shear IKA T-25 Ultra-Turrax mixer for ~

1.5 h followed by ultrasonication using a Misonix S3000 for 10 min. The mixture was then

ultracentrifuged at 20,000 rpm (Beckman Coulter Optima L-80 K) for 2.5 h. The supernatant was

carefully separated from the aggregated nanotubes at the bottom of the tube. For the produced

suspension, visible absorbance spectra of SWNTs were recorded by a Nano-Fluorescence

Nanospectrolyzer (Applied NanoFluorescence, LLC), and the concentrations of SWNTs

suspensions determined by absorbance at 763 nm.

The preparation method for aqueous suspensions of fullerenes (nC60) was adapted from

Deguchi et al. (2001). Briefly, ~15 mg of C60 (99.5%, Term-USA) were added to 500 ml of THF,

bubbled with ultra-high purity (UHP) nitrogen for 1 h to remove oxygen, and then sealed and left

stirring at room temperature for 24 h. Excess solids were later filtered out using a 0.45 μm PTFE

membrane filter, resulting in a transparent pink solution. The filtrate was then added to equal

amount of water in a container placed in a water bath and purged with UHP-N2 until all THF was

evaporated. Next, the obtained solution was vacuum-filtered through 0.45 μm cellulose into a

flask and stored in the dark. To determine the final concentration of the obtained aqueous

fullerene suspension, the fullerene suspension, a 2% NaCl solution and toluene were mixed in a

1:1:2 ratio and sonicated for 10 minutes. After separation of the aqueous and organic phases, the

upper toluene layer was withdrawn for absorbance measurement at 334 nm (Deguchi et al. 2001)

using a Hach DR/4000U Spectrophotometer.

The nano-metals, nano-copper, nano-silver, nano-nickel, and nano-cobalt used in this study

were provided gratis by Quantum Sphere Incorporated (Santa Ana, CA, USA). Nano-aluminum

powder was purchased from NovaCentrix (Austin, TX, USA). Prior to use, nanometallic

particles were characterized as received in the Particle Engineering Research Center at the

33

University of Florida, including specific surface area, particle size distribution (PSD), and zeta

potential (ζ). Density and specific surface area were measured using a Quantachrome Nova 1200

(Quantachrome, Syossett, NY, USA), and zeta potential was measured using a Zeta Reader Mk

21-II (Zeta Potential Instruments, Inc). Particle size distributions were measured with a Coulter

LS 13 320. To evaluate their toxicity, 200 mg/L suspensions were dissolved separately in

nanopure water, and shaken at room temperature for 48 hours.

In each case, preliminary experiments were performed with a range of doses up to 1 g/L

for the surfactants and PVP. Concentrations above 1 g/L were not tested, since they are rarely

encountered in the environment. According to Deguchi et al. (2001), less than 1 ppm of THF

remains in nC60/water suspension when using his proposed method. Therefore, three THF

concentrations were chosen (0.1 ppm, 1 ppm, and 10 ppm) to study the direct toxicity of THF to

Ceriodaphnia dubia.

2.2.3 96-hour Algal Chronic Toxicity Assay Using Pseudokirchneriella subcapitata (Selenatastrum capricornutum)

The preliminary algal assay procedure (PAAP) culture medium was prepared from stock

solutions according to EPA standard method (USEPA 2002), which includes three groups of

salts: major salts, trace salts and micro salts (see PAAP chemical composition in Table 2-2). The

pH of the culture medium was adjusted to 7.5 ± 0.1 with 0.1N NaOH or 0.1N HCl and then

filtered through a 0.45 μm and sterilized by autoclaving. A pure culture of P. subcapitata was

obtained from Hydrosphere Research (Alachua, FL) and grown in PAAP medium with EDTA at

around 25°C. Light source (86 ± 8.6 μE m-2s-1 or 400 ± 40ft-c) and continuous aeration were

provided 24 hours per day. New cultures were prepared every week under sterile conditions by

transferring approximately 20-30 mL of the mature cultures to 1-2 L of fresh sterile media.

34

Toxicity tests were performed in autoclaved 125 ml Erlenmeyer flasks according to EPA

(2002). All sample dilutions (i.e. culture media spiked with increasing concentrations of

individual tested nanomaterials—see Appendix A) and negative controls were run in triplicates

and inoculated with 1ml of a 4 to7 day old algal cultures. They were pre-concentrated to obtain a

cell density of about 5 × 105 cells/ml PAAP. All flasks were placed under the fluorescent lights

in the same growth conditions. The growth inhibition after 96-h was determined by measuring

the concentrations of chlorophyll (chl. a) using a TurnerR QuantechTM digital filter fluorometer.

In this case, the measured toxicity endpoint was growth inhibition (I). On the basis of the

obtained chl. a concentrations, the IC50 (i.e. concentration that inhibits 50% of algal growth) was

determined by first plotting sample concentrations (X-axis) versus the percent inhibition (I%)

determined using Eq. 2-1 (Y-axis), and then performing a regression analysis in the linear

portion of the obtained line, from which the IC50 was determined using Eq. 2-2.

100].[].[

1 ×⎟⎟⎠

⎞⎜⎜⎝

⎛−=

control

sample

achl achl

% I (2-1)

SY

IC erceptint−=

5050 (2-2)

Where [chl.a] corresponds to measured chlorophyll concentration; Yintercept is the inhibition value

corresponding to the intercept of the above regression line with the Y-axis; and S is the slope of

the linear regression.

2.2.4 48-hour Acute Toxicity Assay Using Ceriodaphnia dubia as Test Model Organism

Moderately hard water (MHW) prepared following the EPA standard method (USEPA

2002) was used as culture media in this test. Pure culture of C. dubia were obtained from

Hydrosphere Research (Alachua, FL) and kept in 1L beakers containing 500mL of MHW in a

PervicalTM model # E-30 BX environmental chamber at 25 ºC with constant aeration. The photo

35

period was 16 hr light/8 hr dark. C. dubia was fed with concentrated P. subcapitata cells and

YCT (made from yeast, cereal leaves and trout chow). The daphnia were fed every other day

with 6.67 mL YCT and 6.67 mL algae solution/L culture. The culture medium was also changed

every other day. Neonates of less than 24 hours were separated from adults daily and used for

testing.

Acute toxicity tests were performed according to EPA’s protocol (USEPA 2002). MHW

served as negative control and as the diluent to prepare media with increasing concentrations of

tested nanomaterials. Neonates less than 24 hours were separated from adults and fed 2 hours

prior to test start. For each test, groups of 5 daphnia neonates each were transferred into several

30 ml plastic cups containing 20 ml of MHW (controls) or MHW plus nanomaterials

(treatments—see Appendix A). There was no feeding during the tests and survival was

determined visually after 48 hours using death and/or immobilization as endpoint. To determine

the LC50 (i.e. concentration that causes the death of 50% of test organisms), the organism were

exposed to at least five different and increasing concentrations in triplicates (Appendix A).

Finally, the LC50 values of daphnia test and associated 95% confidence intervals (CIs) were

calculated by the probit analysis (U. S. EPA Probit Analysis Program, Ver. 1.5), and values were

considered different when the calculated CIs did not overlap.

2.2.5 MetPLATE Test

The MetPLATE test kit specific to heavy metal toxicity (Bitton et al. 1994) was used in

this study. The kit contains a bacterial reagent (an E.coli strain), a buffer, chlorophenol red

galactopyranoside (CPRG), which serves as the substrate for β-galactosidase, and moderately

hard water (MHW) used as diluent (Bitton et al. 1994). The kit also includes a positive control

containing copper (Cu). Briefly, the bacterial reagent was hydrated with 5-ml of MHW and

thoroughly mixed by vortexing. Then, a 900 μl aliquot of the MN suspension (or 900 μl of MHW

36

for negative controls) was added to a test tube containing 100 μl of the above-described bacterial

reagent. The test tubes were then vortexed and incubated for 1.5 hours at 35 ºC. Following the

incubation period, a 200-μl aliquot was transferred to a 96-well microplate to which 100 μl of the

enzyme substrate (i.e. CPRG) was added. After mixing, the microplate was incubated at 35 ºC to

allow color development. The response was quantified at 570 nm using a Multiskan microplate

reader. The measured toxicity endpoint was inhibition of color development, and the IC50 values

for MetPLATE assays were determined as described earlier for the IC50 in the algal test by

replacing [chl.a] with absorbance (see equations 2-1 and 2-2). All treatments (see Appendix A)

were run in triplicates.

2.3 Results and Discussion

2.3.1 Characterization of Nano-metal Particles

Particle size distributions for the nano-metals tested are shown in Figure 2-1. While

aggregation resulted in increased mean particle diameter, nano-copper, nano-silver, and nano-

cobalt still had a significant fraction of particles with size <100 nm. Specific surface areas and

zeta potentials for different nanometal powders are shown in Table 2-2. Among the nano-metal

particles, nano-nickel had the largest surface area (50.6 m²/g), and although nano-silver particles

showed rather nanoscale particles (as defined by nanotechnology, <100 nm), the particles surface

area (14.5 m²/g) was relatively small compared to other nanometals.

The zeta potential is related to the stability of particle dispersions. Therefore, particles with

high zeta potential (negative or positive) are electrically stabilized, and nano-copper will have

the highest tendency to agglomerate in water as it has a very low zeta potential (ζ = -0.69 mV,

Table 2-2).

37

2.3.2 Toxicity of Solvents and Surfactants

Using the algal and daphnia tests described above (MetPlate not used because of its

specificity to heavy metal toxicity), the toxicity of a number of surfactants/solvent was

determined (Tables 2-4 and 2-5). Among the tested dispersing agents, only Gum Arabic (GA)

and PVP showed no toxicity at concentration as high as 1 g/L. SDS had the highest toxic

response with both tests (LC50 of 48-h daphnia test: 0.003 g/L, 95% CI = 0.002-0.005 g/L ; IC50

of 96-h algal test: 0.0738 g/L), followed by SDBS (LC50 of 48-h daphnia test: 0.01 g/L, 95% CI

= 0.008-0.012 g/L; IC50 of 96-h algal test: 0.0793 g/L), Triton X-15 (LC50 of 48-h daphnia test:

0.014 g/L, 95% CI = 0.007-0.024 g/L), Triton X-100 (LC50 of 48-h daphnia test: 0.026 g/L, 95%

CI = 0.025-0.028 g/L; IC50 of 96-h algal test: 0.0917 g/L), sodium cholate (LC50 of 48-h daphnia

test: 0.053 g/L, 95% CI = 0.045-0.064 g/L). The effect of different surfactants varied with both

concentration (e.g., Smith et al. 2007) and type of tested model organisms. The results indicate

that solution matrix used to suspend nanomaterials should be tested to select model test

organisms to determine the toxicity of a given nanomaterial. In this study, GA was identified as

one of the least toxic surfactants to model aquatic organisms selected for this study. The latter

was then used to prepare all SWNT suspensions discussed in this dissertation.

Tetrahydrofuran, a bio-aggressive solvent has been used in many studies to prepare

aqueous suspensions of fullerenes (C60). Using C. dubia, the toxicity of relatively low-level THF

solutions was assessed. Table 2-6 showed the mortality of C. daphnia exposed to THF

concentrations of 0.1, 1.0 and 10 ppm. These results imply that trace THF concentrations (~0.1

ppm) present in C60 suspensions (Degushi et al. 2001) should not be toxic to C. dubia, unless it

affects the properties of native C60 to render them toxic.

38

2.3.3 Toxicity of Tested Nanomaterials

2.3.3.1. Fullerene (C60)

The LC50 of the aqueous suspension of fullerenes (nC60) using the 48-hour daphnia test

was 0.395 mg/L (95% CI = 0.287-0.470 mg/L) (Table 2-7). This value compares quite well with

numbers reported by several other studies using freshwater model microorganisms (Table 2-8).

The calculated IC50 based on the 96-h algal test was 0.139 mg/L (Table 2-9).

In this study, specific mechanisms of C60-toxicity were not investigated. However, studies

reporting on the mechanisms of toxicity of several nanomaterials including C60 are available in

the literature. The toxicity of C60 results from oxidative stress induced by production of reactive

oxygen species (ROS) (Oberdorster 2004; Sayes et al. 2005; Yamakoshi et al. 2003), residual of

organic solvent used to disperse C60 particles (Gharbi et al. 2005; Henry et al. 2007), or

interaction between C60-particles and living cells (Usenko et al. 2007). The generation of ROS

(e.g. 1O2) by C60 through energy transfer has been detected in toxicological studies (Isakovic et

al. 2006) and the pathway of ROS production has been summarized as follows (Arbogast et al.

1991):

However, this mechanism is subject to an ongoing debate since C60 has also been identified

as antioxidant that can efficiently scavenge free radicals (Gharbi et al. 2005; Krusic et al. 1991;

Lin et al. 1999; Wang et al. 1999), in addition to the fact that the toxicity effects of C60 have also

been detected in the total absence of light (Isakovic et al. 2006; Sayes et al. 2004).

1C60 1C60* 3C60

* 1C60 3O2

1O2

Triplet Quencher (Q) hν

Intersystem Crossing

39

As stated earlier, C60 particles are hydrophobic and stable water suspension can be

obtained by use of organic solvent (e.g. THF). Although mass spectroscopy of solvent after this

procedure indicated no residual THF in solution (Sayes et al. 2005), it was shown that

approximately 10% THF (w/w) was intercalated into nC60 crystalline lattice after preparation

(Fortner et al. 2005). This trapped THF, although undetectable through analysis of the aqueous

phase, could have toxic implications when nanopaticles interact with living cells. The positive

aspect in favor of THF remains its actual concentration as demonstrated by Harhaji (2007) and

in this study, pure THF would not cause cell death at concentrations up to 10 μg/ml. Also,

Markovic (2007) attributed the higher toxicity of THF-nC60 compared to other fullerene

dispersions prepared with other surfactants/solvents to the less 1O2-quenching power of THF.

