Potential Environmental Implications of Manufactured Nanomaterials

Material Information

Potential Environmental Implications of Manufactured Nanomaterials Toxicity, Mobility, and Nanowastes in Aquatic and Soil Systems
Gao, Jie
Place of Publication:
[Gainesville, Fla.]
University of Florida
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1 online resource (135 p.)

Thesis/Dissertation Information

Doctorate ( Ph.D.)
Degree Grantor:
University of Florida
Degree Disciplines:
Environmental Engineering Sciences
Committee Chair:
Bonzongo, Jean-Claud
Committee Members:
Bitton, Gabriel
Delfino, Joseph J.
Ma, Lena Q.
Kopelevich, Dmitry I.
Wu, Rongling
Graduation Date:


Subjects / Keywords:
Carbon nanotubes ( jstor )
Fullerenes ( jstor )
Nanomaterials ( jstor )
Nanoparticles ( jstor )
Nanotechnology ( jstor )
Sediments ( jstor )
Silver ( jstor )
Slurries ( jstor )
Soils ( jstor )
Toxicity ( jstor )
Environmental Engineering Sciences -- Dissertations, Academic -- UF
mercury, mobility, nanomaterial, nanowaste, toxicity, transport
Electronic Thesis or Dissertation
born-digital ( sobekcm )
Environmental Engineering Sciences thesis, Ph.D.


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. 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. 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. ( en )
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In the series University of Florida Digital Collections.
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Includes vita.
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Thesis (Ph.D.)--University of Florida, 2008.
Adviser: Bonzongo, Jean-Claud.
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by Jie Gao.

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2 2008 Jie Gao


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


4 ACKNOWLEDGMENTS I would like to sincerely thank m y 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 me mbers, 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 st udent 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 collaborati on 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. On ly 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 hi s support, patience, and unwavering love throughout the past seven years of my colle ge undergraduate and graduate studies.


5 TABLE OF CONTENTS page ACKNOWLEDGMENTS...............................................................................................................4 LIST OF TABLES................................................................................................................. ..........8 LIST OF FIGURES.......................................................................................................................10 ABSTRACT...................................................................................................................................12 CHAP TER 1 NANOTECHNOLOGY AND THE ENVIRONM ENT: AP PLICATIONS AND IMPLICATIONS................................................................................................................... .15 1.1 Problem Statement.......................................................................................................... ..15 1.2 Production and Use of Nanomaterials.............................................................................. 18 1.3 Potential Toxicity of Ma nufactured Nanom aterials......................................................... 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 Fullerene (C60).............................................................................................38 Single-Walled Nanotubes (SWNTs)...........................................................39 Metallic nanomaterials................................................................................40 2.4 Conclusions.......................................................................................................................43


6 3 TOXICITY OF SELECTED MANUFACT URED NANOMATERIALS DISPERSED IN NATUR AL WATERS WITH GRADIENTS IN IONIC STRENGTH AND DISSOLVED ORGANIC MATTER CONTENT..................................................................55 3.1 Introduction............................................................................................................... ........55 3.2 Materials and Methods.....................................................................................................56 3.2.1 Collection of Water Samples.................................................................................. 56 3.2.2 Preparation of Nanomaterial Suspen sions in Collected Water Sam ples................ 57 3.2.3 Determination of MNs Concentra tions in Prepared Suspensions .......................... 57 3.2.4 Toxicity of MNs Suspended in Natural Waters..................................................... 58 The Ceriodaphnia dubia assay .....................................................................58 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 Nanom aterials 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 CA RBON NANOT UBES (SWNTS) IN SATURATED HETEROGENEOUS POROUS MEDIA...................................................... 74 4.1 Introduction............................................................................................................... ........74 4.2 Materials and Methods.....................................................................................................76 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 MANUFAC TURED NANOMATERIALS ON BIOGEOCHEMICAL PROCESSES IN SEDIMENTS......................................................... 89 5.1 Introduction............................................................................................................... ........89 5.2 Materials and Methods.....................................................................................................91 5.2.1 Preparation of Nanomaterial Suspensions.............................................................. 91 5.2.2 Sediment Collection...............................................................................................91 5.2.3 Dominant Terminal Electron Accepting Pro cesses (TEAPs) in Sediments and Sediment Manipulation in this Study........................................................................... 91 5.2.4 Analytical Techniques............................................................................................95 5.2.5 Data Analysis.......................................................................................................... 95


7 5.3 Results and Discussion..................................................................................................... 95 5.4 Conclusions.......................................................................................................................98 6 NANOWASTES IN THE ENVIRONMENT: T HE 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 RECOMME NDATIONS............................................................... 111 7.1 Conclusions.....................................................................................................................111 7.2 Recommendations...........................................................................................................112 APPENDIX A TESTED CONCENTRATIONS OF CARBONAND METAL-BASED NANOMATERIALS IN THREE DI FFERENT TOXICITY ASSAYS .............................. 114 B TESTED CONCENTRATIONS OF MANUFACTURED NANOMATERIAL SUSPE NSIONS IN TOXICITY TESTS USING THE C. DAPHNIA 48-H ACUTE TOXICITY ASSAY AND METPLATE TEST...................................................................115 LIST OF REFERENCES.............................................................................................................116 BIOGRAPHICAL SKETCH.......................................................................................................135


8 LIST OF TABLES Table page 1-1 Examples of materials and app lications of nanotechnology (K arn 2004)......................... 26 1-2 Examples of proposed environmental app lications of manufactured nanomaterials......... 27 2-1 Tested surfactants and solvents and their chem ical compositions..................................... 45 2-2 Characteristics of metallic nanoparticles used in toxicity experim ents in this study........46 2-3 Chemical composition of the prelimin ary alg al assay procedure (PAAP) culture medium......................................................................................................................... .....47 2-4 Concentrations of tested surf actants resulting in le thal 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 96h P. subcapitata chronic toxi city assay..............................................................................49 2-6 Percent mortality of C. dubia exposed to solutions with increasing THF concentrations in 48-h a ccute toxicity assay. ..................................................................... 50 2-7 Concentrations of tested m etaland carbonbased nanopa rticles resulti ng 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 m etaland carbonbased nanopa rticles resulti ng 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 metaland carbonbased nanoparticle s 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 th e sandy (G ainesville, Florida) and clayey (Atlanta, Georgia) soils used in co lumn 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 carbonand m etal-based nanomat erials in three different toxicity assays..................................................................................................................114


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


10 LIST OF FIGURES Figure page 1-1 Conceptual diagram of the life cycl e and potential pathways and fate of m anufactured nanomaterials.............................................................................................. 25 2-1 Size distribution of selected metallic MN sam ples as obtained from commercial sources................................................................................................................................44 3-1 Map of the Suwannee River watershed a nd tributaries showin g the th ree sampling locations.............................................................................................................................67 3-2 Concentrations of silver, copper, and C60 in different water samples spiked with individual nanomateria l and then filtered.......................................................................... 68 3-3 Linear correlation of Cu concentrations and dissolved organic m atter (DOC) in water samples SR1, SR2 and DI water (R2=0.9864)...................................................................69 3-4 48-h LC50 values of silverand copp er-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 nanosilverand nanocopper suspensions using MetPLATE test................ 72 4-1 Schematic diagram of the experimental setup for SW NTs 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 sim ulated data of SWNT-GA and SWNTSDS suspensions in sandy soil columns............................................................................ 86 5-1 Kinetics of acetate degradation, and nitrat e, n itrite 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 nitrat e, n itrite and sulfate concentrations in sediment slurries spiked w ith excess nitrate and without (controls) or with tested nanomaterial additions (C60, nanosilver, and CdSe quantum dots)................................. 100 5-3 Kinetics of acetate degradation and nitrat e, n itrite and sulfate concentrations in sediment slurries spiked with excess sulfat e and without (controls) or with tested nanomaterial additions (C60, nanosilver, and CdSe quantum dots)................................. 101


11 5-4 Pseudo-first order kinetics of acetate di sappearan ce 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 sedim ent slurries spiked with SiO2-TiO2-Hg complexes and incubated at different pH......................................................................... 108 6-2 Kinetics of Hg methylation in se diment slurries spiked with S iO2-TiO2-Hg complexes at pH 4 (triangles; native pH), 5 (circles), and 6 (squares)............................ 109 6-3 Toxicity effect of Synthetic Precipi tation Leaching Procedure (SPLP) solutions obtained fro m leaching of virgin and Hg-loaded SiO2-TiO2 nanocomposites in a 1:60 ratio (ml SPLP/mg nanomaterials)................................................................................... 110


12 Abstract of Dissertation Pres ented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy POTENTIAL ENVIRONMENTAL IMPL ICATIONS OF MANUFACTURED NANOMATERIALS: TOXICI TY, MOBILITY, AND NANOW ASTES IN AQUATIC AND SOIL SYSTEMS By Jie Gao August 2008 Chair: Jean-Claude Bonzongo Major: Environmental Engineering Sciences Nanotechnology has been singled out by industr y and governments to become the worlds largest industrial revolution, and it carries the po tential to substantiall y benefit environmental quality through pollution prevention, treatm ent, and remediation. However, nanotechnology could also lead to serious environmental problem s 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 develo p an understanding of the potentially complex interplay between MNs and the health of organi sms and ecosystems. The potential effects of MNs were evaluated by testing th e hypothesis that: chemical elem ents used in the production of MNs could lead to environmental dysfunctions due to : (1) the potential toxici ty of these elements and their derivatives, (2) the small size driv en 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 mi crobial catalyzed reactions, and (4) the large surface area which would allow MN s to act as carriers/delivers of pollutants adsorbed onto them.


