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1 ON THE TRANSFORMATION AND BIO INTERACTION OF NANOSILVER PARTICLES IN NATURAL WATERS: TOXICITY IMPLICATIONS FOR AQUATIC ORGANISMS By JULIANNE MCLAUGHLIN A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 2013
2 2013 Julianne McLaughlin
3 To Agnes Limehouse McElheny
4 ACKNOWLEDGMENTS I thank my advisor, Jean Claude Bonzongo for his dedication and wealth of knowledge. I would like to thank my committee members: Joseph Delfino, Mark Brown, and Kevin Powers for their support, guidance, and their willingness to help throughout my PhD exper ience. I am truly indebted to my boyfriend, Jose Antonio Yaquian, who is my rock and to my brother Daniel McLaughlin, who is the only reason I made it through this process. Finally to my family, Mom, Dad, and b rothers; who have been amazing supports throu ghout my life.
5 TABLE OF CONTENTS page ACKNOWLEDGM ENTS ................................ ................................ ................................ .. 4 LIST OF TABLES ................................ ................................ ................................ ............ 8 LIST OF FIGURES ................................ ................................ ................................ .......... 9 LIST OF ABBREVIATIONS ................................ ................................ ........................... 11 ABSTRACT ................................ ................................ ................................ ................... 13 CHAPTER 1 NANOTECHNOLOGY AND THE ENVIRONMENT: APPLICATIONS AND IMPLICATIONS ................................ ................................ ................................ ...... 15 1.1 Problem Statem ent ................................ ................................ ........................... 15 1.2 Production and Applications of NPs: General Overview ................................ ... 16 1.2.1 Buckminsterfullerene (C 60 ) ................................ ................................ ...... 16 1.2.2 Carbon Nanotubes (CNT) ................................ ................................ ........ 17 1 2 3 Metal Oxide Nanoparticles ................................ ................................ ...... 18 1.2. 4 Nanometals ................................ ................................ ............................. 19 1.2. 5 Quantum dots (QDs) ................................ ................................ ............... 20 1.3 Environmental Implications of Manufactured Nanoparticles: General Overview ................................ ................................ ................................ .............. 21 1.3.1 Toxicity of C 60 and CNT ................................ ................................ ........... 21 1.3.2 Toxicity of Nanometal oxides ................................ ................................ ... 22 1.3.3 Toxicity of Nanometals ................................ ................................ ............ 23 1.3.4 Toxicity of Quantum Dots ................................ ................................ ........ 23 1.4 Rationale for Selection of Nanosilver ................................ ................................ 24 1.5 Scope of Research ................................ ................................ ........................... 26 2 NANOSILVER AND THE ENVIRONMENT: CURRENT STATE OF KNOWLEDGE ................................ ................................ ................................ ........ 29 2.1 Applications and Current Uses of Nanosilver ................................ .................... 29 2.2 Synthesis Methods for Nanosilver ................................ ................................ .... 30 2.3 Toxicity of Nanosilver to Aquatic Organisms ................................ ..................... 32 2.3.1 Bacteria ................................ ................................ ................................ ... 32 2.3.2 Algae ................................ ................................ ................................ ....... 34 2.3.3 Invertebrates ................................ ................................ ........................... 36 2.4. Nanosilver Interactions in the Environment ................................ ...................... 38 2.4.1 Mechanisms of Nanosilver Dissolution ................................ .................... 38 2.4.2 Effects of Organic Matter ................................ ................................ ......... 41 2.4.3 Effects of Inorganic Ligands, Ionic Strength, and pH ............................... 41