In this study, P. subcapiata cells were more sensitive to C60 toxicity than C. daphnia. It

could be because the exposure time for algae was longer and more ROS were produced in the

algal test set-up where artificial light was used.

2.3.3.2. Single-Walled Nanotubes (SWNTs)

As stated earlier, SWNTs suspensions used in this study were prepared in Gum Arabic and

will be referred to herein as SWNT-GA. Toxicity tests using SWNT-GA resulted in LC50 of 0.27

mg/L (95% CI = 0.229-0.294 mg/L) with the 48-h daphnia test and an IC50 of 0.769 mg/L with

the 96-h algal test (Table 2-7 and 2-9).

If released into aqueous systems, carbon nanotubes could be taken up by aquatic

organisms, such as particle feeders with potential negative effects (Roberts et al. 2007). Similar

to C60, SWNTs can also generate ROS and cause oxidative stress (Helland et al. 2007; Manna et

al. 2005), but their potential to induce oxidative stress is mainly due to the presence of metal

impurities such as Fe, Ni, Y and Co (Guo et al. 2007; Lanone and Boczkowski 2006). This is to

some extent supported by the results of toxicological studies comparing the effects of purified

40

(e.g. iron free) versus unrefined (Fe-contaminated) carbon nanotubes (Miyawaki et al. 2008;

Templeton et al. 2006). This is because electron capture by metal impurities induces the

formation of superoxide radical (O2·-) and additional ROS through a Fenton-like reaction (Nel et

al. 2006; Pulskamp et al. 2007). Shvedova et al. (2003) have suggested that iron residual in

unrefined SWNTs could catalyze the production of free radicals through the following reactions:

Fe2+ + O2 → Fe3+ + O2- (2-3)

Fe3+ + Ared- → Fe2+ + Aox (2-4)

O2- + O2

- + 2H+ → O2 + H2O2 (2-5)

H2O2 + Fe2+ → Fe3+ + OH- + ·OH (2-6)

LOOH + Fe2+ → Fe3+ + LO· + -OH (2-7)

On the contrary, Kang et al. (2007) demonstrated that highly purified SWNTs still caused

cell membrane damage, due to direct physical interactions between nanomaterials and cells. In

this case, the degree of particle aggregation plays an important role in toxicity, and refined

SWNTs could actually be more toxic than unrefined because they are better dispersed and could

pierce or permeate cell membranes (Tian et al. 2006).

The SWNTs used in this study were not purified, i.e. they contained a certain amount of

impurities, mostly iron (Dr. Ziegler, UF-Chemical Engineering Dept– personal communication).

Therefore, the toxicity detected here may result from any or a combination of the above listed

factors.

2.3.3.3. Metallic nanomaterials

Nanoscale copper and silver resulted in toxic responses regardless of the toxicity test used,

with the calculated EC50’s (concentration that causes 50% of the maximum effect) ranging from

41

0.4 mg/L (C. dubia) to 23.93 mg/L (MetPLATE) for nano-copper, and from 0.06 mg/L (C.

dubia) to 20.92 mg/L (MetPLATE) for nano-silver. The MetPLATE test generally gave higher

EC50 values (Table 2-10) because of its specificity to bioavailable metals. In addition,

MetPLATE is not very sensitive to nickel, cobalt and aluminum (Bitton et al. 1994). Amongst all

tested nanometals, nano-aluminum particles had the least toxic effects, with EC50 values of 3.91

mg/L with C. dubia and 8.29 mg/L with P. subcapitata.

Due to their small size, nano-copper particles can easily distribute throughout the body of

vertebrates into blood, brain, lung, heart, kidney, spleen, liver, intestine and stomach; and their

large specific surface area could result in high reactivity leading to tissue/organ dysfunctions

(Chen et al. 2006). It has been suggested that once inside organism’s stomach, nanocopper

particles could react with protons (H+) from gastric juice and become quickly ionized, resulting

in an overload of ionic copper (Meng et al. 2007). In this case, the depletion of H+ would then

lead to a massive formation of HCO3- and induction of metabolic alkalosis. Finally, nanocopper

(nano-Cu) toxicological effects are probably not due solely to dissolution properties as

nanocopper particles are capable of catalyzing the production of reactive species as well (Griffitt

et al. 2007).

Silver ions have long been used as efficient antimicrobial agent and similar effects are

anticipated for nano-silver particles. Lok et al. (2006) suggested that the antimicrobial activity of

nano-silver particles is associated with Ag+, formed on the particle surface due to partial

oxidation. In a study of nine different nanomaterials, Soto et al. (2007) found nano-silver to be

amongst the highly toxic nanoparticles. In addition, he found no correlation between toxicity and

particle specific surface areas, which is in agreement with our results. The mechanism of silver

toxicity is rather well understood. Ag+ could enter cells via apical sodium channels because it

42

mimics Na+ (Glover and Wood 2005). Once inside the cell, Ag+ inhibits sodium and chloride

transport and consequently disturbs ion homeostasis (Glover et al. 2005). Similar to most

nanoparticles, the nanosize aggravates the toxicity of silver with the smaller particles being more

toxic.

Overall, no strong relationship between particle surface area (listed in Table 2-2) and

particle toxicity for studied model organisms is apparent. However, particle size is one of the

possible influencing factors, as particles with the smallest particle sizes (nano-copper and nano-

silver) showed the highest toxicity. Compared to copper and silver nanoparticles, less effort has

been made toward understanding the aquatic toxicity of other nanometals. Zhang et al. (2006)

have compared the toxicity of ultrafine and standard-sized nickel particles and found

significantly higher effects with the ultrafine nickel. However, they didn’t elucidate the toxicity

mechanism but speculated that generation of free radicals might be the reason. Although

MetPLATE didn’t show negative results with nano-nickel in our study, Liu et al. (2007)

observed a release of Ni from unrefined SWNTs in aqueous solutions and the mobilization was

more pronounced under acidic conditions, indicating nickel corrosion by acid.

Cobalt is known for its high catalytic activity for the degradation of hydrogren peroxide

and generation of ROS, and while cell membranes are commonly repulsive to toxic metal ions,

they tend to facilitate the uptake of nanosize particles, which later dissolve and damage the cell

(Limbach et al. 2007). Papis et al. (2007) found that nanocobalt could activate cellular pathways

of defense and repair mechanisms but no molecular mechanism was concluded.

A study by Braydich-Stolle et al. (2005) involved cytotoxicity testing of aluminum

nanoparticles. Like our results, they observed much less effect with nano-aluminum (nano-Al)

43

compared with nanosilver (nano-Ag) and nanomolybdenum (nano-Mo). Hussain et al. (2005)

revealed that nanoaluminum didn’t display toxicity up to the concentration of 100 ppm.

For all the nanometals, their tested concentrations in the literature were generally higher

than the EC50 values we obtained from C. daphnia and algal tests, indicating that these two tests

are very sensitive and appropriate for toxicity screening test.

2.4 Conclusions

In this study, we assessed the potential toxicity of various nanoparticles and evaluated the

potential effects of different particle dispersing agents on model test organisms. The results show

that surfactants common in the manufacturing and dispersion of carbon-based nanomaterials can

add to the toxicity of tested nanoparticles. Accordingly, toxicity tests evaluating the biological

impact of nanomaterials should take into account the effect of not only the chemical impurities

associated with the fabrication processes, but also the role of the fluid used to maximize the

dispersion effect. This study also demonstrates the need for multiple toxicity tests in assessing

the potential implications of nanomaterials. Depending on the sensitivity of test organisms and

route of exposure, a nanomaterial can be highly toxic to a given test organism while not toxic at

all to another. Finally, the observed toxic responses in lower trophic level aquatic organisms

support the need for vertebrate toxicity-based studies. In addition, the investigation of the ability

of nanomaterials to bioaccumate and ultimately biomagnify along food chains should be

examined.

44

Figure 2-1. Size distribution of selected metallic MN samples as obtained from commercial sources

45

Table 2-1. Tested surfactants and solvents and their chemical compositions. The chemical formula of GA is not included because it consists of multiple chemical compounds

Surfactants and Solvent Chemical Composition Gum Arabic from acacia tree (reagent grade) NA** Dodecylbenzene-sulfonic acid sodium (SDBS) C18H29NaO3S Sodium dodecyl sulfate (SDS) C12H25O4S.Na Sodium cholate hydrate (98%) C24H39NaO5· xH2O

Triton X-15 glycols, polyethylene, mono ((1, 1, 3, 3-tetramethylbutyl) phenyl) ether

Triton X-100 (C2H4O)nC14H22O Tetrahydrofuran (THF) C4H8O Poly(vinylpyrrolidone) (PVP) (C6H9NO)n

46

Table 2-2. Characteristics of metallic nanoparticles used in toxicity experiments in this study Nanometals Tested for Toxicity

Specific Surface Area (m²/g)

Zeta Potential (ζ) (mV)

Nano-silver 14.53 -27.0 Nano-copper 30.77 -0.69 Nano-aluminum 27.26 + 18.2 Nano-cobalt 36.39 + 17.8 Nano-nickel 50.56 + 21.9

47

Table 2-3. Chemical composition of the preliminary algal assay procedure (PAAP) culture medium

Macronutrient Concentration (mg/L) NaNO3

MgCl2·6H2O CaCl2·2H2O

MgSO4·7H2O K2HPO4 NaHCO3

25.5 12.2 4.41 14.7 1.04 15

Micronutrient Concentration (μg/L) H3BO3

MnCl2·4H2O ZnCl2

CoCl2·6H2O CuCl2·2H2O

Na2MoO4·2H2O FeCl3·6H2O

Na2EDTA·2H2O Na2SeO4

185 416 3.27 1.43 0.012 7.26 160 300 2.39

48

Table 2-4. Concentrations of tested surfactants resulting in lethal effect on 50% of the population (LC50) based on the 48-h Ceriodaphnia dubia assay.

Tested Surfactants LC50 (g/L)

CI

(g/L) Gum Arabic >1 Dodecylbenzene-sulfonic acid sodium 0.01 0.008-0.012 Sodium dodecyl sulfate 0.003 0.002-0.005 Sodium cholate 0.053 0.045-0.064 Triton X-15 0.014 0.007-0.024 Triton X-100 0.026 0.025-0.028 Poly(vinylpyrrolidone) (PVP) >1 * CI = confidence interval; and LC50 values shown as >1 indicate that no toxicity effect was observed for surfactant tested at concentration up to 1g/L.

49

Table 2-5. Surfactant concentrations resulting in the inhibition of 50% of growth (IC50) in a 96-h P. subcapitata chronic toxicity assay.

Tested Surfactants

IC50 (g/L)

Gum Arabic >1 Dodecylbenzene-sulfonic acid sodium 0.0793 Sodium dodecyl sulfate 0.0738 Sodium cholate >1 Triton X-15 >1 Triton X-100 0.0917 Poly(vinylpyrrolidone) (PVP) >1 * IC50 values shown as > 1 indicate that no toxicity effect was observed for surfactant tested concentration up to 1g/L.

50

Table 2-6. Percent mortality of C. dubia exposed to solutions with increasing THF concentrations in 48-h accute toxicity assay.

Concentration of THF (ppm) Mortality of C. Dubia (%) 0.1 0 1 13.33 ± 11.55 10 26.67 ± 11.55

51

Table 2-7. Concentrations of tested metal- and carbon- based nanoparticles resulting in lethal effect on 50% of the population (LC50) based on the 48-h Ceriodaphnia dubia assay.

Tested Nanomaterials LC50 (mg/L)

Confidence Interval(mg/L)

Nano-silver 0.066 0.053-0.075 Nano-copper 0.401 0.329-0.456 Nano-cobalt 1.647 1.522-1.734 Nano-nickel 0.658 0.549-0.984 Nano-aluminum 3.906 3.321-4.410 Fullerenes (C60) 0.395 0.287-0.470 SWNT suspended in Gum Arabic (GA) 0.74 0.63-0.81

52

Table 2-8. Examples of published EC50 values for fullerenes (C60), single-walled carbon nanotubes (SWNTs), and nano-copper (nano-Cu) on daphnia and zebrafish

Tested Nanomaterials Test Model Organisms EC50 References

C60 Daphnia sp 460 ppb Lovern and Klaper 2006 C60 Embryonic zebrafish <200 ppb Usenko et al. 2007 C60 Daphnia magna 800 ppb Zhu et al. 2006

SWNTs Daphnia magna 10 - 20 ppm Roberts et al. 2007 Nano-copper Danio rerio (Zebrafish ) 1.56 ppm Griffitt et al. 2007

53

Table 2-9. Concentrations of tested metal- and carbon- based nanoparticles resulting in growth inhibition of 50% of the population (IC50) based on the 96-h P. subcapitata chronic toxicity assay.

Tested Nanomaterials IC50 (mg/L)

Nano-silver 0.196 Nano-copper 0.542 Nano-cobalt 0.707 Nano-nickel 0.348

Nano-aluminum 8.285 Fullerenes (C60) 0.139

SWNT suspended in Gum Arabic (GA) 2.11

54

Table 2-10. Concentrations of tested metal- and carbon- based nanoparticles resulting in 50% inhibition of color development in MetPLATE test. Tested Nanomaterials IC50 (mg/L)

Nano-silver 20.92 ± 0.61 Nano-copper 23.93 ± 0.51 Nano-cobalt Not detected Nano-nickel Not detected

Nano-aluminum Not detected Fullerenes (C60) Not detected

SWNT suspended in Gum Arabic (GA) Not detected * Besides nanosilver and nanocopper, no toxicity effect was observed with all the other tested nanoparticles.