13 To address this broad hypothesis, three well-e stablished small-scale toxicity tests (i.e. the Ceriodaphnia dubia acute toxicity test, the Pseudokirchneriella subcapitata chronic toxicity test, and MetPLATE), were used. In addition, studie s at the system level were conducted using a combination of column and batch experiments to investigate the transpor t behavior of MNs in heterogeneous porous media and the interactions of MNs with microbial-c atalyzed oxidation of organic matter in sediments. Carbon (i.e. fullerenes (C60), single-walled carbon nanot ubes (SWNTs)) and metal (i.e. CdSe quantum dots, and powders of the following nanometalsAg, 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 nanoAg being the most toxic. The use of experimental conditions that mimic li kely scenarios of MNs introduction to aquatic systems showed that the toxicity response of te st model organisms to MNs under such conditions would be affected by ke y water quality parameters such as organic matter content and solution chemistry. Column st udies of SWNTs transport in heterogeneous porous soils showed that soils characteristics and the chemical composition of MN suspensions affect transport behaviors, and that the latt er can be quantitativel y predicted by use of mathematical models such as the convection-di spersion 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 sedi mentary environments as a function of pH.


14 Overall, these findings help shed light in the poorly studied environm ental implications of MNs. However, several questions remain unanswered as these short-term laboratory investigations may not be able to predict the fate and transpor t of MNs on a long-term basis.


15 CHAPTER 1 NANOTECHNOLOGY AND THE ENVIRONM ENT: AP PLICATIONS AND IMPLICATIONS 1.1 Problem Statement Nanoscience and nanotechnologies are currently generating an extraordinary interest as they carry expectations that are believed to br ing about changes as profound as the industrial revolution, antibiotics, and nucle ar weapons all in one (Ajaya n et al. 1999). Manufactured nanoparticles (MNs) find use in a wide variety of hum an activities, from medical, electronics, to environmental research and a pplications. But, while the a dvantages of nanoscience and nanotechnologies are multiple with many more still to be discovered, the potential implications of this new technology on the environmen t and living organisms remain largely unknown (Fischer and Chan 2007; Oberdorster 2004; Seetha ram 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 compar tments. 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 en ter 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 pot ential 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, en vironmental, and medi cal applications (Bor derieux 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 1 999). Unfortunately, and despite the current


16 ongoing discussion on MNs, the study on the environm ental fate and impacts of nanoparticles remains a frontier science. Although growing ra ther 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 Roya l 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 ( Amongst several ke y points highlighted by the study, the report noted the lack of published data on negative imp acts 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, e xperimental data along the lines of the above-mentioned recommendations have been increasing in the lite rature. For instance, re search on toxicity of MNs on living organisms is increasing very rapidl y, starting from some in itial laboratory studies exposing fish to carbonaceous MNs (Oberdorst er 2004) to the widesp read 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 mode l 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 e xposing 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 (Oberdorst er 2004). The observed damage, the lipid peroxidation is known to impa ir the normal functioning of cell membranes and


17 has been linked to illnesses such as Alzheimers disease in humans. Besides the carbon based MNs, metal nanoparticles have also shown a tendency for both bioaccumulation and toxicity. Using optical and transmissi on electron microscopy, Xu et al. (2004) showed that nonfunctionalized Ag nanoparticles as large as 80 nm in diameter could cross the cell membrane of Pseudomonas aeruginosa. However, the tested Ag-nanopa rticles did not damage the cell membrane. Similar tests on E. coli and red blood cells using functionalized gold (Au) nanoparticles showed moderate toxicity caus ed by lysis of cell membranes mediated by nanoparciles cross-linked with catio nic 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 trophi c levels. Unfortunately, the exte nt and potential effects of such transfer, if any, remain unknown. In addition to toxicity studies, and although sti ll 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 coul d exhibit different transport be haviors in porous media. These findings are important with regard to the asse ssment of both the efficacy (e.g., when used for remediation purposes) and envir onmental impacts of MNs. The field of nanotechnology is so new that ther e has been little time to develop substantive data on exposure and hazards. To date, a few studi es have investigated the toxicological effects of both direct and indirect expos ure to nanomaterials, but unfort unately, no clear guidelines for these effects exist (Colvin 2003). On the other ha nd, the environmental toxicity of nanomaterials is not predictable from that of similar but co mmon larger compounds. Finally, despite the fact that there has been a recent increasing effort to understand ecosystems at the system level, both


18 theoretical and experimental i nvestigations of the impacts of MNs on ecosystem functions are still lacking. 1.2 Production and Use of Nanomaterials Nanotechnology has been pegged by industry an d governm ents to become the worlds largest and fastest industrial revo lution (Roco et al. 1999), due to its vast potential for use in different fields such as medi cine, electronics, chemistry, a nd engineering (Kung and Kung 2004; Navalakhe and Nandedkar 2007). According to the U.S. National Institute fo r 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 applic ations, as well as consumer products ranging from anti-aging cream to sunblocks (ETCGroupGenotype 2004; Hund-Rinke and Simon 2006; Wiesner 2003). The pr oduction and synthesis of MNs have been accomplished by a wide variety of techniques incl uding 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 antimicrobi al 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 na notechnology 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 (ETC Group 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 pro cesses for pollution prev ention and treatment


19 (Zhang and Karn 2005). Table 1-2 summarizes th e 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 th an their bulk particles be cause of larger surface area, enhanced chemical reactivity and potenti al for cell penetration (Monteiro-Riviere and Orsire 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 degr adation 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 biologica l 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 th e system (Karakoti et al. 2006). They can enter human tissues via the lungs after inhalation, thr ough the digestive system, and thr ough 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 observe d 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


20 al. 2005). Nano-copper particles can distribute throughout the body into blood, brain, lung, heart, kidney, spleen, liver, intestine and stomach, and pr ovoke dysfunctions of the organs (Chen et al. 2006). Nano-silver particles, quant um dots and some other metal ox ide nanoparticles are able to produce ROS as well (Adams et al. 2006; Cho et al. 2007; Jain and Pr adeep 2005; Kim et al. 2007; Stoimenov et al. 2002; Wang et al. 2007). Although still in its infancy, the emerging fi eld of nanotoxicology is fast growing. For instance, standardized te st methods are lacking in order to assess MNs safety, interpret extralaboratory data, and generate an online databank th at is accessible to al l users and manufacturers (Kovochich et al. 2007). Therefore, knowledge of exposure to and hazard of nanomaterials are in great need to fully understa nd their risks and impacts. 1.4 Potential Effects of MNs on Ecosystem Functions Ecosystem s around the world accomplish numerous natural services and most if not all of them seem to have common main characteris tics, 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 quan titatively characterize any of the above listed na tural services can be used to assess the impact of pollutants on ecosy stem functions. Such ab ilities are provided by thermodynamics, which has been successfully ap plied in describing th e basic properties of ecosystems (e.g. flow of matter and energy). By co mbining 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 car bon can be used as proxy to de tect the potential impacts of MNs on specific basic ecosystem functions. For inst ance, in pristine soils and sediments, the composition and distribution of different micr obial populations are usually well-established, although changes associated with shift in seasons and other major parameters are also common.


21 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 respons e may not be expressed. Although the ability of MNs to behave like some lipophylic pollutants su ch 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 eco logical consequences due to the combination of bioaccumulation/biomagnificati on phenomena and the rather well-established toxicity of some of these nanoparticles. Theref ore, 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 th at 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 actua lly 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 shou ld raise concerns about MN-impacts at a system level.


22 1.5 Environmental Fate and Transport of MNs Nanom aterials 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 Zha ng 2003). If MN applications deve lop as projected, nanoparticle introduction to ground waters and soils could pr esent significant challe nges with regard to exposure risk for both aquatic/soil organisms and human health (Colvin 2003). The aquatic environment may become contaminated through di scharges 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 additi on, their use in newly developed environmental remediation techniques (e.g. groundwater and soil remediation) would lead to intentional introducti on of MNs to natural systems. Based on current knowledge, the small size of MNs increases th e potential for their dispersi on (e.g. enhanced mobility) and exposure (e.g. crossing of cell membranes); whil e their large specific surface area increases chemical reactivityand could facilitate adsorpti on and transport of other toxic pollutants in the environment. Also, when released to the envi ronment, MNs might undergo transformations due to aggregation, sorption/desorption, deposit ion and bio-uptakeand identifying factors controlling these processes is the key to predicting their environmental fate and impact (Arnall 2003). This is because changes in MN surface ch emistry 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 therefor e, 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) show ed that the mobility of different MNs in porous media could differ substant ially. For instance, fullerol (C60 hydroxide, C60(OH)m, m = 2226) 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-bycase 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 asse ss the potential impact s 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), a nd (iv) the fate of pollutants adsorbed onto MNs, once released to the environment. Therefore, following this introductory chapter on background information on nanotechnology a nd 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 smallscale bioassays: (1) an aquatic invertebrate based testthe 48-h Ceriodaphnia dubia short-term assay; (2) a freshwater algal-based testthe 96-h Pseudokirchneriella subcapitata or S. capricornutum chronic assay; and (3) a bacterial produced enzyme based testthe Me tPLATE 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 iden tified as toxic in Chapter 2 were tested for toxicity after being suspended into natural waters of diffe rent chemical compositions. The effects of organic matter content and ioni c strength are investigated. Chapter 4 deals with the mobility aspects of sing le-walled carbon nanotubes (SWNTs) in heterogeneous porous media and the use of a modeling tool to predict the transport behavior of SWNTs versus soil char acteristics 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 bi oavailability of adsorbed Hg to sediment microorganisms. Chapter 7 provides a general conclusion and emphasi zes 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 (ada pted from Helland et al. 2007). Synthesis methods NANOPRODUCTS AIR SOIL WATER End of Life Property Change-post-synthesis -intermediate production Environmental Health End Points Living Organisms Environment:Property Change ExposureRelease Release Release Release End of Life -post-synthesis -intermediate production Environmental Health End Points Environment:Property Change ExposureRelease Release Release ReleaseNanoProducts Synthesis methods NANOPRODUCTS AIR SOIL WATER End of Life Property Change-post-synthesis -intermediate production Environmental Health End Points Living Organisms Environment:Property Change ExposureRelease Release Release Release End of Life -post-synthesis -intermediate production Environmental Health End Points Environment:Property Change ExposureRelease Release Release ReleaseNanoProducts


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 manuf actured 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 Nanom aterials and nanotechnologies have become the worlds largest and fastest industrial revolution, with an anticipated cap acity to affect many industrial ac tivities and lead to discovery and implementation of unique materials and devices from electronics to engineered tissues (Roco et al. 1999). Despite these anticipated applications, th e products of nanoscience and nanotechnologies also raise concerns about thei r potential health and environmental impacts (Isobe et al. 2006). Thus, while much research effort is currently directed toward exploring the properties and applicati ons of nanomaterials in medicina l, industrial, agricultural, and environmental fields, the body of experimental work on the potential implications of nanomaterials on living organisms and ecosystem f unctions 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; Ober drster 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 impli cation 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 potenti al applications in biological sensing and drug delivery systems (Zhang et al. 2005). Howeve r, current preparation methods lead to the production of SWNTs aggregates, and without fu rther treatment, the de gree of aggregation becomes a limiting factor in using these nanotub es (Zhang et al. 2005). On the other hand, since the toxicity of such nanoparticles is size-depe ndent, their uncontrollable aggregation behavior would likely make the interpretation of experiment al results rather difficult. Therefore, the


29 determination of different physical characteristics of prepared suspensions is necessary prior, and if possible, during toxicity stud ies (Oberdrster et al. 2005). Fullerenes (C60 or buckyballs) and SWNTs have been among the most widely studied and used carbon-based nanometerials due to their un ique structural and el ectronic properties that enable numerous industrial, medical and environm ental 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 Byrn e 2008). For instance, methods have been proposed for preparation of fullerene suspensi on in water using a step-wise approach that includes dissolution into organic solvent followed by a back extracti on 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 g/mL (Yamakoshi et al. 1994). Additionally, it has been reporte d that SWNTs can be disper sed in a number of aqueous surfactant solutions such as s odium dodecyl sulfate (SDS), sodium dodecylbenzene sulfonate (NaDDBS), and Triton X-100 (Wan g et al. 2004; Zhang et al. 2005). Unfortunately, although the surfactant induced functi onalization of SWNTs enha nces their suspension/d ispersion, it can also affect their inherent properties (Garg and Sinnott 1998), and likel y, modify their interactions with and impacts on living cells. It is estimated that the worldwide market for nano-products woul d reach $1 trillion by 2015 (Roco 2005). This anticipated large produ ction 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 th e human body via respiratory routes, skin or digestive system exposures and persist without being phagocytosed (Karakoti et al. 2006). Nanomaterials could even distri bute throughout out the entire body, cross the bl ood-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 activ ely (Tetley 2007). Realizing the toxicological effects of nanoparticles, scien tists have tested and demonstr ated 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 in cluding (i) the chemical composition and the ability of a given nanoparticle to re lease free radicals, (ii) the particle size and geometry, (iii) surface area and reactivity, (i v) surface treatments or modifications, (v) the degree of aggregation/agglomerati on, and (vi) particle electrost atic 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 manufact ured 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 enviro nmental 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; Pa pis et al. 2007). However, results obtained from most of thes e toxicity experiments are not eas ily used to predict the actual effects of nanomaterials in natura l systems (Fischer and Chan 2007).


31 In this study, a wide variety of carbonand metal-based nanoma terials is screened for their potential toxicity using thr ee different small-scale micro-biotests (i.e. MetPLATE, Ceriodaphnia dubia and Pseudokirchneriella subcapitata tests). The ultimate objective is to identify nanomaterials with toxicity eff ects and use such toxic nanoparticl es in further investigations focusing on their potential imp acts in natural systems. 2.2 Materials and Methods 2.2.1 Chemicals Besides water, different solvents and surf actants used in the preparation of MNs suspensions could lead to erroneous toxicity resu lts. This is because certain solvents/surfactants can be highly toxic to some organisms. To elimin ate this uncertainty, the toxicity of 6 surfactants and 2 organic solvents commonly used to obtai n highly dispersed car bon-based nanomaterial suspensions was determined first (see Table 2-1). Besides the poly(viny lpyrrolidone) (PVP) and tetrahydrofuran (THF) which were obtained from Fisher Scientific (Atlant a, 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 concentrat ed solution of 50 g/L for PVP, and 10 g/L solutions for each of the other surfactants. Th ese stock solutions were then used to prepare necessary dilutions for toxicity tests. THF solutions were prepar ed by direct dilution of aliquot volumes into Nanopure water to produce samples with incr easing concentration gradients and then tested for toxicity. 2.2.2 Preparation of Nanomaterial Suspensions Single walled carbon nanotubes (SWNTs) suspensi ons were prepared using an initial mass of ~ 40 mg of raw SWNTs (R ice 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 g/mL. To obtain a highly disperse d suspension from this initial


32 solution, the mixture was homogenized using a hi gh-shear IKA T-25 Ultra-Turrax mixer for ~ 1.5 h followed by ultrasonication using a Miso nix 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. Exce ss solids were later filtered out using a 0.45 m PTFE membrane filter, resulting in a transparent pi nk 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 so lution was vacuum-filtered through 0.45 m cellulose into a flask and stored in the dark. To determine th e 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 w ithdrawn for absorbance measurement at 334 nm (Deguchi et al. 2001) using a Hach DR/4000U Spectrophotometer. The nano-metals, nano-copper, nano-silver, nanonickel, and nano-cobalt used in this study were provided gratis by Quantum Sphere Incorp orated (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 ( ). Density and specific surface area were measured using a Quantachrome Nova 1200 (Quantachrome, Syossett, NY, USA) and zeta potential was measur ed 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 pe rformed with a range of doses up to 1 g/L for the surfactants and PVP. Concentrations abov e 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 pr oposed 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 prelim inary 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 m and sterilized by autoclaving. A pure culture of P. subcapitata was obtained from Hydrosphere Resear ch (Alachua, FL) and grown in PAAP medium with EDTA at around 25C. Light source (86 8.6 E m-2s-1 or 400 40ft-c) and c ontinuous aeration were provided 24 hours per day. New cultures were prep ared 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 medi a spiked with increasi ng concentrations of individual tested nanomaterialss ee Appendix A) and negative cont rols 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 chlor ophyll (chl. a) using a TurnerR QuantechTM digital filter fluorometer. In this case, the measured toxicity endpoint wa s growth inhibition (I). On the basis of the obtained chl. a concentrations, the IC50 (i.e. concentration that inhi bits 50% of algal growth) was determined by first plotting sample concentratio ns (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 sampleachl achl % I (2-1) S Y ICerceptint 5050 (2-2) Where [chl. a] corresponds to measured chlorophyll concentration; Yintercept is the inhibition value corresponding to the intercep t 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 C 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 trou t chow). The daphnia were fed every other day with 6.67 mL YCT and 6.67 mL alg ae 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 acco rding to EPAs protoc ol (USEPA 2002). MHW served as negative control and as the diluent to prepare media with increasing concentrations of tested nanomaterials. Neonates less than 24 hour s were separated from adults and fed 2 hours prior to test start. For each test, groups of 5 daphnia neonate s each were transferred into several 30 ml plastic cups containing 20 ml of MH W (controls) or MHW plus nanomaterials (treatmentssee Appendix A). There was no f eeding during the tests and survival was determined visually after 48 hours using death a nd/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 concentratio ns in triplicates (Appendix A). Finally, the LC50 values of daphnia test and associated 95% confidence intervals (CIs) were calculated by the probit an alysis (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 MetPL ATE 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 -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 hydr ated with 5-ml of MHW and thoroughly mixed by vortexing. Then, a 900 l aliquot of the MN suspension (or 900 l of MHW