55

CHAPTER 3 TOXICITY OF SELECTED MANUFACTURED NANOMATERIALS DISPERSED IN NATURAL WATERS WITH GRADIENTS IN IONIC STRENGTH AND DISSOLVED

ORGANIC MATTER CONTENT

3.1 Introduction

The success of nanotechnologies and the resulting widespread production and use of

manufactured nanomaterials (MNs) will likely lead to their introduction to natural systems.

Rivers/lakes are likely to behave as primary sinks as they integrate pollutants from atmospheric

deposition, terrestrial surface runoffs, and groundwater discharges. In the past few years, several

studies focused on the potential toxicity of MNs, using existing experimental procedures, which

are not always adequate to assess the environmental implications of these new pollutants. If

current efforts to understand the biological effects of MNs on human health have relied on

appropriate techniques such as inhalation exposures, studies emphasizing the environmental

implications use primarily very drastic approaches to facilitate the contact between MNs and test

model organisms.

Man made nanoparticles can enter aquatic systems through direct discharges (e.g.

accidental spills and surface runoffs), but also through industrial and/or domestic wastewater

effluents. For instance, the use of nano-silver (nano-Ag) particles in washing machines (e.g.

washers manufactured by Samsung Electronics) could release nano-Ag particles into sewer

systems (Christen 2007; Lovern et al. 2007). Since current wastewater treatment plants are ill-

equipped for MNs removal, these MNs would ultimately enter natural waterways. So far,

laboratory studies on toxicity of MNs have focused on their impacts on model organisms (Du et

al. 2008; Fortner et al. 2005; Griffitt et al. 2007; Lovern et al. 2007; Rameshbabu et al. 2007;

Wei et al. 2007; Yoon et al. 2007; Zhu et al. 2006). Although highly informative, results from

most of these toxicity studies are usually obtained under conditions that are far different from

56

those encountered in natural systems. For example, the use of drastic mixing methods (e.g.

ultrasound, sonication, etc), toxic solvents/surfactants (e.g. toluene, sodium dodecyl sulfate,

triton X-100, etc) or a combination the above techniques in the preparation of MNs suspensions

lead to ideal MN-dispersion levels, which are unfortunately unattainable through direct

introduction of MN into natural waters. Accordingly, it is quite difficult to extrapolate such

laboratory results to natural systems.

To address this gap in current knowledge on the environmental implications of

nanotechnologies, this study was designed to investigate the potential impacts of selected MNs

that previously showed toxicity effects (i.e. C60, nano-copper and nano-silver), when suspended

directly in natural waters. The rationale for this study is that MNs released directly to waterways

would likely be impacted primarily by solution composition and weak mechanical dispersion

processes such as waves and bioturbation. The type and extent of the biological effects of MNs

under these specific conditions could therefore be different from those currently reported in the

literature.


3.2.1 Collection of Water Samples

Water samples used in this study were collected from the Suwannee River (Fig. 3-1),

which provides an excellent opportunity to study the effect of water chemistry on potential

toxicity of MNs. The Suwannee River system contains three linked hydrologic units (upper,

middle and estuary), each providing distinct hydrological characteristics and gradients in

dissolved organic carbon (DOC) and ionic strength (I). In the upper watershed, confinement of

the Floridan aquifer provides surface drainage and sources of organic carbon from wetlands (e.g.

the Okefenokee swamp in southern Georgia, Fig. 3-1). The boundary between the upper confined

and middle unconfined watershed is a geomorphic feature called the Cody Scarp, below which

57

ground water returns to the surface from many large springs, increasing the ionic strength of

surface waters. Finally, the river delta leads to the Gulf of Mexico, and provides sites for

collection of water samples with lower DOC and higher salinity.

Three samples were collected from (i) the headwaters or SR1, (ii) river mid-section or

SR2, and (iii) the river delta or SR3 (Figure 3-1) in DI-prewashed and site-water rinsed 2 liter

PE-containers. Soon after collection, the samples were placed in coolers and transported back to

UF campus. In the laboratory, samples were filtered (0.45 μm) to remove large size debris and

kept refrigerated at 4°C until used in laboratory experiments.

3.2.2 Preparation of Nanomaterial Suspensions in Collected Water Samples

Suspensions of pre-selected three MNs (C60, nano-Ag, and nano-Cu) were prepared in

volumetric flasks by mixing 200 mg of each of the MNs in 200ml of water. In addition,

Nanopure® water was also used as matrix to the tested MNs. Obtained mixtures were gently

agitated using a New Brunswick G24 horizontal shaker (Edison, NJ) to mimic the mechanical

mixing effect of waves. Following a 7-day of gentle mixing period, MNs were filtered to remove

agglomerated particles with size > 1.6 μm (1.6 μm Whattman filter paper, Florham Park, NJ),

and the filtrate collected for determination of MNs concentrations and use in toxicity

experiments.

3.2.3 Determination of MNs Concentrations in Prepared Suspensions

The concentration of C60 in the filtrates was determined by measuring the absorbance at

336 nm using a Hach DR/4000U spectrophotometer, and then correlating to standard solution of

known concentration prepared in tetrahydrofurane (THF) (Lyon et al. 2006). Following the acid

digestion of water MN-suspensions by a mixture of concentrated nitric and hydrofluoric acids,

concentrations of Cu and Ag in aqueous MN-suspensions were determined in comparison with

58

non-spiked natural water samples by inductively coupled plasma, atomic emission spectroscopy

(ICP-AES).

3.2.4 Toxicity of MNs Suspended in Natural Waters

Two toxicity tests, the 48-hr Ceriodaphnia dubia assay and MetPlateTM were used in this

study. The concentration ranges of tested MNs are given in Appendix B, and a brief description

of used toxicity tests is given below.

3.2.4.1 The Ceriodaphnia dubia assay

Moderately hard water (MHW) (USEPA 2002) was used as culture media for C. dubia.

Pure culture of C. dubia were obtained from Hydrosphere Research (Alachua, FL) and kept in

1L beakers containing 500mL of MHW in a PervicalTM model # E-30 BX environmental

chamber at 25 ºC with constant aeration. The photo period was 16 hr light/8 hr dark. C. dubia

was fed with concentrated P. subcapitata solution and a mixture made from yeast, cereal leaves

and trout chow (YCT). The daphnia were fed every other day through the addition of 6.67mL of

YCT and 6.67mL algal suspension per liter daphnia culture. Neonates of less than 24 hours were

separated from adults daily and used for testing.

Acute toxicity tests were performed according to EPA’s protocol (USEPA 2002), in which

MHW serves as diluent and as negative control. MNs tested for toxicity were added to the

culture as natural water suspensions to produce an increasing concentration gradient. Neonates

less than 24 hours were separated from adults and fed 2 hours prior to test start. Groups of five

neonates were then transferred into 30 ml plastic cups containing 20 ml of MN-suspensions.

There was no feeding during the tests and survival was determined visually after 48 hours. The

measured toxicity endpoint was death and/or immobilization, and for each tested MN, at least

five concentrations were used for LC50 determination (i.e. concentration resulting to the

death/immobilization of 50% of the population). All experiments were run in triplicates, and the

59

95% confidence intervals (CIs) associated with obtained LC50 values were calculated by the

probit analysis (U. S. EPA Probit Analysis Program, Ver. 1.5), and considered different when not

overlapping.

3.2.4.2 MetPLATE test

The MetPLATE test kit specific to heavy metal toxicity (Bitton et al. 1994) was also used

in this study. This kit contains a bacterial reagent (an E.coli strain), buffer, chlorophenol red

galactopyranoside (CPRG), which serves as substrate for β-galactosidase, and moderately hard

water (MHW) as a diluent (Bitton et al. 1994). The kit also includes a positive control (Cu).

Briefly, the bacterial reagent was rehydrated with 5-ml of diluent and thoroughly mixed by

vortexing. Next, 900 μl aliquot of the MN-suspensions prepared in natural waters was added to a

test tube containing 100 μl of the above-described bacterial reagent. In these assays, MHW

containing natural waters without MNs served as negative controls. The test tubes were vortexed

and then incubated for 1.5 hours at 35 ºC. A 200-μl aliquot of the above mixture was then

transferred to a 96-well microplate to which 100 μl of CPRG, the enzyme substrate, was added.

The microplates were then shaken and incubated at 35 ºC for color development. The response

was quantified at 570 nm using a Multiskan microplate reader. The measured toxicity endpoint

was inhibition of color development, and the IC50 (i.e. concentration that inhibits 50% of color

development) was determined by first plotting sample concentrations (X-axis) versus the percent

inhibition (%I) determined using Eq. 3-1 (Y-axis), and then performing a regression analysis in

the linear portion of the obtained line, from which the IC50 was determined using Eq. 3-2.

1001 ×⎟⎟⎠

⎞⎜⎜⎝

⎛−=

control

sampleA

A% I (3-1)

SY

IC erceptint50

50 −= (3-2)

60

Where A is the absorbance; Yintercept is the inhibition value corresponding to the intercept of

the above regression line with the Y-axis; and S is the slope of the linear regression. All samples

were run in triplicates.


3.3.1 Characterization of Water Samples

The chemical composition of the three water samples is presented in Table 3-1. Sample

SR1 was characterized by a very high DOC content (47.71 mg C/L) and low ionic strength (0.94

mM). The second sample (SR2) had an ionic strength of 3.34 mM and a DOC content of 10.18

mg C/L. The sample collected along the salinity gradient in the river delta had a much higher

ionic strength (475.17 mM). Although the DOC concentration of SR3 samples could not be

determined in our laboratory due to the high chloride concentration of the samples, the reported

DOC in this portion of the Suwannee River was about 2.3 mg C/L (Del Castillo et al. 2000).

Overall, the ionic strength of these water samples shows a steep gradient indicative of the

downstream in increasing concentration of inorganic carbon, primarily bicarbonate. Other major

ions do not vary significantly between SR1 and SR2. Finally, in all three samples the original

concentrations of total-Ag and total-Cu were below the detection limits (<10 μg/L) of the ICP-

AES used in this study.

3.3.2 Total Concentration of Dispersed Nanomaterials

The concentrations of C60 and nanometals in collected filtrates were analyzed by

spectrophotometry and ICP-AES respectively (Fig 3-2), representing the fractions of suspended

nanoparticles passing through a 1.6 μm filter. Measured concentration varied significantly with

the water chemical composition. All nanomaterials were rather well-dispersed in SR1 with

concentrations ranging from ~ 0.54 mg/L to ~12.58 mg/L. The lowest concentrations were

61

obtained in SR3 (from 0.04 ± 0.02 mg/L to 0.66 ± 0.23 mg/L). Suspension of the same

nanomaterials in Nanopure water yielded final concetrations ranging from ~ 0.13 mg/L to ~1.67

mg/L after filtration.

The highest Ag concentration was observed in Nanopure-water suspensions followed by

SR1 and SR3, but no statistically significant difference between values measured in SR1 and

SR3 samples was found. Measured Ag levels in SR2 samples were negligible. Nano-silver has

previously been found to be partially oxidized when exposed to oxygen and Ag+ will form on the

particle’s surface (Lok et al. 2007). The positive charges can probably keep some silver

nanoparticles from forming large aggregates and thus go through the filter membrane. Fan and

Bard (2002) demonstrated that oxidized silver films could be solubilized as Ag (I) and Ag (0)

species. This may account for the larger fraction of suspended or dissolved Ag in DI water as

opposed to other waters (Fig. 3-2). Soluble Ag is known to have a high affinity with organic

matter (logK of 7.5, Bury et al. 2002; 9.0-9.2, Janes and Playle 1995) and chloride (logKAgCl of

9.8, Lide and Frederikse 1997). This is probably also true for the Ag nanoparticles. They can be

stabilized by natural ligands and polymers in water (Mayer et al. 1999; Mucalo et al. 2002;

Pastoriza-Santos and Liz-Marzán 2002; Zhou et al. 2006). Accordingly, steric stabilization effect

from the organic ligand layers coating the nanoparticles can help prevent particle aggregation.

Therefore, in SR2, removal of Ag is most likely due to low DOC levels. These results suggest

that organic matter and solution chemical composition can influence the suspension of nano-

silver particles in water, although the quantitative relationships remain unknown.

In contrast to nano-Ag, Nanopure water had lowest Cu-level while most Cu nanoparticles

were dispersed in SR1 water. Similar to nano-Ag, soluble Cu tends to form very stable

complexes with natural organic ligands, e.g. log K'Cu(II)L varies from 10 to 14 (Leal and Van den

62

Berg 1998; Pastoriza-Santos and Liz-Marzán 2002; Xie et al. 2004). The detected total-Cu

concentration showed a positive correlation with DOC concentrations (Fig. 3-3), indicating

strong binding effects of NOM on nano-copper particles. This trend is in agreement with

previously published work (Cantwell and Burgess 2001; Wen et al. 1999).