36 for negative controls) was added to a test tube containing 100 l of the above-described bacterial reagent. The test tubes were then vortexed a nd incubated for 1.5 hours at 35 C. Following the incubation period, a 200l aliquot was transferred to a 96-well microplate to which 100 l of the enzyme substrate (i.e. CPRG) was added. After mi xing, the microplate was incubated at 35 C to allow color development. The response was quan tified 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 determin ed 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-m et als tested are shown in Figure 2-1. While aggregation resulted in increased mean partic le diameter, nano-copper, nano-silver, and nanocobalt still had a significant fraction of particle s 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 la rgest surface area (50.6 m/g), a nd although nano-silver particles showed rather nanoscale particles (as defined by nanotechnology, <100 nm), the particles surface area (14.5 m/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 ( = -0.69 mV, Table 2-2).


37 2.3.2 Toxicity of Solvents and Surfactants Using the algal and daphnia tests described above (MetP late not used because of its specificity to heavy metal toxicity), the toxi city of a number of surfactants/solvent was determined (Tables 2-4 and 2-5). Among the test ed dispersing agents, on ly 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 surfacta nts varied with both concentration (e.g., Smith et al. 2007) and type of tested model organisms. The results indicate that solution matrix us ed 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 sele cted for this study. The latter was then used to prepare all SWNT susp ensions 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 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; Yamakos hi 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 detect ed in toxicological studies (Isakovic et al. 2006) and the pathway of ROS production has b een 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 li ght (Isakovic et al. 2006; Sayes et al. 2004). 1C60 1C60 3C60 1C60 3O2 1O2 Triplet Quencher (Q) h Intersystem Crossing


39 As stated earlier, C60 particles are hydrophobic and st able water suspension can be obtained by use of organic solvent (e.g. THF). A lthough mass spectroscopy of solvent after this procedure indicated no residual THF in solu tion (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, alt hough undetectable through analysis of the aqueous phase, could have toxic implications when nanopa ticles interact with living cells. The positive aspect in favor of THF remains its actual concen tration as demonstrated by Harhaji (2007) and in this study, pure THF would not cause cell death at concentrations up to 10 g/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 ar tificial light was used. 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, car bon nanotubes could be taken up by aquatic organisms, such as particle feeders with poten tial negative effect s (Roberts et al. 2007). Similar to C60, SWNTs can also generate ROS and cause oxida tive stress (Helland et al. 2007; Manna et al. 2005), but their potential to i nduce 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 toxicolo gical 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 el ectron 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 th at iron residual in unrefined SWNTs could catalyze the production of free radicals through th e following reactions: Fe2+ + O2 Fe3+ + O2 (2-3) Fe3+ + AredFe2+ + 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) demonstr ated that highly purified SWNTs still caused cell membrane damage, due to direct physical in teractions between nanom aterials 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 unrefin ed because they are better dispersed and could pierce or permeate cell memb ranes (Tian et al. 2006). The SWNTs used in this study were not purifie d, i.e. they contained a certain amount of impurities, mostly iron (Dr. Ziegle r, UF-Chemical Engineering Dept personal communication ). Therefore, the toxicity detected here may resu lt from any or a combination of the above listed factors. Metallic nanomaterials Nanoscale copper and silver resulted in toxic re sponses regardless of th e toxicity test used, with the c alculated EC50s (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 spec ificity to bioavailable metals. In addition, MetPLATE is not very sensitive to nickel, coba lt and aluminum (Bitton et al. 1994). Amongst all tested nanometals, nano-aluminum particle s 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, nanocopper particles can easily distribute throughout the body of vertebrates into blood, brain, lung, heart, kidney, spleen, liver, in testine 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 organisms 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 di ssolution 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 efficien t 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 additio n, he found no correlation between toxicity and particle specific surface areas, wh ich 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 homeos tasis (Glover et al. 2005 ). Similar to most nanoparticles, the nanosize aggravates the toxicity of silver with the sma ller particles being more toxic. Overall, no strong relationship between partic le 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 nanosilver) showed the highest toxici ty. Compared to copper and silver nanoparticles, less effort has been made toward understanding the aquatic toxic ity of other nanometals. Zhang et al. (2006) have compared the toxicity of ultrafine a nd standard-sized nickel particles and found significantly higher effects with the ultrafine nickel However, they didnt elucidate the toxicity mechanism but speculated that generation of free radicals might be the reason. Although MetPLATE didnt show negative re sults 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 pa rticles, which later dissolve and damage the cell (Limbach et al. 2007). Papis et al. (2007) found that nanocobalt could ac tivate cellular pathways of defense and repair mechanisms but no molecular mechanism was concluded. A study by Braydich-Stolle et al. (2005) invol ved 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 nanom olybdenum (nano-Mo). Hussain et al. (2005) revealed that nanoaluminum didnt display toxicity up to th e concentration of 100 ppm. For all the nanometals, their test ed 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 par ticle dispersing agents on m odel te st organisms. The results show that surfactants common in the manufacturing and dispersion of carbon-ba sed nanomaterials can add to the toxicity of tested nanoparticles. A ccordingly, 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 al so the role of the fluid used to maximize the dispersion effect. This study also demonstrates th e need for multiple toxicity tests in assessing the potential implications of nanomaterials. De pending on the sensitivity of test organisms and route of exposure, a nanomaterial can be highly toxic to a given te st organism while not toxic at all to another. Finally, the obs erved toxic responses in lower trophic level aquatic organisms support the need for vertebrate toxi city-based studies. In addition, the investigation of the ability of nanomaterials to bioaccumate and ultimat ely 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 cons ists 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, 3tetramethylbutyl) phenyl) ether Triton X-100 (C2H4O)nC14H22O Tetrahydrofuran (THF) C4H8O Poly(vinylpyrrolidone) (PVP) (C6H9NO)n


46 Table 2-2. Characteristics of me tallic nanoparticles used in toxi city experiments in this study Nanometals Tested for Toxicity Specific Surface Area (m/g) Zeta Potential ( ) (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 MgCl2H2O CaCl2H2O MgSO4H2O K2HPO4 NaHCO3 25.5 12.2 4.41 14.7 1.04 15 Micronutrient Concentration ( g/L) H3BO3 MnCl2H2O ZnCl2 CoCl2H2O CuCl2H2O Na2MoO4H2O FeCl3H2O Na2EDTA2H2O Na2SeO4 185 416 3.27 1.43 0.012 7.26 160 300 2.39


48 Table 2-4. Concentrations of tested surfactants re sulting 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 resulti ng 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 metaland carbonbased 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 metaland carbonbased nanoparticle s 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 te sted metaland carbonbased 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 e ffect was observed with all the other tested nanoparticles.


55 CHAPTER 3 TOXICITY OF SELECTED MANUFACTUR ED NANOMATERIALS DISPERSED IN NATURAL WAT ERS WITH GRADIENTS IN IONIC STRENGTH AND DISSOLVED ORGANIC MATTER CONTENT 3.1 Introduction The success of nanotechnologies and the resu lting widespread production and use of m anufactured nanomaterials (MNs) will likely le ad to their introduction to natural systems. Rivers/lakes are likely to behave as primary sinks as they inte grate pollutants from atmospheric deposition, terrestrial surface runoffs, and groundwate r discharges. In the past few years, several studies focused on the potential t oxicity of MNs, using existing experimental procedures, which are not always adequate to asse ss the environmental implications of these new pollutants. If current efforts to understand th e biological effects of MNs on human health have relied on appropriate techniques such as inhalation exposures, studie s emphasizing the environmental implications use primarily very dr astic 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) c ould release nano-Ag pa rticles into sewer systems (Christen 2007; Lovern et al. 2007). Since current wastewat er treatment plants are illequipped for MNs removal, these MNs would ultimately enter natural waterways. So far, laboratory studies on toxicity of MNs have fo cused on their impacts on model organisms (Du et al. 2008; Fortner et al. 2005; Gr iffitt 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 obta ined under conditions that are far different from


56 those encountered in natural systems. For ex ample, the use of drastic mixing methods (e.g. ultrasound, sonication, etc), toxic solvents/surfact ants (e.g. toluene, sodium dodecyl sulfate, triton X-100, etc) or a combination the above te chniques in the prepara tion of MNs suspensions lead to ideal MN-dispersion levels, which ar e unfortunately unattainable through direct introduction of MN into natural waters. Accord ingly, 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 imp acts of selected MNs that previously showed toxicity effects (i.e. C60, nano-copper and nano-silver), when suspended directly in natural waters. The ra tionale for this study is that MNs released di rectly to waterways would likely be impacted primarily by solutio n composition and weak mechanical dispersion processes such as waves and biot urbation. 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 Materials and Methods 3.2.1 Collection of Water Samples Water sam ples used in this study were coll ected 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 hyd rologic units (upper, middle and estuary), each providing distinct hydrological char acteristics and gradients in dissolved organic carbon (DOC) and ionic streng th (I). In the upper wa tershed, confinement of the Floridan aquifer provides surface drainage and s ources of organic carbon from wetlands (e.g. the Okefenokee swamp in southern Georgia, Fig. 3-1). The boundary betw een the upper confined and middle unconfined watershed is a geomorph ic feature called the Cody Scarp, below which


57 ground water returns to the surface from many larg e 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 h eadwaters 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 c oolers and transported back to UF campus. In the laboratory, samples were filtered (0.45 m) to remove large size debris and kept refrigerated at 4C 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 ge ntle mixing period, MNs were filtered to remove agglomerated particles with size > 1.6 m (1.6 m 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 tetrahydrofuran e (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-sus pensions 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. 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 Re search (Alachua, FL) and kept in 1L beakers containing 500mL of MHW in a PervicalTM model # E-30 BX environmental chamber at 25 C with constant aeration. Th e 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 da y 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 accordi ng to EPAs protocol (USEPA 2002), in which MHW serves as diluent and as negative control. MNs tested for toxici ty were added to the culture as natural water suspen sions to produce an increasing c oncentration 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 cont aining 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 i mmobilization, and for each tested MN, at least five concentrations were used for LC50 determination (i.e. conc entration resulting to the death/immobilization of 50% of the population). All experiments we re 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 consider ed different when not overlapping. MetPLATE test The MetPL ATE 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 -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 l aliquot of the MN-suspensions prepar ed in natural waters was added to a test tube containing 100 l of the above-described bacteria l 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 C. A 200l aliquot of the above mixture was then transferred to a 96-well microplate to which 100 l of CPRG, the enzyme substrate, was added. The microplates were then shaken and incubate d at 35 C 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 samp le concentrations (X-axi s) versus the percent inhibition (%I) determined using Eq. 3-1 (Y-axis) and then performing a regression analysis in the linear portion of the obt ained line, from which the IC50 was determined using Eq. 3-2. 100 1 control sampleA A % I (3-1) S Y ICercept int 5050 (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 Results and Discussion 3.3.1 Characterization of Water Samples The chem ical composition of the three water samples is presented in Table 3-1. Sample SR1 was characterized by a very high DOC conten t (47.71 mg C/L) and lo w 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 c oncentration of SR3 samples could not be determined in our laboratory due to the high chlo ride 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 sa mples shows a steep gradient indicative of the downstream in increasing concentr ation of inorganic carbon, pr imarily 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 g/L) of the ICPAES 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 m 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 ~12.5 8 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 di fference 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 partiall y oxidized when exposed to oxygen and Ag+ will form on the particles surface (Lok et al. 2007). The posit ive 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. 32). 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-Marzn 2002; Zhou et al. 2006). Accordingl y, steric stabilization effect from the organic ligand layers coating the nanopa rticles can help prevent particle aggregation. Therefore, in SR2, removal of Ag is most likel y due to low DOC levels. These results suggest that organic matter and solution chemical co mposition can influence the suspension of nanosilver 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 nanoAg, soluble Cu tends to form very stable complexes with natural or ganic ligands, e.g. log K'Cu(II)L varies from 10 to 14 (Leal and Van den


62 Berg 1998; Pastoriza-Santos a nd Liz-Marzn 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 part icles. 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 solubili ty in water (Heymann 1996; Ruoff et al. 1993), an ap proximate 3.09 ppm of C60 nanoparticles was suspended in SR2. This is probably because C60 particles could be negatively charged ( up to -40 mV) in aqueous systems (Brant et al. 2005; Deguchi et al 2001). However, when in contac t 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 explanat ion 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) ha s 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. Espi nasse 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 Nano materials Suspended in Natural Waters to Ceriodaphnia dubia As particle-feeding organism s, 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-coppe r were found to be toxic, as reflected by the LC50 data shown in Fig 3-4. The highest toxi city was observed on organisms exposed to nanopure-water suspended nanometals, with a LC50 of 0.462 g/L for Ag (95% CI = 0.449 0.473 g/L) and 2.14 g/L for Cu (95% CI = 1.97 2.33 g/L). With regard to Ag, SR1 had a LC50 of 6.18 g/L (95% CI = 5.52-6.66 g/L), 0.771 g/L (95% CI = 0.74546-0.798 g/L) for SR2, and 0.696 g/L (95% CI = 6.56-0.730 g/L) for SR3. The LC50 values for Cu in SR1 and SR2 had trends similar to that of Ag and with LC50 of 46 g/L for nano-Cu suspended in SR1 water (95% CI = 43-52 g/L) and 7.12 g/L (95% CI = 6.93-7.30 g/L) in SR2 water. Unlike Ag, nano-copper in SR3 was less toxic with a LC50 value of 48 g/L (95% CI = 41-74 g/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 toxi city 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 de spite its much smaller Ag content. This could suggest that Ag+ was likely the dominant form after susp ension of nano-silver SR2 water.


64 Cu can bind to the fish gills, inhibit ion tr ansport, and cause disturbance of multiple physiological processes (Grosell et al. 2007; Schw artz et al. 2004). Organic matter can not only facilitate dissolution and disper sion of copper nanoparticles, but decrease the toxicity of copper as well (Kim et al. 1999; Richar ds et al. 2001; Schwartz et al 2004; Vasconcelos et al. 1997). Similar to other results (Kramer et al. 2004; Pere z 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 approxima tely one or two orders of magnitude, indicating much higher toxicity asso ciated 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; Sc iera et al. 2004). In a separate study, we te sted the toxicity of C60 water suspensions prepared by the Deguchis 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 THFproduced C60 suspensions to that of C60 directly suspended in DI-w ater. They showed a much poorer suspension in DI-water, resulting in lower C60 toxicity. In Zhus study, the LC50 for THFnC60 to Daphnia magna was about 0.8 ppm while water-stirred-nC60 didnt cause 50% mortality at concentrations up to 35 ppm (Zhu et al. 2006). It has been suggested th at the residual THF is responsible for the observed highe r toxicity, but others argued that differences in surface chemistry and/or morphology play a more importa nt role in the observe d 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 fo rms. 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 indica te the presence of bioavailable forms of the m etal (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 g/L and 48 6 g/L, respectively. Similar to the well known strong antimicrobial e ffects 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 surfac e (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 w ith 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-AgDI-water suspensions indicated presence of bioavailable Ag species (B itton et al. 1994). According to the literature, the bioavailabl e 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, impl ying 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 na no-aggregates owing to the particle-feeding


66 behavior of the cladocerans. C ontrary to silver, Cu-NOM comple xes 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 system s varies significantly in different aqueous systems. The suspended concentr ations 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 be havior 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 mani pulated nano-water systems, much still remains to be answered and further investigated.


67 Figure 3-1. Map of the Suwannee River watershe d and tributaries show ing 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 Ag0 0.5 1 1.5 2 2.5 3 SR1 SR2 SR3 DIConcentration (mg/L)0.54 0.043 0.66 1.67 Cu0 5 10 15 20 25 SR1 SR2 SR3 DIConcentration (mg/L)12.58 1.45 0.51 0.134 C600 0.5 1 1.5 2 2.5 3 3.5 SR1 SR2 SR3 DIConcentration (mg/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 a nd then filtered. Numbers above bars are average concentrations of two samples and vertical li nes represent 1 standard deviation of the mean.