Although it is well accepted that C60 has a very low solubility in water (Heymann 1996;

Ruoff et al. 1993), an approximate 3.09 ppm of C60 nanoparticles was suspended in SR2. This is

probably because C60 particles could be negatively charged (ξ up to -40 mV) in aqueous systems

(Brant et al. 2005; Deguchi et al. 2001). However, when in contact with even weak electrolyte

solutions, suspensions of C60 would tend to become destabilized and agglomerate due to surface

complexation with cations (Brant et al. 2005), a potential explanation for lower solubility of C60

in SR3 samples. Like nanometals, it has also been previously shown that carbon nanomaterials

can be stabilized by organic matter or natural polymers (Chen and Elimelech 2007; Espinasse et

al. 2007; Hyung et al. 2007). Deguchi (2007) has demonstrated the stabilization of C60 by human

serum albumin and showed that dispersion stability of C60 nanoparticles in complex

environments could be different from simple model systems (Deguchi et al. 2007) common to

the majority of current experimental studies. Espinasse et al. (2007) stated that humic and fulvic

acids could enhance the stability of colloidal nC60 by increasing the negative charge on the

surface. However, if true, the higher concentration of C60 measured in SR2 as compared to SR1

cannot be easily explained. It is therefore proposed that acidity in SR1 (pH = 4.7) leads to the

interaction of H+ with negatively charged C60 surfaces, resulting in net surface charges that favor

aggregation.

63

3.3.3 Evaluation of Acute Toxicity of Nanomaterials Suspended in Natural Waters to Ceriodaphnia dubia

As particle-feeding organisms, C. dubia were chosen as model test organisms because of

their high sensitivity to toxic fine particles. Although no toxicity to C. dubia was detected

through exposure to C60 suspensions with concentration up to 10% of the original suspension,

water samples containing nano-silver and nano-copper were found to be toxic, as reflected by the

LC50 data shown in Fig 3-4. The highest toxicity was observed on organisms exposed to

nanopure-water suspended nanometals, with a LC50 of 0.462 μg/L for Ag (95% CI = 0.449 –

0.473 μg/L) and 2.14 μg/L for Cu (95% CI = 1.97 – 2.33 μg/L). With regard to Ag, SR1 had a

LC50 of 6.18 μg/L (95% CI = 5.52-6.66 μg/L), 0.771 μg/L (95% CI = 0.74546-0.798 μg/L) for

SR2, and 0.696 μg/L (95% CI = 6.56-0.730 μg/L) for SR3. The LC50 values for Cu in SR1 and

SR2 had trends similar to that of Ag and with LC50 of 46 μg/L for nano-Cu suspended in SR1

water (95% CI = 43-52 μg/L) and 7.12 μg/L (95% CI = 6.93-7.30 μg/L) in SR2 water. Unlike

Ag, nano-copper in SR3 was less toxic with a LC50 value of 48 μg/L (95% CI = 41-74 μg/L).

The toxicity of soluble Ag species has been well studied (Lee et al. 2005; Morgan et al.

2005; VanGenderen et al. 2003). Free Ag ions are typically very toxic to freshwater organisms

(Bury et al. 1999), but Ag complexed with dissolved organic compounds has much lower

toxicity (Hogstrand and Wood 1998; Rodgers et al. 1997). Therefore, this could explain the

lower toxicity of Ag-SR1 suspensions (i.e. higher LC50 to Daphnia). In SR3 samples, although

chloride will also form complexes with silver, AgCl0 colloids can still diffuse across biological

membranes, enhancing bioavailability and toxicity of Ag (Erickson et al. 1998; Glover and

Wood 2005; Hogstrand and Wood 1998). Interestingly, Ag-SR2 suspension was almost as toxic

as Ag-DI suspension; and this despite its much smaller Ag content. This could suggest that Ag+

was likely the dominant form after suspension of nano-silver SR2 water.

64

Cu can bind to the fish gills, inhibit ion transport, and cause disturbance of multiple

physiological processes (Grosell et al. 2007; Schwartz et al. 2004). Organic matter can not only

facilitate dissolution and dispersion of copper nanoparticles, but decrease the toxicity of copper

as well (Kim et al. 1999; Richards et al. 2001; Schwartz et al. 2004; Vasconcelos et al. 1997).

Similar to other results (Kramer et al. 2004; Perez et al. 2006; Ryan et al. 2004), our data also

demonstrated that toxicity to C. daphnia is linearly correlated to DOC (Fig 3-5). However, the

calculated slope is smaller than the above studies by approximately one or two orders of

magnitude, indicating much higher toxicity associated with nano-copper particles. For SR3,

increased water hardness and alkalinity could also reduce Cu toxicity owing to the competition

between Cu and other cations for binding sites or carbonate complexation (Erickson et al. 1996;

Hyne et al. 2005; Sciera et al. 2004).

In a separate study, we tested the toxicity of C60 water suspensions prepared by the

Deguchi’s method (Deguchi et al. 2001), which uses THF to transfer C60 to water (Kopelevich et

al. 2008). The LC50 for C. dubia exposed to such C60-suspensions was 0.395 mg/L (95% C. I. =

0.287-0.470 mg/L). Other researchers (e.g. Zhu et al. 2006) also compared the toxicity of THF-

produced C60 suspensions to that of C60 directly suspended in DI-water. They showed a much

poorer suspension in DI-water, resulting in lower C60 toxicity. In Zhu’s study, the LC50 for THF-

nC60 to Daphnia magna was about 0.8 ppm while water-stirred-nC60 didn’t cause 50% mortality

at concentrations up to 35 ppm (Zhu et al. 2006). It has been suggested that the residual THF is

responsible for the observed higher toxicity, but others argued that differences in surface

chemistry and/or morphology play a more important role in the observed toxicity responses

(Brant et al. 2005; Lyon et al. 2006). However, results from this study suggest that C60

nanoparticles could be less toxic if released directly to natural waters, although the potential exist

65

for long-term in-situ transformation of C60 to more reactive/toxic forms. Also, the effect on C.

dubia may not be representative of the response of most aquatic organisms.

3.3.4 Acute Toxicity with MetPLATE

Toxicity observed with MetPLATE would indicate the presence of bioavailable forms of

the metal (Bitton et al. 1994). No toxicity was detected with C60 samples, Ag-SR1, Ag-SR2, and

Cu-DI suspensions (Fig. 3-6). The EC50 values for SR3 and DI-water nano-Ag containing

suspensions were 112 ± 0.94 μg/L and 48 ± 6 μg/L, respectively.

Similar to the well known strong antimicrobial effects of silver (Ki et al. 2007; Sondi and

Salopek-Sondi 2004; Yoshida et al. 1999), recent investigations have also demonstrated the

antimicrobial effects of silver nanoparticles (Jain and Pradeep 2005; Lok et al. 2006; Yu 2007).

Besides the formation of Ag+ on the surface of nano-silver surface (Lok et al. 2007), the toxicity

mechanism is also related to the fact that silver nanoparticles can penetrate the cell membrane

and lead to bacteria death through reactions with the cell metabolic processes (Lok et al. 2006;

Sondi and Salopek-Sondi 2004). Our results showed that NOM coating in SR1 samples reduces

the availability of free Ag ions in the solution, while the toxicity of nano-Ag-SR3 and nano-Ag-

DI-water suspensions indicated presence of bioavailable Ag species (Bitton et al. 1994).

According to the literature, the bioavailable form of silver chloride could be AgCl0

(Hogstrand and Wood 1998). MetPLATE results for nano-Ag-DI-water suspension have

somewhat proved our speculations about Ag concentration in water, i. e., Ag+ was formed on the

surface by oxidation.

No toxicity was detected with Cu-DI, implying little formation of bioavailable copper

without organic matter and other ions. Therefore, the high toxicity of the same sample to C.

dubia is probably caused by Cu nanoparticles or nano-aggregates owing to the particle-feeding

66

behavior of the cladocerans. Contrary to silver, Cu-NOM complexes could still be bioavailable

to bacteria, and are thus more toxic (Apte et al. 2005). This is consistent with our Cu-SR1

MetPLATE toxicity result.

3.4 Conclusions

In conclusion, dispersion of nanoparticles in aqueous systems varies significantly in

different aqueous systems. The suspended concentrations of nanomaterials are mainly dependent

on the chemical properties of the particles and water characteristics. However, the aquatic

toxicity was not linearly correlated with the suspended content of MNs. As revealed by C. dubia

and MetPLATE tests, toxicity was not the result of bio-availabity solely; instead, it could be

largely affected by the metabolism and feeding behavior of the aquatic organisms. Our results

are of great significance in the context of understanding the fate, transformation and biological

effects of nanomaterials in the aqueous environments. Although some of the findings are

consistent with previous observed trends in manipulated nano-water systems, much still remains

to be answered and further investigated.

67

Figure 3-1. Map of the Suwannee River watershed and tributaries showing the three sampling locations. SR1-samples were collected near the Florida-Georgia border, SR2-samples were obtained from the river mid-section, and SR3 samples were collected along the salinity gradient in the river estuary.

68

Ag

0

0.5

1

1.5

2

2.5

3

SR1 SR2 SR3 DI

Con

cent

ratio

n (m

g/L

)

0.54

0.043

0.66

1.67

Cu

0

5

10

15

20

25

SR1 SR2 SR3 DI

Con

cent

ratio

n (m

g/L

)

12.58

1.45 0.51 0.134

C60

0

0.5

1

1.5

2

2.5

3

3.5

SR1 SR2 SR3 DI

Con

cent

ratio

n (m

g/L

)

1.62

3.09

0.038

0.83

Figure 3-2. Concentrations of (A) silver, (B) copper, and (C) C60 in different water samples spiked with individual nanomaterial and then filtered. Numbers above bars are average concentrations of two samples and vertical lines represent ± 1 standard deviation of the mean.

69

SR1

SR2DI

-2

0

2

4

6

8

10

12

14

0 10 20 30 40 50

DOC concentration (mg/L)

Cu

conc

entr

atio

n (m

g/L

)

Figure 3-3. Linear correlation of Cu concentrations and dissolved organic matter (DOC) in water samples SR1, SR2 and DI water (R2=0.9864).

70

Ag

0

1

2

3

4

5

6

7

SR1 SR2 SR3 DI

LC

50 (u

g/L

)6.18

(5.52-6.66)

0.77(0.745-0.798)

0.70(0.656-0.73)

0.462(0.449-0.473)

Cu

0

10

20

30

40

50

60

SR1 SR2 SR3 DI

LC

50 (u

g/L

)

46(43-52)

7.12(6.93-7.3)

48(41-74)

2(1.97-2.33)

Figure 3-4. 48-h LC50 values of (A) silver- and (B) copper-spiked water samples to Ceriodaphnia dubia. Numbers above the bars are the estimated LC50 values and the 95% confidence interval in parentheses determined by probit analysis.

71

SR1

SR2DI

-10

0

10

20

30

40

50

0 10 20 30 40 50

DOC concentration (mg/L)

LC

50 o

f Cu

(ug/

L)

Figure 3-5. Relationship between the 48-h LC50 values of nanocopper suspensions to

Ceriodaphnia dubia and dissolved organic matter (DOC) concentrations in SR1, SR2, and DI water.

72

Ag

0

20

40

60

80

100

120

SR1 SR2 SR3 DI

IC50

(ug/

L)

NA

112.14

NA

47.79

Cu

0

50

100

150

200

250

SR1 SR2 SR3 DI

IC50

(ug/

L)

NA

204

46.0830.91

Figure 3-6. IC50 values of (A) nanosilver- and (B) nanocopper suspensions using MetPLATE

test. Numbers above the bars are the estimated IC50 values and vertical lines represent the ± 1 standard deviation of the mean.

73

Table 3-1. Characteristics of water samples prior to contact with C60, Ag and Cu nanoparticles Dispersion

water Sample

type pH Alkalinity (mg/L as CaCO3)

Ionic Strength (I) (mM)

DOC (mg/L)

Ag (μg/L)

Cu (μg/L)

SR 1 Freshwater 4.70 6 0.94 45.71 <10 <10 SR 2 Freshwater 7.15 88 3.34 10.18 <10 <10 SR 3 Seawater 7.56 132 475.17 N/A* <10 <10

Note*: DOC concentration of SR3 is not available due to the high salinity. The reported DOC in this portion of the Suwannee river was about 2.3 mg C/L (Del Castillo et al. 2000).

74

CHAPTER 4 MOBILITY OF SINGLE-WALLED CARBON NANOTUBES (SWNTS) IN SATURATED

HETEROGENEOUS POROUS MEDIA

4.1 Introduction

Single-walled carbon nanotubes (SWNTs) have attracted the attention of both engineers

and scientists since their discovery in 1993 (Iijima and Ichihashi 1993). The unique size-related

characteristics of SWNTs, such as strength, elasticity, high adsorption capacity, and controllable

conductivity, have led to their use in a growing number of industrial processes as well as to their

introduction into a wide variety of commercial product (Isobe et al. 2006; Jia et al. 2005; Lyon et

al. 2005; Templeton et al. 2006). It has been estimated that 65 tons of nanotubes and fibers were

produced worldwide in 2004, resulting in revenue predictions that would exceed $4.5 billion by

2010, based on an annual growth rate of ~60% (Cientifica 2005). On the other hand, this

dramatic increase in production rates of SWNTs and their anticipated uses in a wide variety of

commercial and industrial applications suggest that they will inevitably enter the environment

and potentially impact the biosphere. Therefore, while the applications of SWNTs are exciting,

there is concern over potential environmental and human health problems.

Besides toxicity issues common to nanosized materials, the environmental and human

health implications of SWNTs seem to be complicated by the presence of impurities, such as

metal catalyst contaminants, which have been so far impossible to remove entirely without

destroying the sp2 structure of SWNTs. Additionally, the above listed potential applications of

SWNTs have motivated research on the production of highly dispersed suspensions

(Matarredona et al. 2003), and SWNTs are now routinely dispersed in a wide variety of

surfactants including sodium dodecyl sulfate (SDS) and sodium dodecylbenzene sulfonate

(NaDDBS), to preserve their inherent properties (Wang et al. 2004; Zhang et al. 2005). However,

75

some of the surfactants with efficient SWNT dispersion abilities are potentially toxic to living

organisms (see Chapter 2).