69 SR1 SR2 DI -2 0 2 4 6 8 10 12 14 01 02 03 04 05 0 DOC concentration (mg/L) Cu concentration (mg/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 Ag0 1 2 3 4 5 6 7 SR1SR2SR3DILC50 (ug/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 ) Cu0 10 20 30 40 50 60 SR1SR2SR3DILC50 (ug/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) silverand (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 SR2 DI -10 0 10 20 30 40 50 01 02 03 04 05 0 DOC concentration (mg/L)LC50 of 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 Ag0 20 40 60 80 100 120 SR1SR2SR3DIIC50 (ug/L) N A 112.14 N A 47.79 Cu0 50 100 150 200 250 SR1 SR2 SR3 DIIC50 (ug/L) N A 204 46.08 30.91 Figure 3-6. IC50 values of (A) nanosilverand (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 ( g/L) Cu ( g/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 Cas tillo et al. 2000).


74 CHAPTER 4 MOBILITY OF SINGLE-WA LLE D CARBON NANOTUBES (S WNTS) IN SATURATED HETEROGENEOUS POROUS MEDIA 4.1 Introduction Single-walled carbon nanotubes (S WNTs) have attracted the attention of both engineers and scientists since their di scovery in 1993 (Iijima and Ichiha shi 1993). The unique size-related characteristics of SWNTs, such as strength, el asticity, high adsorption cap acity, and controllable conductivity, have led to th eir use in a growing number of industr ial processes as well as to their introduction into a wide variety of commercial product (Isobe et al. 2006; Ji a et al. 2005; Lyon et al. 2005; Templeton et al. 2006). It has been estima ted 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 ra tes of SWNTs and their anticipated uses in a wide variety of commercial and industrial applicatio ns 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 hu man health problems. Besides toxicity issues comm on to nanosized materials, the environmental and human health implications of SWNTs seem to be comp licated 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 pr oduction 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 inhe rent properties (Wang et al. 2004; Zhang et al. 2005). However,


75 some of the surfactants with efficient SWNT di spersion 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 (L anone 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 combin ation of two or more of the above parameters. Currently, research on the imp lications of SWNTs has been devoted largely to toxicity issues while the investig ation of their fate and transport in both aquatic and terrestrial systems remains limited. Changes in nanopar ticle surface chemistry in comb ination 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; Leco anet 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 zerova lent iron nanoparticles by electro static repulsion of similarly charged particles. Implicitly, the lack of aggregation would then favor the transport of nanosized particles through porous media. Cu rrently published research on transport of nanomaterials in porous media are so far limited to qualitative desc riptions of transport be havior (Heller et al. 2004; Lecoanet et al. 2004), and modeling st udies 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 da ta that can be used to predict the ability of


76 SWNTs to move through heterogeneous porous medi a 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 colu mns 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 pred ict the transport of nanosize particles such as viruses (20 200 nm) and colloids (<10 m) in soils (Chendorain et al. 1998; Close et al. 2006; Jin et al. 2000; VidalesContreras et al. 2006). 4.2 Materials and Methods 4.2.1 Single-Walled Carbon Nanotube Sample Preparation SW NTs were obtained from Rice University (Rice HPR 145.1). SWNT suspensions were prepared by suspending an initia l 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 fr om 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 centrif uged at 20,000 rpm (Beckman Coulter Optima L80 K) for 2.5 h. The supernatant was then carefu lly separated from the aggregated nanotubes collected at the bottom of the tube. In para llel, an aqueous suspen sion of SWNT without surfactant was prepared by direct mixing of SWNT and Nanopure water followed by a 15minute sonication period. For all of these produced suspensions, visible absorbance spectra of SWNTs were recorded by using a NanoFluorescence Nanospectrolyzer (Applied NanoFluorescence, LLC), and the concentratio ns 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 W oods on the University of Florida campus (sandy soil) and from a construction site n ear 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). Partic le size distribution was determined according to the USDA Soil Survey Lab Method (USDA 1992). The Wa lkley & Black Method (Walkley and Black 1934) was used to measure the soil organic carb on 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-dri ed, sieved (1.19mm) to remove gross debris and then homogenized. 4.2.3 Column Experiments Colum ns used in this study were 30 cm long cl ear 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 th e leachate through. For sandy soil experiments, each column was f illed 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 conser vative tracer solution, bromide (Br-) in this case, and SWNTs suspension into the columns, packed soils were first saturated with nanopure water (>18 M cm, Barnstead Corp) to enhance packing homogeneity and help clean the effluent from loose soil fine particles. Next, 50 ml of Brsolution, 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 inte rvals and analyzed for Brand SWNT concentrati ons. In case of clay soil, a similar procedure was used, except that 15 0 g of soil were packed into the column and 25 ml of Brsolution, and 25 ml of SWNT-GA or SW NT-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 Nanospectroly zer. 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 trans port in soils was selected and used in this study to simulate SW NT 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): c x c v x c D t c R 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 the removal rate coefficient (min-1). For linear instantaneous sorption, the above retardation factor R is defined as: dbK R 1 (4-2)


79 where b is the bulk density of sa ndy soils (dry weight) (g cm-3), 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 pr ogram CXTFIT 2.1, was used to estimate transport parameters with experimentally collec ted breakthrough curve (BTC) data (Toride et al. 1999). The bromide data were analy zed by setting the removal rate ( ) as zero and the retardation factor (R ) as 1. 4.3 Results and Discussion 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 42. Table 4-2 lists the parameters used in the model simulations. As a conservative tracer, Bris assumed to have no retardation (R = 1) or reaction ( = 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 Brtransport very well (with correlation coefficients ( r2) equal to 0.98 and 0.85 for sandy soils and clay soils, resp ectively 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 m odel, the pore water velocity was measured experimentally. Dispersion on the other hand is ma inly 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, th e dispersion coefficient


80 of SWNT was assumed to be similar to that of Br-, and the reaction rate 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, ex tensive 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; Funde rburg et al. 1981; Han et al. 2006; Jin and Yates 2002; Jin et al. 1997). Given the nanosize of these biotic par ticles, results on viruses can be used for comparison purposes. For instance, one of the majo r 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 adsorp tion on soil particles (Schijven and Hassanizadeh 2000). The CDE model, which has b een successfully used to predic t 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 BrBTC (Figure 4-3), indicating the likely occurrence of kinetic adso rption 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 soil s 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 resu lts 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 surf ace tension of particles and increase the capacity of the mobile aqueous phase by forming micelles (Chen et al. 2006). The ma ss balance was 37% and 47% for SWNT-GA and SWNT-SDS, respectively. The mass recovery results suggest ed that more SWNT-GA was retained in the soil column th an 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 of SWNT-GA than SWNTSDS indicate stronger adsorption of SW NT-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 ch arged tannic acid resulted in le ss attachment efficiency than polysaccharide-like organic compou nds. 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 or ganic matter can provide negative charges or cover positively charged sites, thus increasing the repulsion be tween virus and soil particles. Wall et al. (2008) suggested it is due to orga nic matter occupying adsorption sites or physically blocking pores. From these findings it can be co ncluded that surfactant modified surface charges of the nanoparticles could influence the re tention 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 colu mns 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 cl ay 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, th e flow rate of SWNTSDS 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 re tention 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 agr ee 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 typi cal Florida sandy soil, raising concerns that surfactant-dispersed SWNTs could rather easily trav el 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 SW NT on microorganisms could ra ise concerns on the long-term impact of such retained SWNT on soil microor ganisms and associated ecosystem services. Finally, this study also demonstrat ed 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 por ous media, and ideally, such studies should investigate changes in physicochemical charact eristics of MNs followi ng transport through soil