Overall, the above observations suggest that the toxic effects of SWNTs could be related to

several parameters including metal impurities (Lanone and Boczkowski 2006; Miyawaki et al.

2008; Shvedova et al. 2003), the degree and kind of aggregation in produced SWNTs (Tian et al.

2006; Wick et al. 2007), the nanosize and shape of SWNTs, the type of solvent and surfactant

used in the preparation processes, or the combination of two or more of the above parameters.

Currently, research on the implications of SWNTs has been devoted largely to toxicity

issues while the investigation of their fate and transport in both aquatic and terrestrial systems

remains limited. Changes in nanoparticle surface chemistry in combination with shifts in key

environmental parameters such as levels and type of organic matter, pH, and ionic strength could

affect the mobility of nanomaterials, and therefore, their environmental fate and transport

(Espinasse et al. 2007; Kanel et al. 2007; Lecoanet et al. 2004; Lecoanet and Wiesner 2004;

Schrick et al. 2004). For example, Schrick et al. (2004) found that negative surface charge could

prevent the aggregation of zerovalent iron nanoparticles by electrostatic repulsion of similarly

charged particles. Implicitly, the lack of aggregation would then favor the transport of nanosized

particles through porous media. Currently published research on transport of nanomaterials in

porous media are so far limited to qualitative descriptions of transport behavior (Heller et al.

2004; Lecoanet et al. 2004), and modeling studies focusing on transport of nanotubes in

heterogeneous porous media are still lacking.

In this study, we investigated the mobility of SWNTs suspended in either ionic (SDS) or

non-ionic (Gum Arabic) surfactants in natural sandy and clay soils using column experiments.

The ultimate objective of this study was to generate data that can be used to predict the ability of

76

SWNTs to move through heterogeneous porous media as a function of (i) soil characteristics and

(ii) the chemical composition of the solvent used to suspend SWNTs. For this work, transport

patterns of SWNT suspensions in soil columns were analyzed using a well-established

convection-dispersion equation (CDE), referred to herein as CDE-model. The CDE model was

selected because of its efficient ability to predict the transport of nanosize particles such as

viruses (20 – 200 nm) and colloids (<10μm) in soils (Chendorain et al. 1998; Close et al. 2006;

Jin et al. 2000; Vidales-Contreras et al. 2006).


4.2.1 Single-Walled Carbon Nanotube Sample Preparation

SWNTs were obtained from Rice University (Rice HPR 145.1). SWNT suspensions were

prepared by suspending an initial mass of about 40 mg of SWNTs into 200 ml of either an

aqueous Gum Arabic (GA) or Sodium Dodecyl Sulfate (SDS) surfactant solution (1% w/v). GA

and SDS used in this study were obtained from Aldrich-Sigma. To obtain highly dispersed

SWNT suspension, mixtures of SWNTs were first homogenized using a high-shear IKA T-25

Ultra-Turrax mixer for about 60 to 90 minutes followed by a 10-min ultra-sonication with a

Misonix S3000. Next, the mixture was centrifuged at 20,000 rpm (Beckman Coulter Optima L-

80 K) for 2.5 h. The supernatant was then carefully separated from the aggregated nanotubes

collected at the bottom of the tube. In parallel, an aqueous suspension of SWNT without

surfactant was prepared by direct mixing of SWNT and Nanopure® water followed by a 15-

minute sonication period. For all of these produced suspensions, visible absorbance spectra of

SWNTs were recorded by using a Nano-Fluorescence Nanospectrolyzer (Applied

NanoFluorescence, LLC), and the concentrations of SWNTs suspensions determined by

absorbance at 763 nm.

77

4.2.2 Soil Sample Collection and Characterization

Two types of soils were collected from the McCarty Woods on the University of Florida

campus (sandy soil) and from a construction site near Atlanta, Georgia (Georgia clay soil). After

removing the top 2 to 3 cm, soil samples were collected from the top 1 to 2 feet. The

physicochemical characteristics of these soils have been reported previously (Feng et al. 2007),

and are summarized here in Table 4-1. The soil pH was measured according to the U.S. EPA

method 9045D (USEPA 2004). Particle size distribution was determined according to the USDA

Soil Survey Lab Method (USDA 1992). The Walkley & Black Method (Walkley and Black

1934) was used to measure the soil organic carbon content. Soil organic matter was calculated by

multiplying the soil organic carbon content by a coefficient of 1.72 (Nelson and Sommers 1996).

Prior to use in column experiments, collected soils were first air-dried, sieved (1.19mm) to

remove gross debris and then homogenized.

4.2.3 Column Experiments

Columns used in this study were 30 cm long clear plexiglass tubes with ~4 cm internal

diameter (see Figure 4-1). The bottom of the columns were sealed with glass wool and secured

with a non transparent PVC ring equipped with an opening to allow the leachate through.

For sandy soil experiments, each column was filled with 300 g of soil from the top. The pore

volume of each of packed column was measured by gravimetry and averaged 93.5 ml. The

porosity was determined to be 0.37. Prior to the introduction of a conservative tracer solution,

bromide (Br-) in this case, and SWNTs suspension into the columns, packed soils were first

saturated with nanopure® water (>18 MΩ cm, Barnstead Corp) to enhance packing

homogeneity and help clean the effluent from loose soil fine particles. Next, 50 ml of Br-

solution, and 50 ml of SWNT-GA, SWNT-SDS or sonicated SWNTs suspension were added to

each column from the top, followed by continuous addition of DI water. Effluent samples were

78

collected at discrete intervals and analyzed for Br- and SWNT concentrations. In case of clay

soil, a similar procedure was used, except that 150 g of soil were packed into the column and 25

ml of Br- solution, and 25 ml of SWNT-GA or SWNT-SDS suspensions were used. The pore

volume for the clay soil column was determined to be 76.625 ml and a calculated porosity of

0.41. On collected leachates, the concentrations of SWNTs were determined as described earlier

using the Applied NanoFluorescence Nanospectrolyzer. Bromide concentrations were measured

using Ion Chromatography (Dionex Model DX-320 IC system). All experiments were conducted

in duplicates.

4.2.4 Modeling

As stated earlier, a CDE model of solute transport in soils was selected and used in this

study to simulate SWNT transport through the soil columns. The CDE model can simulate

quantitatively the retention and transport of solutes in soils and reveal the effect of soil

characteristics on solute fate and transport (Chen et al. 2006). The one-dimensional governing

equation for this model can be written as follows (Parker and Vangenuchten 1984):

cxcv

xcD

tcR μ−

∂∂

−∂∂

=∂∂

2

2

(4-1)

where c is the solute concentration in mg/L, t is the time (min), R is the retardation factor, D is

the hydrodynamic dispersion coefficient (cm2 min-1), x is the distance (cm), v is the pore-water

velocity (cm min-1), and μ the removal rate coefficient (min-1). For linear instantaneous sorption,

the above retardation factor R is defined as:

θ

ρ dbKR +=1 (4-2)

79

where ρb is the bulk density of sandy soils (dry weight) (g cm-3), θ is the volumetric water

content (cm3 cm-3), and Kd is the equilibrium sorption distribution coefficient between the liquid

and the sorbed phases (cm3 g-1). The value of R is evidently equal to or greater than 1.

The nonlinear least-square curve-fitting program CXTFIT 2.1, was used to estimate

transport parameters with experimentally collected breakthrough curve (BTC) data (Toride et al.

1999). The bromide data were analyzed by setting the removal rate (μ) as zero and the

retardation factor (R) as 1.


4.3.1 Bromide Transport and Breakthrough Curves in Sandy and Clay soils

Breakthrough curves (BTCs) observed and simulated with the CDE model for the

conservative tracer (i.e. Br-) through sandy soil and clay soil columns are presented in Figure 4-

2. Table 4-2 lists the parameters used in the model simulations. As a conservative tracer, Br- is

assumed to have no retardation (R = 1) or reaction (μ = 0) within the soil columns, and based on

experimentally measured pore water velocities, v, and the dispersion coefficients, D, determined

by the model as a fitting parameter, the output of model simulations matched the experimental

data of Br- transport very well (with correlation coefficients (r2) equal to 0.98 and 0.85 for sandy

soils and clay soils, respectively – see Figure 4-2).

4.3.2 SWNTs Transport in Sandy Soils

Figure 4-3 shows the BTC of experimental data and simulated data of SWNT suspended in

GA and SDS through sandy soil columns. For CDE model, the pore water velocity was measured

experimentally. Dispersion on the other hand is mainly determined by the characteristics of the

porous media. It is therefore acceptable to assume the dispersion coefficient to be constant for

different solutes traveling through the same packed soil. Accordingly, the dispersion coefficient

80

of SWNT was assumed to be similar to that of Br-, and the reaction rate μ and retardation

coefficient R were then determined by model simulations (Table 4-2).

Although studies focusing on fate and transport of most nanomaterials in heterogeneous

porous media are either lacking or preliminary, extensive investigations have been carried out on

transport and retention of virus particles in soil media (Bales et al. 1991; Bales et al. 1993; Bitton

et al. 1979; Bitton et al. 1984; Funderburg et al. 1981; Han et al. 2006; Jin and Yates 2002; Jin et

al. 1997). Given the nanosize of these biotic particles, results on viruses can be used for

comparison purposes. For instance, one of the major factors controlling the transport of viruses

in soil columns is adsorption (Bitton et al. 1979). Under saturated soil conditions, viruses may

reach equilibrium and/or undergo kinetic adsorption on soil particles (Schijven and Hassanizadeh

2000). The CDE model, which has been successfully used to predict the transport behavior of

viruses and colloids in soils, seems to work quite well in simulating SWNTs transport in porous

media, with correlation coefficients (r2) of 0.96 and 0.86 for SWNT-GA and SWNT-SDS,

respectively. However, observed SWNT BTCs tend to have longer tails compared to Br- BTC

(Figure 4-3), indicating the likely occurrence of kinetic adsorption of SWNT on tested sandy

soils (Schijven and Hassanizadeh 2000). R-values close to 1 were observed in both SWNTs-GA

and SWNTs-SDS leaching experiments, reflecting minimal equilibrium adsorption and

retardation of SWNTs by sandy soils at tested flow rates. The quality of the model fitting is

obviously affected by the number of data points as SWNT-GA yielded twice more data points

than SWNT-SDS, resulting in a better fit.

While SWNTs were detected in effluents collected from columns spiked with SWNT-GA

and SWNT-SDS suspensions, no SWNT was detected in the first 2 L effluent collected from

sonicated aqueous SWNT suspensions. These results indicate that surfactants do facilitate the

81

transport of SWNTs in the soils. This is mainly due to ability of surfactants to reduce the

interfacial and surface tension of particles and increase the capacity of the mobile aqueous phase

by forming micelles (Chen et al. 2006). The mass balance was 37% and 47% for SWNT-GA and

SWNT-SDS, respectively. The mass recovery results suggested that more SWNT-GA was

retained in the soil column than SWNT-SDS. It is probably because SDS can provide negative

charges on the surface of SWNT micelles and cause more electrostatic repulsion and less

attachment efficiency. Furthermore, the greater values of R and μ of SWNT-GA than SWNT-

SDS indicate stronger adsorption of SWNT-GA to the sandy soils as well.

In previous investigations involving other carbonaceous nanomaterials, Espinasse et al.

(2007) compared fullerene (C60) transport in the presence of tannic acid and polysaccharides and

suggested that the negatively charged tannic acid resulted in less attachment efficiency than

polysaccharide-like organic compounds. Several studies on transport of virus particles in soil

columns (Chattopadhyay et al. 2002; Cheng et al. 2007; Pieper et al. 1997; Wall et al. 2008;

Zhuang and Jin 2003) also observed increased transport using surfactants or natural organic

matters. Zhuang and Jin (2003) proposed that organic matter can provide negative charges or

cover positively charged sites, thus increasing the repulsion between virus and soil particles.

Wall et al. (2008) suggested it is due to organic matter occupying adsorption sites or physically

blocking pores. From these findings it can be concluded that surfactant modified surface charges

of the nanoparticles could influence the retention and transport in the sandy soils.

4.3.3 SWNTs Transport in Clay Soils

Regardless of the composition of the SWNT suspension used (i.e. GA, SDS), SWNTs did

not leach out of the clay soil columns after more than 10 pore volumes. This reveals that all

SWNTs were strongly retained by clay particles, therefore delaying their downward motion. To

some extent, this is not surprising because previous studies on viruses noted that the sorption

82

capacity virus particles increased with increasing clay fractions in soils (Bitton 1975; Bitton et al.

1978; Chu et al. 2003; Juhna et al. 2003). In addition, impeded flow rates were observed as

compared to sandy soils. Also, this effect was particularly pronounced in clay soil columns

spiked with SWNT-SDS. However, as shown in sandy soil experiments, the flow rate of SWNT-

SDS treated columns were higher than those in SWNT-GA treated columns. This opposite trend

is probably due to pore size exclusion of the SWNT-GA nanoparticles (Schijven and

Hassanizadeh 2000).

Overall, it is obvious that the retention of SWNT in soil varies with soil characteristics, and

soils with high clay content could significantly adsorb/retain SWNTs, or at least significantly

delay their downward transport.

4.4 Conclusions

The observed trends of SWNTs mobility agree well with some other investigations on

nanoparticles and nano-sized microorganisms such as viruses. The surface properties of

nanomaterials are important in determining the transport of nanomaterials through porous media.