83 columns, as well as the potential toxicity of elut ed MNs and/or their deri vatives. It appears that under specific conditions, the transp ort behavior of SWNT can be m odeled, 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 experime ntal 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 0.51 1.5 2 2.53Pore volumeC/C 0 Experiment data Simulated data Bromide-0.1 0 0.1 0.2 0.3 0.4 0.5 0.6 00.511.522.53Pore volumeC/C 0 Experimental data Simulated data Figure 4-2. Breakthrough curves of experime ntal 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 012345Pore volumeC/C 0 Experiment data Simulated data SWNT in SDS-0.1 0 0.1 0.2 0.3 0.4 0.5 00.511.522.53Pore volumeC/C 0 Experiment data Simulated data Figure 4-3. Breakthrough curves of experiment al and simulated data of SWNT-GA and SWNTSDS suspensions in sandy soil columns. Vertic al 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 co lumn 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) (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 MANU FAC TURED 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 fl ow of energy, material and information, and the participation of biota and wate r. Therefore, the ability to qua litatively and/or quantitatively characterize any of the above lis ted natural processes can be us ed 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 ba sic properties of ecosystem s (e.g. flow of matter and energy). In sediments, the composition and distributi on of microbial populations are usually wellestablished, although changes associ ated with shifts in seasons and other major parameters are also common. However, the input of pollutants can have significant im pacts on the composition of microbial communities and/or their activi ties in sediments. In such cases, potential consequences could range from a simple delay in biodegradation of or ganic matter to major environmental impacts such as the production of more toxic derivatives, with bioaccumulation potential and negative effect s 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 basi c ecosystem functions. This is because the anticipated widespread production and use of MN s could lead to new environmental pollutants (Masciangioli and Zhang 2003), which could either directly impact living cells, or undergo environmental transformations to pr oduce 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 t oxicity response of test model organisms and human cells to MNs exposure, onl y limited number of studies have investigated the fate, transport and the resulting impact s of MNs at the system level (H yung et al. 2007; Lecoanet et al. 2004; Lecoanet and Wiesner 2004). They showed that MNs could exhibit different transport behaviors in porous media or beco me stabilized in organic rich waters with potential for longrange 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 cytot oxic due to the release of Cd2+ ions (Derfus et al. 2004; Lanone and Boczkowski 2006; Zhang et al. 2006) an d/or through direct interaction of the quantum dots w ith 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 co nducted in our laboratory based on kinetics of acetate biodegradation showed that the addition of C60 to sediment slurries negatively impacted rates of acetate oxidation (Kopelevi ch et al. 2008). In contrast, T ong 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 (nit rate and sulfate) and the eff ect of tested MNs assessed. 5.2 Materials and Methods 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 ultrahigh purity (UHP) nitrogen for 1 hour to remove oxygen, and then sealed and left stirring at room temperature for 24 hours. Excess solid s were later filtered out using a 0.45 m PTFE membrane filter, resulting in a transparent pink solution. The filtrate was then added to equal amount of water and placed in a wa ter bath prior to purging with UHP-N2 to evaoparte THF. Following the THF evaporation, the obtaine d solution was vacuum-filtered through 0.45 m 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 purch ased from NN-Labs, LLC (Fayetteville, AR). 5.2.2 Sediment Collection Sediments used in this study were collected from a polluted lake (Lak e Alice) located on University of Florida campus in Gainesville, Florida. Surface sediments (top 10 cm) were collected using pre-cleaned high density polyethylene scoops, transfer red 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 redo x 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 syst ems 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 equilibriu m, and that platinum el ectrodes 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 processe s 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: methanoge nesis: 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 indicat or 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 predom inance of sulfate reduction in samples that lacked significant sulfate c oncentrations (Chapelle et al. 1996) Also, its occurrence at trace levels requires highly sensitive and expensive an alytical techniques. Accordingly, for studies conducted in batch reactors, one of the reliab le and inexpensive approaches used electron-


93 acceptor availability and the presen ce of final products of microbial metabolism as illustrated in half-reactions (equations 5-1 through 5-8). OHeHaqO2 22 1 )( 4 1 (5-1) OHNOeHNO2 2 32 1 2 1 2 1 (5-2) OHgONeHNO2 2 38 5 )( 8 1 4 5 4 1 (5-3) OHgNeHNO2 2 35 3 )( 10 1 5 6 5 1 (5-4) OHMneHSMnO2 2 22 2 1 2)( 2 1 (5-5) )aq( )s(FeeFe2 3 (5-6) OHHSeHSO2 2 42 1 8 1 8 9 8 1 (5-7) OHgCHeHgCO2 4 24 1 )( 8 1 )( 8 1 (5-8) The occurrence of the above geochemical r eactions is thermodynamically driven, as reactions with the lowest Gibbs free energy are favored when several TEAs are present in the system. Accordingly, simple meas urements 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 potentia l impacts, if any, of MNs on rates of these microbial-catalyzed geochemical reactions. This study focused primarily on sediment micr oorganisms that utilize nitrate (equations 5-2, 5-3, and 5-4) or sulfate (e quation 5-7) as TEAs. In the labor atory, and prior to MNs testing,


94 relatively diluted sediment slurries (1:5 m/v) were prepared under continuo us 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 aliquo ts from slurries showed that nitrate was mostly converted to nitrite and sulf ate concentrations averaged ~24 mg/L. At this point, aliquots of well-ho mogenized 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 CH3COOper liter of slurry. In parallel, sediment slurries with final concentrations of 150 ppm of CH3COOand 100 ppm of either 3NOor 2 4SOwere also prepared to re-establish the prevalence of these specific TEAs and favor biogeochemical r eactions catalyzed by either nitrateor sulfatereducing bacteria. Hermetically sealed vials were placed in the dark and the microbial degradation of acetate monitored over time in bot h 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, respectivel y. 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 m syringe-filters in the anaerobic chamber. The filtrates were then analyzed forCOOCH3, 3NO 2NOand 2 4SO 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 i ndigenous sediments microorganisms. All experiments were run in triplicates.


95 5.2.4 Analytical Techniques The concentrations ofCOOCH3, 3NO, 2NOand 2 4SO 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 to luene 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 meas urement at 334 nm using a Hach DR/4000U Spectrophotometer. 5.2.5 Data Analysis All experiments were run in triplicates and obt ained data presented as mean 1 standard deviation. Differences amongst the treatment eff ects were analyzed using one-way analysis of variance (ANOVA) followed by Tukeys test for multiple comparisons of anion concentrations. Differences were considered significant when p values were < 0.05. 5.3 Results and Discussion In these experiments, nitrate and sulfate were added to sediment slurries to establish predominant TEAPs, and then asse ss 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, w ithout 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 natura lly occurring in the sedi ment slurries prior to sample incubations. Figures 5-2 and 5-3 are sim ilar 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 decr ease over time, suggesting an inhibition of microbiological pr ocesses leading to the oxidation of organic matter. During the 17days of incubation, trends in ni trate (Fig. 5-1b) and nitrite (Fig. 5-1c) in these sediment slurries paralleled that of acetate, suggesting that bot h 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 fa ster (Figs. 5-2a and 54b). 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 CdSeand nano-Ag spiked samples paralleled trends observed in control samples. However, some tre nds 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 observe d 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 C60treated samples) or decreasing c oncentrations (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 reducer s 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 un dergo 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 chan ge in sulfate concentrations (Fig. 5-3d) shows that sulfate reducing bacteria (SRB) were not in volved in any of the acetate degradation observed in these experiments (Fig. 5-3a). To efficiently test the impacts on natura lly occurring sediment SRB, it would have been necessary to deplete th e tested sediment slurries of their original background sulfate through a longterm anaerobic incubation in or der to expect a significant microbiological response attributab le 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 disapp earance 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 se vere microbial inhibition by the C60-suspensions. In our previous studies, we observed only impeded bi odegradation of acetate even with a higher nC60 concentration of 0.5 ppm (Kopelevi ch 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 sma ller in CdSe and nanosilver spiked slurries than the control for acet ate 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, th ese results point to the potential of C60 to inhibit the microbiologically driven oxidation of organic matte r. 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 sediment ary biogeochemical processes were studied. Fullerene aqueous suspensions at a concentratio n 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 coul d be an indication that higher concentrations of thes e MNs would likely result in negative effects on sediment microorganisms as well. Future research should combine the use of molecular probe and welldefined redox conditions to pinpoi nt specific microbial groups and assess their responses when they become exposed to increasing MNs concentrations.


99 Figure 5-1. Kinetics of acetate de gradation (a), and nitrate (b), nitrite (c) and sulfate (d) concentrations in sediment slurries w ithout (controls) or spiked with tested nanomaterials (C60, nanosilver, and CdSe quantum dot s). Vertical bars represent 1 standard deviation of the mean of three replicates.

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100 Figure 5-2. Kinetics of acetate de gradation (a), and nitrate (b), nitrite (c) and sulfate (d) 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.

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101 Figure 5-3. Kinetics of acetate de gradation (a), and nitrate (b), nitrite (c) and sulfate (d) 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.

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102 Figure 5-4. Pseudo-first order ki netics of acetate disappearance fr om sediment-slurries treated with either silver nanoparticles or CdSe quantum dots as compared to the non-treated controls. Ln(C) represents the natural l og of acetate concentration in mg/L. (a) = slurries containing acetate only, (b) = slurries with both acetate and nitrate additions, and (c) = slurries with ace tate and sulfate additions.