With the help of dispersing surfactant, SWNTs moved rapidly in a typical Florida sandy soil,

raising concerns that surfactant-dispersed SWNTs could rather easily travel through such soils

and reach the groundwater in case of accidental spill for example. In contrats, finer clay soil

particles could significantly impede the mobility of SWNT. In this latter case, however, the

reported toxicity effects of SWNT on microorganisms could raise concerns on the long-term

impact of such retained SWNT on soil microorganisms and associated ecosystem services.

Finally, this study also demonstrated that the CDE model could be used to predict the transport

behavior of SWNT suspension in sandy soils. This initial investigation opens the door for future

studies on transport of nanomaterials in porous media, and ideally, such studies should

investigate changes in physicochemical characteristics of MNs following transport through soil

83

columns, as well as the potential toxicity of eluted MNs and/or their derivatives. It appears that

under specific conditions, the transport behavior of SWNT can be modeled, and this effort could

be extended to other types of MNs to examine the physicochemical properties governing the fate

and transport of MNs in porous media.

84

Figure 4-1. Schematic diagram of the experimental setup for SWNTs transport in packed heterogeneous sandy or clay soils.

NM suspension and leaching

solution

Tested soil

PVC column

PVC ring

Collection container

85

Bromide

-0.1

0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

0 0.5 1 1.5 2 2.5 3

Pore volume

C/C

0

Experiment dataSimulated data

Bromide

-0.1

0

0.1

0.2

0.3

0.4

0.5

0.6

0 0.5 1 1.5 2 2.5 3

Pore volume

C/C

0

Experimental dataSimulated data

Figure 4-2. Breakthrough curves of experimental and simulated data of bromide (Br-) in sandy (A) and clay (B) soils. Each point represents the average of two replicates and vertical bars are standard deviations.

(A)

(B)

86

SWNT in GA

-0.05

0

0.05

0.1

0.15

0.2

0.25

0.3

0 1 2 3 4 5

Pore volume

C/C

0

Experiment data

Simulated data

SWNT in SDS

-0.1

0

0.1

0.2

0.3

0.4

0.5

0 0.5 1 1.5 2 2.5 3

Pore volume

C/C

0

Experiment dataSimulated data

Figure 4-3. Breakthrough curves of experimental and simulated data of SWNT-GA and SWNT-SDS suspensions in sandy soil columns. Vertical bars are standard deviations of two replicates.

87

Table 4-1. Physicochemical characteristics of the sandy (Gainesville, Florida) and clayey (Atlanta, Georgia) soils used in column experiments (Feng et al. 2007).

Characteristic Sandy soil clay soil pH 5.7 5.7

% Organic carbon 0.5 0.6 % Organic matter 1.6 1.8

% Sand 96.92 56.4 % Silt 0.02 22.6

% Clay 3.06 21.0

88

Table 4-2. Transport parameters estimated by CXTFIT for bromide and SWNT in GA and SDS in sandy soils.

r2 v (cm min-1) R D (cm2 min-1) μ (min-1)

Bromide in sandy soils 0.98 7.77 1 13.03 0

SWNT-GA in sandy soils 0.96 6.34 1.28 - 0.34

SWNT-SDS in sandy soils 0.86 7.28 1.06 - 0.22

89

CHAPTER 5 POTENTIAL IMPACTS OF MANUFACTURED NANOMATERIALS ON

BIOGEOCHEMICAL PROCESSES IN SEDIMENTS

5.1 Introduction

Ecosystems accomplish numerous natural services and most, if not all of them, seem to

have common main characteristics including the flow of energy, material and information, and

the participation of biota and water. Therefore, the ability to qualitatively and/or quantitatively

characterize any of the above listed natural processes can be used to assess the impact of

pollutants on ecosystem functions, and could be provided by thermodynamics, which has been

successfully applied in the description of the basic properties of ecosystems (e.g. flow of matter

and energy).

In sediments, the composition and distribution of microbial populations are usually well-

established, although changes associated with shifts in seasons and other major parameters are

also common. However, the input of pollutants can have significant impacts on the composition

of microbial communities and/or their activities in sediments. In such cases, potential

consequences could range from a simple delay in biodegradation of organic matter to major

environmental impacts such as the production of more toxic derivatives, with bioaccumulation

potential and negative effects on ecosystem functions.

By combining the flow of material and energy to microbial activity, a series of reactions

involved cycling of organic carbon in sediments can be used as proxy to detect potential impact

of manufactured nanomaterials (MNs) on basic ecosystem functions. This is because the

anticipated widespread production and use of MNs could lead to new environmental pollutants

(Masciangioli and Zhang 2003), which could either directly impact living cells, or undergo

environmental transformations to produce secondary toxic derivatives.

90

In contrast to overwhelming literature on MNs production and application (e.g.,

Borderieux et al. 2004; Davis 1997; Eng 2004; Florence et al. 1995; Jensen et al. 1996; Li et al.

2006; Pitoniak et al. 2003; Tungittiplakorn et al. 2004; Wang and Zhang 1997; Zajtchuk 1999

Kotelnikova et al., 2003), and growing focus on toxicity response of test model organisms and

human cells to MNs exposure, only limited number of studies have investigated the fate,

transport and the resulting impacts of MNs at the system level (Hyung et al. 2007; Lecoanet et al.

2004; Lecoanet and Wiesner 2004). They showed that MNs could exhibit different transport

behaviors in porous media or become stabilized in organic rich waters with potential for long-

range transport. Relevant to this study are findings that C60 could be toxic to soil microbial

species (Fortner et al. 2005), depending on the composition of culture medium and C60

speciation. Also, CdSe quantum dots are believed to be cytotoxic due to the release of Cd2+ ions

(Derfus et al. 2004; Lanone and Boczkowski 2006; Zhang et al. 2006) and/or through direct

interaction of the quantum dots with cells (Kirchner et al. 2005; Liang et al. 2007). Likewise, the

reported antimicrobial effects of silver nanoparticles can become a “double-edged sword” owing

to the extremely high toxicity of Ag+ that forms on Ag-nanoparticle surfaces (Lok et al. 2007;

Lok et al. 2006; Yu 2007). Preliminary studies conducted in our laboratory based on kinetics of

acetate biodegradation showed that the addition of C60 to sediment slurries negatively impacted

rates of acetate oxidation (Kopelevich et al. 2008). In contrast, Tong et al. (2007) and Nyberg et

al. (2008) did not detect negative responses by soil microbial communities by monitoring gas

production (i.e. CO2 and CH4) in C60 spiked soils.

Overall, the above observations tend to point to the possibility of certain MNs to act as

bactericides, and therefore, to negatively impact soil and sediment microorganisms. Therefore, in

this study, the effects of aqueous suspensions of C60, CdSe-quantum dots and nano-silver on the

91

sediment microorganisms were investigated. Sediment slurries were manipulated to favor certain

terminal electron acceptors (nitrate and sulfate) and the effect of tested MNs assessed.


5.2.1 Preparation of Nanomaterial Suspensions

The preparation method for nC60 suspensions was adapted from Deguchi et al. (2001).

Briefly, ~15 mg of C60 (99.5%, Term-USA) were added to 500 ml of THF, bubbled with ultra-

high purity (UHP) nitrogen for 1 hour to remove oxygen, and then sealed and left stirring at

room temperature for 24 hours. Excess solids were later filtered out using a 0.45 μm PTFE

membrane filter, resulting in a transparent pink solution. The filtrate was then added to equal

amount of water and placed in a water bath prior to purging with UHP-N2 to evaoparte THF.

Following the THF evaporation, the obtained solution was vacuum-filtered through 0.45 μm

cellulose into a flask and stored in the dark. A suspension of nano-silver (Quantum Sphere, Inc.)

at 200 mg/L was prepared in nanopure water by simple shaking at room temperature for 48

hours. CdSe quantum dot suspensions were purchased from NN-Labs, LLC (Fayetteville, AR).

5.2.2 Sediment Collection

Sediments used in this study were collected from a polluted lake (Lake Alice) located on

University of Florida campus in Gainesville, Florida. Surface sediments (top 10 cm) were

collected using pre-cleaned high density polyethylene scoops, transferred into sieves (2 mm), and

the sieved fraction was stored in pre-cleaned plastic containers.

5.2.3 Dominant Terminal Electron Accepting Processes (TEAPs) in Sediments and Sediment Manipulation in this Study

Understanding the dynamics of redox processes is key to predicting the fate and impact of

certain environmental contaminants (Davis et al. 1999; Kampbell et al. 1996; Van Stempvoort et

al. 2002). However, methods for evaluating redox processes are quite problematic. Platinum

92

electrode measurements of redox potential (Eh) are commonly used, and this in spite of the fact

that unique redox potentials in natural systems do not exist (Thorstenson 1984)—and that

measured Eh values do not agree with Eh values calculated from measured concentrations of

redox couples (Lindberg and Runnells 1984). This is likely due to the fact that most natural

systems are seldom in a state of redox equilibrium, and that platinum electrodes are subject to a

variety of interferences.

The use of molecular hydrogen (H2) as indicator of predominant TEAPs provides an

alternative method for evaluating redox processes in natural systems (Chapelle et al. 1996;

Lendvay and Adriaens 1999; Lovley et al. 1994; Lovley and Goodwin 1988; McGuire et al.

2000). In fact, fermentative microorganisms continuously produce H2 during anoxic

decomposition of organic matter, and the produced H2 is consumed by respiratory

microorganisms that use Fe(III), sulfate, or CO2 as terminal electron acceptors (Chapelle et al.

1996). Previous research (Lovley et al. 1994; Lovley and Goodwin 1988) has shown that in

sediment, H2 concentrations associated with specific TEAPs fall in quite well defined range of

values, which can be divided as follows: methanogenesis: from ~7 to >10 nM; sulfate reduction:

1 to 1.5 nM; Fe (III)-reduction: ~0.2 nM; Mn (IV) and/or nitrate-reduction: <0.05 nM.

However, while H2 concentrations are better indicator of redox potential than Eh

measurements, it is worth noting that H2 concentrations do have drawbacks as well. In certain

specific cases, H2 measurements have indicated the predominance of sulfate reduction in samples

that lacked significant sulfate concentrations (Chapelle et al. 1996). Also, its occurrence at trace

levels requires highly sensitive and expensive analytical techniques. Accordingly, for studies

conducted in batch reactors, one of the reliable and inexpensive approaches used electron-

93

acceptor availability and the presence of final products of microbial metabolism as illustrated in

half-reactions (equations 5-1 through 5-8).

OHeHaqO 22 21)(

41

=++ −+ (5-1)

OHNOeHNO 223 21

21

21

+=++ −−+− (5-2)

OHgONeHNO 223 85)(

81

45

41

+=++ −+− (5-3)

OHgNeHNO 223 53)(

101

56

51

+=++ −+− (5-4)

OHMneHSMnO 22

2 2212)(

21

+=++ +−+ (5-5)

)aq()s( FeeFe +−+ =+ 23 (5-6)

OHHSeHSO 224 2

181

89

81

+=++ −−+− (5-7)

OHgCHeHgCO 242 41)(

81)(

81

+=++ −+ (5-8)

The occurrence of the above geochemical reactions is thermodynamically driven, as

reactions with the lowest Gibbs free energy are favored when several TEAs are present in the

system. Accordingly, simple measurements of certain reactants ( −3NO , +3Fe , and −2

4SO ) and

products ( −2NO , ON2 , +2Fe , and 4CH ) of redox reactions associated with the degradation of a

tracer organic compounds can help identifiy the potential impacts, if any, of MNs on rates of

these microbial-catalyzed geochemical reactions.

This study focused primarily on sediment microorganisms that utilize nitrate (equations

5-2, 5-3, and 5-4) or sulfate (equation 5-7) as TEAs. In the laboratory, and prior to MNs testing,

94

relatively diluted sediment slurries (1:5 m/v) were prepared under continuous stirring in 3 L glass

containers, purged with N2, and hermetically sealed, The TEAs (e.g. nitrate, Mn and Fe

oxyhydroxides, and sulfate) naturally present in sediments were allowed to be consumed over

time, and after about three weeks, IC analyses of filtrate aliquots from slurries showed that

nitrate was mostly converted to nitrite and sulfate concentrations averaged ~24 mg/L. At this

point, aliquots of well-homogenized sediment slurries were transferred into 50 ml serum vials

within an anaerobic chamber and then spiked with a small volume of a highly concentrated

solution of sodium acetate (CH3COONa) to obtain a final concentration of ~150 mg CH3COO-

per liter of slurry. In parallel, sediment slurries with final concentrations of 150 ppm of

CH3COO- and 100 ppm of either −3NO or −2

4SO were also prepared to re-establish the prevalence

of these specific TEAs and favor biogeochemical reactions catalyzed by either nitrate- or sulfate-

reducing bacteria. Hermetically sealed vials were placed in the dark and the microbial

degradation of acetate monitored over time in both MNs non-spiked (i.e. controls) and spiked

slurries. Tested MNs included aqueous suspensions of C60, CdSe quantum dots, and nano-Ag at

final concentrations of 0.14 ppm, 0.5 ppm and 0.2 ppm, respectively. These tested concentrations

were based on the IC50 values of these MNs determined previously in toxicity tests using the P.

subcapitata 96-h chronic toxicity assay. Disposable syringes were used to withdraw samples

from the vials, followed by filtration through 0.45 μm syringe-filters in the anaerobic chamber.

The filtrates were then analyzed for −COOCH3 , −3NO , −

2NO and −24SO by ion chromatography

(see analytical techniques below). Obtained results were fit to a pseudo-first order kinetic model

to help assess the impact of MNs on rates of acetate oxidation by indigenous sediment’s

microorganisms. All experiments were run in triplicates.