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103 CHAPTER 6 NANOWASTES IN THE ENVIRONMENT: T HE TROJAN HORSE EFFECT OF NANOMATERIALS 6.1 Introduction The environmental fate and transport of manuf actured nanomaterials (MNs) as well as the fate of pollutants sorbed ont o 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 gr owth and development for the early 21st century. Most scientists and engineers are confident that nanoscience and nanotechnology will revolutionize medicinal, industrial, agricultura l, and environmental research as a wide variety of MNs are being produced (Hurt et al. 2006). Although in its infancy, resear ch on both the environmental impacts and health implications of MNs is fa st 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 pollutant s 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 effl uents 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 pollu tants (Daughton 2004). Unfortunately a major disadvantage of this late corrective approach is the difficulty to deal with we ll-established economic activities that generate the pollutants of concern. Theref ore, a proactive approach is ideal to limit the complex ramifications associated with delaye d prevention and remediation measures (Daughton 2004). We assessed the fate and potential imp acts of Hg sorbed onto MNs (i.e. SiO2-TiO2 nanocomposites) by mimicking the Trojan horse effect in a sediment matrix. Spent SiO2-TiO2

PAGE 104

104 nanocomposites used in the removal of Hg from a simulated coal combustion effluent were used in laboratory studies to determine the bioa vailability of inorgani c Hg sorbed onto SiO2-TiO2 nanocomposites by using Hg methylation as a proxy for bioavailability, and the toxicity of HgSiO2-TiO2 complexes using FluoroMetPLATE, a bacterial based microbiotest. 6.2 Materials and Methods A thorough description of the proc edure 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 meas urement 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 be tween 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 me thylating bacteria. Hg methyla tion 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 solu tion was used to produce slurries with pH>4. All prepared sediment slurries were then de-aerated with ultr a high purity (UHP) N2 to help accelerate the development of anoxic conditions a nd favor methyl-Hg production. Overall, the experiment consisted of control slurries with no Hg addition and slurries spiked with Hg as HgSiO2-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, totaland methyl-Hg concentrations in aqueous phase and sediments were determined following previously published

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105 methods (Bloom 1989; Bloom and Crecelius 19 83; 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 (SPL P) 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 SiO2TiO2 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 follo wing digestion with bromine monochloride (Bloom and Crecelius 1983). 6.3 Results and Discussion 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 Hgnanocomposite complexes. The percentage of THg methylated from the Hg-SiO2-TiO2complexes added to slurries and corrected from control samples is shown in Figure 6-1 as a function of pH. These results show an incr easing and pH-dependent trend of methyl-Hg production over time, with more methyl-Hg produced at the lo west tested pH. This trend suggests that microbial Hg methyl ation could be controlled by the solubility of adsorbed Hg onto MNs, which decreases with increasi ng 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 methylHg under these tested conditions, regardless of the amount produced, is a strong indication of the bioavailability of Hg-sorbed onto SiO2-TiO2 nanocomposites.

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106 Using the linear portion of the above describe d Hg methylation trends, and assuming that the methylation rates of inorgani c 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 cons tant rate approached zero at pH 6 as methyl-Hg levels in sediment slurries becam e barely detectable. Additionally, the rate determined at pH 4 was about one order of magn itude 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 bioacce ssibility. Overall, Hg adsorbed onto MN and introduced into sedimentary environments could qui ckly become bioavailab le 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 concentrati ons on plain materials we re at or below our analytical detection limits. The SPLP extracti on procedure was comparatively applied to SiO2 and TiO2 nanoparticles, ultra-violet (UV) irradiated virgin SiO2-TiO2 nanocomposites, and the Hg-contaminated UV-irradiated nanocomposites (P itoniak 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 i nhibition 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 Hgcontaminated nanocomposites show a much higher toxicity with average inhibition values of

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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 inhibi tion recorded with the Hg-contaminated SiO2-TiO2 nanocomposites indicates an additional toxicity eff ect 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 b acteria involved in Hg biotransformation become impacted. Consequently, the detection of methyl-Hg in these methylation expe riments 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 e nvironmental fate of adsorbed pollutants as they ultimately enter natural (e.g., waterways) or engin eered (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 n ear total inhibition of Hg methylation at pH6. Overall, these results point to the potential for bio accessbility of MNadsorbed pollutants as a function of certain key environmental pa rameters 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.

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108 0 2 4 6 8 10 12 14 16 02468 Time (days)% THg converted to methyl-H g pH 4 pH 5 pH 6 Figure 6-1. Percent THg convert ed 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 analyti cal detection limit of 0.05 ng/g.

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109 pH 6 pH 5 k = 0.0041/day r2=0.9755 pH 4 k = 0.02/day r2=0.9328 0.00 0.05 0.10 0.15 0.20 0.25 024681 0 Time (days)-Ln(1-([MeHg]/THg)) Figure 6-2. Kinetics of Hg methylation in sediment slurries spiked with SiO2-TiO2-Hg complexes at pH 4 (triangles ; native pH), 5 (circles), a nd 6 (squares). The methylation of inorganic Hg decreases with increasing pH to reach values below the analytical detection limit at pH 6.

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110 0 10 20 30 40 50 60 70 80 90 100 SiO2 nanoparticleTiO2 nanoparticleUV-irradiated non-used TiO2SiO2 nanocomposites Hg-TiO2-SiO2 complexes% Inhibitio n 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 pe rcent (%) inhibition.Not Detected

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111 CHAPTER 7 CONCLUSIONS AND RECOMMENDATIONS 7.1 Conclusions Nanotechnology is a highly promising and ex citing 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 incr easingly 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 bi otic and abiotic transf ormations of MNs in natural systems are limited. This research focused on potential environmental impacts of MN s, and the use of toxicological, hydrological a nd biogeochemical approaches has resulted in the following conclusions. Toxicity testing was based on three different microbiotests emphasizing the interactions of MNs with (i) biochemical proc esses (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 concentrati on and testing methods. The use of different toxicity methods is n ecessary to avoid erroneous results. This is because different test organisms respond diff erently 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 surf actant used in this study) showed higher degrees of toxicity as compared to water sonicated suspensions. Although the toxicity

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112 mechanisms were not addressed in this study, the toxicity of these MNs could be attributable to their ability to generate hi ghly reactive and toxic fr ee radicals, their degree of purity as toxic impurities increase toxicity, or simple surface interac tion 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 th at toxicity results obtained from laboratory experiments that use drastic MNs suspension methods may not be realistic. It was found that th e 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/ch aracteristics and solution chem istry (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 ad sorption and dispersion in the porous media. Finally, the use of a convection-dispersion m odel 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 ba cteria involved in organic matter oxidation, primarily nitrate and nitrite reducer s. Nano-silver and CdSe quantum dots were less toxic at tested concentrations, but gave the indication of potenti al 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 a dvantage with regard to medical research, it could have negative implications in the environment. 7.2 Recommendations Based on our findings, the following recommenda tions are made to fu rther the extent of knowledge of the environmenta l 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 test s 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.

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113 The use of mathematical models to explor e the transport patterns of MNs in different types of soils should be considered as it can he lp predict the potential dispersal of MNs in porous media. Although our study observed negative impact s of MNs on sediment microorganisms, further investigations are needed to pinpoint the microbial groups that are sensitive to MN toxicity. Molecular techniques could reveal th e 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 partic ulate binding properties of MNs in both freshwater and seawater. The relative importance of differen t biological routes of uptake al so needs to be assessed in representative aquatic species, since this will be a crucial factor governing intracellular behavior, distribution, fate and toxic ity of internalized MNs. A major challenge remains the derivation of t oxicity 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 w ill 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 functio nal properties.

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114 APPENDIX A TESTED CONCENTRATIONS OF CARBONAND MET AL-BASED NANOMATERIALS IN THREE DIFFERENT TOXICITY ASSAYS Table A-1. Tested concentrati ons of carbonand metal-based na nomaterials 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,

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115 APPENDIX B TESTED CONCENTRATIONS OF MANUFA CTURED NANOMAT ERIAL SUSPENSIONS IN TOXICITY TESTS USING THE C. DAPHNIA 48-H ACUTE TOXICITY ASSAY AND METPLATE TEST Table B-1. Tested concentrati ons of manufactured nanomaterial suspensions in toxicity tests using the C. daphnia 48-h acute toxicity as say and MetPLATE test Nanomaterial Suspensions Concentrations Used in 48-h Ceriodaphnia dubia Assay (mg/L) Concentrations Used in MetPLATE Test ( g/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

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135 BIOGRAPHICAL SKETCH Jie Gao was born in Huaian, Jiangsu Province, C hina. She received her bachelors degree in environmental science from Nanjing Univer sity in 2003. After one year study in Beijing University graduate school, Jie Gao was admitte d as a PhD student with an Alumni Fellowship award in the Department of Environmental Engineering Sciences at the Un iversity of Florida. She received Masters degree in 2005 and has been ever since wo rking on her doctoral research in assessment of environmental impacts of na nomaterials under the supervisory of Dr. JeanClaude Bonzongo.