95

5.2.4 Analytical Techniques

The concentrations of −COOCH3 , −3NO , −

2NO and −24SO were determined by ion

chromatography (Metrohm 700-series IC system). Concentration of obtained aqueous fullerene

suspension (nC60) was measured by a method adapted from Deguchi et al. (2001). Briefly, the

fullerene suspension, a 2% NaCl solution and toluene were mixed in a 1:1:2 ratio and then

sonicated for 10 minutes. After separation of the aqueous and organic phases, the upper toluene

layer was withdrawn for absorbance measurement at 334 nm using a Hach DR/4000U

Spectrophotometer.

5.2.5 Data Analysis

All experiments were run in triplicates and obtained data presented as mean ± 1 standard

deviation. Differences amongst the treatment effects were analyzed using one-way analysis of

variance (ANOVA) followed by Tukey’s test for multiple comparisons of anion concentrations.

Differences were considered significant when p values were < 0.05.


In these experiments, nitrate and sulfate were added to sediment slurries to establish

predominant TEAPs, and then assess the impacts of tested MNs on bacteria that use nitrate or

sulfate as TEA during organic matter oxidation.

Figure 5-1 shows trends of acetate degradation (5-1a), nitrate (5-1b), nitrite (5-1c), and

sulfate (5-1d) in sediment slurries spiked with acetate, tested MNs, without either nitrate or

sulfate. Accordingly, nitrate (~1 mg/L), nitrite (~5.7 mg/L), and sulfate (~24 mg/L)

concentrations at time zero here are levels naturally occurring in the sediment slurries prior to

sample incubations. Figures 5-2 and 5-3 are similar to Figure 5-1, except that the sediment

slurries were spiked with acetate and MNs, and with nitrate (Figure 5-2) or sulfate (Figure 5-3).

96

The acetate added to control and CdSe or nano-Ag spiked sediment slurries decreased in a

similar manner (Fig. 5-1a), resulting in comparable reaction rates of ~0.1 day-1 (Fig. 5-4a). In

contrast, acetate concentration in C60-treated slurries did not decrease over time, suggesting an

inhibition of microbiological processes leading to the oxidation of organic matter. During the 17-

days of incubation, trends in nitrate (Fig. 5-1b) and nitrite (Fig. 5-1c) in these sediment slurries

paralleled that of acetate, suggesting that both nitrate and nitrite were used as TEAs during

acetate oxidation, except in C60-treated slurries. Also, sample redox levels were not low enough

to favor sulfate reduction, as the concentrations of sulfate remained constant throughout the

experiment (Fig. 5-1d).

Figure 5-2 shows the importance of nitrate addition to sediment slurries and the response

of sediment microorganisms assessed through their ability to oxidize acetate. In these sediment

slurries, acetate degradation rates were much faster (Figs. 5-2a and 5-4b). But overall, observed

trends remain similar to those described earlier (Fig. 5-1), in that C60-treated samples resulted in

inhibition of acetate degradation, while CdSe- and nano-Ag spiked samples paralleled trends

observed in control samples. However, some trends were observed for both nitrate (Fig. 5-2b)

and nitrite (Fig. 5-2c). Soon after 5 days of incubation, divergent trends were observed and

generally decreased for nitrate, except in C60-treated sediment slurries. With regard to nitrite, and

besides the increasing trend observed in CdSe-treated slurries (i.e. likely conversion of nitrate to

nitrite), the other treatments resulted in either flat nitrite trend (i.e. no removal or addition in C60-

treated samples) or decreasing concentrations (Fig. 5-2c). These results suggest that different

pathways were used by microorganisms present in sediment slurries to oxidize acetate. In CdSe

treated slurries, nitrate reducers tended to produce nitrite (i.e. −3NO −

2NO ; Dou et al. 2008),

while the decrease in both nitrate and nitrite in sediment slurries spiked with nano-Ag and

97

acetate alone seemed to involve bacteria that convert these TEAs to either N2 or N2O (Radehaus

1997). Finally, sulfate concentrations did not undergo changes, as nitrate remained abundantly

available in the system after acetate was fully consumed.

Figure 5-3 shows the response of microorganisms in sediment slurries spiked with excess

sulfate to MNs addition. The lack of visible change in sulfate concentrations (Fig. 5-3d) shows

that sulfate reducing bacteria (SRB) were not involved in any of the acetate degradation observed

in these experiments (Fig. 5-3a). To efficiently test the impacts on naturally occurring sediment

SRB, it would have been necessary to deplete the tested sediment slurries of their original

background sulfate through a long-term anaerobic incubation in order to expect a significant

microbiological response attributable primarily to SRB. Figures 5-3b and 5-3c show that the

acetate degradation observed here is still tied to the coupled biogeochemical cycling of nitrogen

and organic carbon.

Overall, these results show that the disappearance of acetate from aqueous phase follows a

pseudo-first order kinetic in both MNs-treated and non-treated sediment slurries with the

exception of C60 treated samples. In the latter, concentrations of acetate remained almost

constant up to 17 days, implying a severe microbial inhibition by the C60-suspensions. In our

previous studies, we observed only impeded biodegradation of acetate even with a higher nC60

concentration of 0.5 ppm (Kopelevich et al. 2008). This difference is probably due to sediment

slurry preparation methods, i.e., use of highly diluted (in this study) versus thick (in study by

Kopelevich et al. 2008) slurries. The apparent rate of reaction (kapp, determined as the slope of

the regression of ln[Cacetate] versus time) is comparatively smaller in CdSe and nanosilver spiked

slurries than the control for acetate and nitrate spiked samples. The determined apparent reaction

rate (kapp) of acetate degradation was 0.44 day-1 for non-MNs-treated (i.e. controls) sediment

98

slurries and about two times lower for nanosilver (kapp= 0.24 day-1) and CdSe (kapp= 0.20 day-1)

treated sediment slurries

Although molecular probes such as the phosphor-lipid fatty acid (PLFA) profile would

have been helpful in identifying the dominant groups of microorganisms in sediments under

these specific experimental conditions, these results point to the potential of C60 to inhibit the

microbiologically driven oxidation of organic matter. This could be due to their bactericidal

effect (Fortner et al. 2005; Lyon et al. 2006).

5.4 Conclusions

In summary, the effects of MNs on sedimentary biogeochemical processes were studied.

Fullerene aqueous suspensions at a concentration that results in 50% inhibition growth of P.

subcapitata (i.e. 0.14 ppm) stopped the degradation of acetate by sediment microorganisms. In

contrast, nanosilver and CdSe treated sediment slurries slowed down rates of acetate oxidation

rather slightly in the presence of high nitrate concentrations. This could be an indication that

higher concentrations of these MNs would likely result in negative effects on sediment

microorganisms as well. Future research should combine the use of molecular probe and well-

defined redox conditions to pinpoint specific microbial groups and assess their responses when

they become exposed to increasing MNs concentrations.

99

Figure 5-1. Kinetics of acetate degradation (a), and nitrate (b), nitrite (c) and sulfate (d)

concentrations in sediment slurries without (controls) or spiked with tested nanomaterials (C60, nanosilver, and CdSe quantum dots). Vertical bars represent ± 1 standard deviation of the mean of three replicates.

100


concentrations in sediment slurries spiked with excess nitrate and without (controls) or with tested nanomaterial additions (C60, nanosilver, and CdSe quantum dots). Vertical bars represent ± 1 standard deviation of the mean of three replicates.

101


concentrations in sediment slurries spiked with excess sulfate and without (controls) or with tested nanomaterial additions (C60, nanosilver, and CdSe quantum dots). Vertical bars represent ± 1 standard deviation of the mean of three replicates.

102

Figure 5-4. Pseudo-first order kinetics of acetate disappearance from sediment-slurries treated with either silver nanoparticles or CdSe quantum dots as compared to the non-treated controls. Ln(C) represents the natural log of acetate concentration in mg/L. (a) = slurries containing acetate only, (b) = slurries with both acetate and nitrate additions, and (c) = slurries with acetate and sulfate additions.

103

CHAPTER 6 NANOWASTES IN THE ENVIRONMENT: THE “TROJAN HORSE EFFECT” OF

NANOMATERIALS

6.1 Introduction

The environmental fate and transport of manufactured nanomaterials (MNs) as well as the

fate of pollutants sorbed onto MNs through nanotechnology-based remediation processes are of

growing concern. This is because nanoscience and nanotechnology are now poised to become the

most important drivers of economic growth and development for the early 21st century. Most

scientists and engineers are confident that nanoscience and nanotechnology will revolutionize

medicinal, industrial, agricultural, and environmental research as a wide variety of MNs are

being produced (Hurt et al. 2006). Although in its infancy, research on both the environmental

impacts and health implications of MNs is fast growing (Biswas and Wu 2005; Goodman et al.

2004; Oberdorster 2004; Oberdorster et al. 2006; Sayes et al. 2004; Xu et al. 2004). In contrast,

the fate and potential impacts of pollutants adsorbed onto MNs through nanotechnology-based

remediation processes have been simply ignored. For instance, the use of MNs in the removal of

pollutants from either aqueous and/or gaseous effluents will generate nanowastes that need to be

either recycled or disposed of safely. So far, a reactive approach has been the most common way

of dealing with emerging pollutants (Daughton 2004). Unfortunately， a major disadvantage of

this late corrective approach is the difficulty to deal with well-established economic activities

that generate the pollutants of concern. Therefore, a proactive approach is ideal to limit the

complex ramifications associated with delayed prevention and remediation measures (Daughton

2004).

We assessed the fate and potential impacts of Hg sorbed onto MNs (i.e. SiO2-TiO2

nanocomposites) by mimicking the “Trojan horse” effect in a sediment matrix. Spent SiO2-TiO2

104

nanocomposites used in the removal of Hg from a simulated coal combustion effluent were used

in laboratory studies to determine the bioavailability of inorganic Hg sorbed onto SiO2-TiO2

nanocomposites by using Hg methylation as a proxy for bioavailability, and the toxicity of Hg-

SiO2-TiO2 complexes using FluoroMetPLATE®, a bacterial based microbiotest.


A thorough description of the procedure used to prepare the SiO2-TiO2 nanocomposites as

well as the mechanisms of Hg sorption onto MN surfaces have been reported previously

(Pitoniak et al. 2005). For measurement of Hg bioavailability using microbial catalyzed

methylation as surrogate for bioaccessibilty and to minimize the Hg methylation signal

associated with the sediment native Hg, we intentionally used pristine sediments with a

background total Hg (THg) concentration of only 5.68 ± 0.44 ng g-1 wet weight (Odum Wetland,

Gainesville, FL, USA). This THg value falls on the low end of the reported global background

range, for which common values are mostly between 200 and 400 ng Hg/g. Sediments were first

sieved (<2 mm) to produce a homogeneous fine material, and used later in laboratory

experiments as source of Hg methylating bacteria. Hg methylation experiments were conducted

using sediment slurries prepared with Nanopure® water in a 1:5 ratio (mass/volume). The initial

pH of the slurries was 4, and a 0.1N NaOH solution was used to produce slurries with pH>4. All

prepared sediment slurries were then de-aerated with ultra high purity (UHP) N2 to help

accelerate the development of anoxic conditions and favor methyl-Hg production. Overall, the

experiment consisted of control slurries with no Hg addition and slurries spiked with Hg as Hg-

SiO2-TiO2 complexes. In this latter treatment, Hg was added to increase the background amount

of THg naturally present in sediments by about 40%. Tubes were then sacrificed at different time

periods and analyzed for produced methyl-Hg. In the present study, total- and methyl-Hg

concentrations in aqueous phase and sediments were determined following previously published

105

methods (Bloom 1989; Bloom and Crecelius 1983; Bonzongo and Lyons 2004; Warner et al.

2003).

To assess the toxicity of SiO2-TiO2-nanocomposites used in Hg removal from gaseous

effluents, both virgin and spent SiO2-TiO2-nanocomposites were leached separately using the

Synthetic Precipitation Leaching Procedure (SPLP) solution, which is a mixture of HNO3 and

H2SO4 with a final pH of 4.22 ± 0.05 (USEPA 1996). Based on preliminary determinations of

percent inhibition, a 1:60 ratio (ml/mg) was used and for each replicate, about 0.18g of either

single nano-oxides (i.e. SiO2 and TiO2), virgin SiO2-TiO2 nanocomposites or Hg-loaded SiO2-

TiO2 nanocomposites were leached with 3 ml of SPLP solution on a rotoshaker® for 18 h. After

centrifugation, aliquots of the supernatant were removed and immediately used for toxicity

assays, and analyzed for THg determination following digestion with bromine monochloride

(Bloom and Crecelius 1983).


Our results show that the treatment imposed upon these slurries induces the methylation of

both background Hg initially present in the sediment (control) and Hg added as Hg-

nanocomposite complexes. The percentage of THg methylated from the Hg-SiO2-TiO2-

complexes added to slurries and corrected from control samples is shown in Figure 6-1 as a

function of pH. These results show an increasing and pH-dependent trend of methyl-Hg

production over time, with more methyl-Hg produced at the lowest tested pH. This trend

suggests that microbial Hg methylation could be controlled by the solubility of adsorbed Hg onto

MNs, which decreases with increasing pH. Although this trend can also be attributed to the effect

of pH change on Hg methylating microorganisms, it is rather clear that the detection of methyl-

Hg under these tested conditions, regardless of the amount produced, is a strong indication of the

bioavailability of Hg-sorbed onto SiO2-TiO2 nanocomposites.

106

Using the linear portion of the above described Hg methylation trends, and assuming that

the methylation rates of inorganic Hg would be far in excess of the rates of demethylation of

produced methyl-Hg (i.e., Km>>Kd), a pseudo first order kinetics assumption would then allow

the determination of the reaction rates at different pH (Figure 6-2). The determined constant rates

for Hg methylation in slurries spiked with Hg-SiO2-TiO2 complexes were km=0.02 day-1 and

km=0.004 day-1 at pH 4 and 5, respectively, and the constant rate approached zero at pH 6 as

methyl-Hg levels in sediment slurries became barely detectable. Additionally, the rate

determined at pH 4 was about one order of magnitude lower than the reaction rate observed in

sediments slurries spiked with free ionic Hg added as HgCl2. Although the experiment with

HgCl2 additions was used at the native sediment pH only (pH 4); the obtained results suggest that

Hg sorption onto nanocomposites delays its bioaccessibility. Overall, Hg adsorbed onto MN and

introduced into sedimentary environments could quickly become bioavailable and therefore toxic

in more acidic systems.

The concentration of THg bound onto the SiO2-TiO2 nanocomposites averaged 639.04 ±

366.31 ng Hg/g nanomaterials, while THg concentrations on plain materials were at or below our

analytical detection limits. The SPLP extraction procedure was comparatively applied to SiO2

and TiO2 nanoparticles, ultra-violet (UV) irradiated virgin SiO2-TiO2 nanocomposites, and the

Hg-contaminated UV-irradiated nanocomposites (Pitoniak et al. 2005). Obtained leachates were

then used for toxicity testing with FluoroMetPLATETM, which is specific to heavy metal toxicity

(Bitton et al. 1994). The toxicity results expressed as percent inhibition are presented in Figure

6-3. Based on this toxicity test, SiO2 and TiO2 nanoparticles are rather non-toxic despite the 4%

inhibition response obtained with SiO2 leachate. In contrast, both the virgin and Hg-

contaminated nanocomposites show a much higher toxicity with average inhibition values of

107

57% and 84%, respectively. It appears that both the support material (i.e. SiO2-TiO2

nanocomposites) and Hg adsorbed onto it contribute to the recorded inhibition. The toxicity of

virgin SiO2-TiO2 nanocomposites (57% inhibition) is likely due to its physicochemical

characteristics, while the much higher inhibition recorded with the Hg-contaminated SiO2-TiO2

nanocomposites indicates an additional toxicity effect due to the adsorbed toxic Hg. In sediment

slurries spiked with SiO2-TiO2 nanocomposites, the above observed toxicity could translate into

reduced Hg methylation as bacteria involved in Hg biotransformation become impacted.

Consequently, the detection of methyl-Hg in these methylation experiments points to the

potential bioavailability of Hg sorbed onto SiO2-TiO2 nanocomposites.

6.4 Conclusions

In summary, pollutants loaded MNs from nanotechnology-based remediation processes

could result in potential concerns related to the environmental fate of adsorbed pollutants as they

ultimately enter natural (e.g., waterways) or engineered (e.g. landfills) systems as waste streams.

When compared to sediments spiked with free HgCl2, the addition of MN-adsorbed Hg to

sediments slurries resulted in slower rates of methyl-Hg production and near total inhibition of

Hg methylation at pH≥6. Overall, these results point to the potential for bioaccessbility of MN-

adsorbed pollutants as a function of certain key environmental parameters such as acidic pH.

These preliminary results clearly illustrate the need for further research that could lead to

guidelines for the handling and disposal of nanowastes.

108

0

2

4

6

8

10

12

14

16

0 2 4 6 8

Time (days)

% T

Hg

conv

erte

d to

met

hyl-H

g

pH 4 pH 5 pH 6

Figure 6-1. Percent THg converted to methyl-Hg in sediment slurries spiked with SiO2-TiO2-Hg complexes and incubated at different pH. At pH 6, methyl-Hg concentrations were either below or at the analytical detection limit of 0.05 ng/g.

109

pH 6

pH 5k = 0.0041/day

r 2 =0.9755

pH 4 k = 0.02/dayr 2 =0.9328

0.00

0.05

0.10

0.15

0.20

0.25

0 2 4 6 8 10

Time (days)

-Ln(

1-([M

eHg]

/TH

g))

Figure 6-2. Kinetics of Hg methylation in sediment slurries spiked with SiO2-TiO2-Hg complexes at pH 4 (triangles; native pH), 5 (circles), and 6 (squares). The methylation of inorganic Hg decreases with increasing pH to reach values below the analytical detection limit at pH 6.

110

0

10

20

30

40

50

60

70

80

90

100

SiO2 nanoparticle TiO2 nanoparticle UV-irradiatednon-used TiO2-

SiO2nanocomposites

Hg-TiO2-SiO2complexes

% In

hibi

tion

Figure 6-3. Toxicity effect of Synthetic Precipitation Leaching Procedure (SPLP) solutions obtained from leaching of virgin and Hg-loaded SiO2-TiO2 nanocomposites in a 1:60 ratio (ml SPLP/mg nanomaterials). Results are expressed as percent (%) inhibition.

Not Detected

111

CHAPTER 7 CONCLUSIONS AND RECOMMENDATIONS

7.1 Conclusions

Nanotechnology is a highly promising and exciting cross-cutting molecular technology

that spans many areas of science and technological applications. However, due to the relative

novelty of this technology very little has been done to assess the risks to biological systems; and

concerns about the use of the products of nanotechnology are being increasingly expressed in

public and in the media (Colvin et al. 2003). Our current knowledge of the harmful effects of

nanoparticles remains very limited and data on biotic and abiotic transformations of MNs in

natural systems are limited.

This research focused on potential environmental impacts of MNs, and the use of

toxicological, hydrological and biogeochemical approaches has resulted in the following

conclusions.

• Toxicity testing was based on three different microbiotests emphasizing the interactions of

MNs with (i) biochemical processes (i.e. the MetPLATE test), (ii) the growth of a

unicellular freshwater green algae (P. subcapitata), and (iii) the survival of an aquatic

invertebrate (C.dubia). Both the suspending medium for MNs and MNs themselves

exhibited different degrees of toxicity as a function of concentration and testing methods.

The use of different toxicity methods is necessary to avoid erroneous results. This is

because different test organisms respond differently to different toxicants. Amongst the

nanometal particles tested, nano-silver and nano-copper displayed the highest toxicity.

Aqueous fullerene suspensions prepared by use of organic solvent (THF) and SWNT

suspensions in Gum Arabic (a non-toxic surfactant used in this study) showed higher

degrees of toxicity as compared to water sonicated suspensions. Although the toxicity

112

mechanisms were not addressed in this study, the toxicity of these MNs could be

attributable to their ability to generate highly reactive and toxic free radicals, their degree

of purity as toxic impurities increase toxicity, or simple surface interaction with cell

membranes.

• The suspension of selected toxic MNs (i.e. C60, nano-silver, and nano-copper) in natural water matrices with varying DOC content and ionic strength showed that toxicity results obtained from laboratory experiments that use drastic MNs suspension methods may not be realistic. It was found that the suspensions of MNs in natural waters varied significantly with water chemistry and particle chemical composition and reactivity.

• Using soil columns to assess the transport of SWNTs in heterogeneous porous media, it was found that soil texture/characteristics and solution chemistry (i.e. the composition of the liquid used to suspend SWNTs) affect the transport of this highly hydrophobic MN, as surface charges of the MNs influence their adsorption and dispersion in the porous media. Finally, the use of a convection-dispersion model was able to accurately predict SWNTs transport in sandy soils, with a strong correlation between data obtained experimentally and simulated ones.

• The effects of C60, nano-Ag and CdSe quatum dots on sediment microbial activity were studied in slurries. C60 appeared to be highly toxic to bacteria involved in organic matter oxidation, primarily nitrate and nitrite reducers. Nano-silver and CdSe quantum dots were less toxic at tested concentrations, but gave the indication of potential pronounced negative effects on microorganisms at much higher concentrations.

• The fate and transformation of an example pollutant adsorbed onto MNs indicated that under specific environmental conditions, MNs could act as carriers of the pollutant adsorbed onto them. If this constitutes an advantage with regard to medical research, it could have negative implications in the environment.

7.2 Recommendations

Based on our findings, the following recommendations are made to further the extent of

knowledge of the environmental implications of MNs:

• Conduct tests focusing on long-term effects of MNs. Although toxicity was detected in short term for all MNs tested, long term tests may be needed to explore their toxicity mechanism and the effect of their transformation in the environment.

• Detailed characterization of MNs (e.g. speciation) in aqueous systems is needed in order to better understand the behavior and fate of MNs in natural environments.

113

• The use of mathematical models to explore the transport patterns of MNs in different types of soils should be considered as it can help predict the potential dispersal of MNs in porous media.

• Although our study observed negative impacts of MNs on sediment microorganisms, further investigations are needed to pinpoint the microbial groups that are sensitive to MN toxicity. Molecular techniques could reveal the shifts in sediment microbial populations following contamination by nanomaterials.

Overall, since there is so little data available for aquatic environments, research is required

to test the behavior and particulate binding properties of MNs in both freshwater and seawater.

The relative importance of different biological routes of uptake also needs to be assessed in

representative aquatic species, since this will be a crucial factor governing intracellular behavior,

distribution, fate and toxicity of internalized MNs.

A major challenge remains the derivation of toxicity thresholds for MNs; and determining

whether or not currently available biomarkers of harmful effects are effective for the new

discipline of environmental nanotoxicology. If new methods are required to adequately assess

the toxicity of MNs, then such new methods will also need to be, linked if possible, with

functional ecosystem indices. Such linkages would be desirable in order to bridge the gap

between individual organism “health-status” and ecosystem-level functional properties.

114

APPENDIX A TESTED CONCENTRATIONS OF CARBON- AND METAL-BASED NANOMATERIALS

IN THREE DIFFERENT TOXICITY ASSAYS

Table A-1. Tested concentrations of carbon- and metal-based nanomaterials in three different toxicity assays

Tested Nanomaterials

Concentrations Used in 48-h Ceriodaphnia dubia Assay (mg/L)

Concentrations Used in 96-h P. subcapitata Chronic Assay (mg/L)

Concentrations Used in MetPLATE

Test (mg/L)

Nano-silver 0.05, 0.06, 0.07, 0.08, 0.09, 0.1 0.1, 0.15, 0.2, 0.25, 0.3 4, 8, 16, 24, 32

Nano-copper 0.3, 0.4, 0.5, 0.6, 0.7 0.5, 0.52, 0.54, 0.56, 0.58, 0.6 4, 8, 16, 24, 32

Nano-cobalt 1.5, 1.6, 1.7, 1.8, 1.9 0.5, 0.6, 0.7, 0.8, 0.9 NA

Nano-nickel 0.3, 0.4, 0.5, 0.6, 0.8 0.25, 0.3, 0.35, 0.4, 0.45 NA

Nano-aluminum 3, 3.5, 4, 4.5, 5, 5.5 4, 6, 7, 8, 9 NA

Fullerenes (C60) 0.27, 0.41, 0.54, 0.68,

0.81 0.11, 0.14, 0.16, 0.19,

0.22

SWNT suspended in Gum Arabic

(GA) 0.24, 0.26, 0.28, 0.3, 0.32 1.37, 1.92, 2.47, 3.02,

3.57,

115

APPENDIX B TESTED CONCENTRATIONS OF MANUFACTURED NANOMATERIAL SUSPENSIONS

IN TOXICITY TESTS USING THE C. DAPHNIA 48-H ACUTE TOXICITY ASSAY AND METPLATE TEST

Table B-1. Tested concentrations of manufactured nanomaterial suspensions in toxicity tests using the C. daphnia 48-h acute toxicity assay and MetPLATE test

Nanomaterial Suspensions

Concentrations Used in 48-h Ceriodaphnia dubia Assay

(mg/L)

Concentrations Used in MetPLATE Test (μg/L)

Ag-SR1 5.4, 5.94, 6.48, 7.02, 7.56 NA

Ag-SR2 0.68, 0.72, 0.765, 0.81, 0.85 NA

Ag-SR3 0.59, 0.66, 0.73, 0.79, 0.83 66.27, 79.52, 92.78, 106.03, 119.29, 132.54

Ag-DI 432.9, 449.55, 466.2, 482.85, 499.5 33.3, 41.63, 49.95, 58.28, 66.6

Cu-SR1 32.72, 35.24, 37.75, 40.27, 45.30 12.58, 18.88, 25.17, 31.46, 37.75

Cu-SR2 6.66, 6.95, 7.24, 7.53, 7.82 28.96, 57.92, 86.88, 115.84, 144.8, 173.76, 202.72

Cu-SR3 25.5, 30.6, 35.7, 40.8, 45.9, 51 127.5, 153, 178.5, 204, 229.5, 255

Cu-DI 1.61, 1.88, 2.14, 2.41, 2.68 NA

C60-SR1 NA NA

C60-SR2 NA NA

C60-SR3 NA NA

C60-DI NA NA

116

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BIOGRAPHICAL SKETCH

Jie Gao was born in Huaian, Jiangsu Province, China. She received her bachelor’s degree

in environmental science from Nanjing University in 2003. After one year study in Beijing

University graduate school, Jie Gao was admitted as a PhD student with an Alumni Fellowship

award in the Department of Environmental Engineering Sciences at the University of Florida.

She received Master’s degree in 2005 and has been ever since working on her doctoral research

in assessment of environmental impacts of nanomaterials under the supervisory of Dr. Jean-

Claude Bonzongo.

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