6 2.5 Considerations and Suggestions ................................ ................................ ...... 44 3 TRANSFORMATION AND FATE OF NANOSILVER PARTICLES IN WATERS OF DIFFERENT CHEMICAL COMPOSITION ................................ ........................ 46 3.1. Introductio n ................................ ................................ ................................ ...... 46 3.2. Materials and Methods ................................ ................................ ..................... 47 3.2.1. Collection and Characterization of the Different Waters Used in this Study ................................ ................................ ................................ ............. 47 220.127.116.11. Sample collection and handling ................................ .................... 47 18.104.22.168. Sample analysis ................................ ................................ ............ 48 3.2.2. Preparation and Characterization of nAg Stock Suspensions ................ 49 3.2.3. Imaging of the differ ent nAg suspensions ................................ ............... 50 3.2.4. Dissolution Kinetics ................................ ................................ ................ 51 3.3. Results and Discussion ................................ ................................ .................... 52 3.3.2. Characterization of Prepared nAg Suspensions ................................ ..... 53 22.214.171.124. Initial particle characterization ................................ ....................... 53 126.96.36.199 Characterization of nAg suspensions ................................ ............. 53 188.8.131.52. Effect of storage on nAg particle surface charge and PSD ........... 54 3.3.3. Dissolution Kinetics of nAg Suspended in ACT, SPG, and NP Waters .. 57 3.4. Conclusion ................................ ................................ ................................ ....... 59 4 EFFECTS OF NATURAL WATER CHEMISTRY ON NANOSILVER BEHAVIOR AND TOXICITY TO CERIODAPHNIA DUBIA AND PSEUDOKIRCHNERIELLA SUBCAPITATA ................................ ................................ ................................ ....... 75 4.1 Introduction ................................ ................................ ................................ ....... 75 4.2 Materials and Methods ................................ ................................ ...................... 77 4.2.1 Collection and Characterization of Test Waters ................................ ...... 77 4.2.2 Preparation and Characterization of nAg Stock Suspensions ................. 78 4.2.3 Microscopy ................................ ................................ .............................. 79 4.2.4 Toxicity Assays ................................ ................................ ........................ 80 184.108.40.206 Ceriodaphnia dubia bioassays ................................ ....................... 80 220.127.116.11 Pseudokirchneriella subcapitata bioassays ................................ .... 81 4.3 Results and Discussion ................................ ................................ ..................... 82 4.3.1 Characterization of Test Waters ................................ .............................. 82 4.3.2 Characterization of Prepared nAg Suspensions ................................ ...... 84 18.104.22.168 Initial particle characterization ................................ ........................ 84 22.214.171.124 Characterization of nAg suspensions ................................ ............. 84 4.3.3 Biological Impacts of nAg ................................ ................................ ........ 86 4.4 Conclusion ................................ ................................ ................................ ........ 88 5 INTERACTION OF WATER TRANSFORMED NANOSILVER PARTICLES WITH AQUATIC ORGANISMS: LINKING SOLUTION CHEMISTRY TO OBSERVED BIOLOGICAL RESPONSES ................................ .............................. 97 5.1 Introduction ................................ ................................ ................................ ....... 97
7 5.2 Materials and Methods ................................ ................................ ...................... 98 5.2.1 Collection and Characterization of Test Waters ................................ ...... 98 126.96.36.199 Sample collection and handling ................................ ..................... 98 188.8.131.52. Sample analysis ................................ ................................ ............ 99 5.2.2 Preparation and Characterization of nAg and AgNO 3 Stock Suspensions ................................ ................................ ................................ 100 5.2.3 Toxicity Assays ................................ ................................ ...................... 101 184.108.40.206 Ceriodaphnia dubia bioassays ................................ ..................... 102 220.127.116.11 Pseudokirchneriella subcapitata bioassays ................................ .. 103 5.3 Results and Discussion ................................ ................................ ................... 105 5.3.1 Characterization of Test Waters ................................ ............................ 105 5.3.2. Characterization of nAg Suspensions Used in Toxicity Assays ............ 107 5.3.3 Biological Impacts of nAg and AgNO 3 ................................ ................... 108 18.104.22.168 Exposure to silver and biological responses of organisms grown in MHW growth medium and in the two natural waters ......................... 108 22.214.171.124 Effects of model organic compounds on nAg and AgNO 3 toxicity 112 5.4 Conclusion ................................ ................................ ................................ ...... 115 6 CONCLUSIONS AND RECOMMENDATIONS ................................ ..................... 133 APPENDIX A CHEMICAL COMPOSITION OF MODERATELY HARD WATER ........................ 139 B CHEMICAL COMPOSITION OF THE ALGAL CULTURE MEDIUM ..................... 140 LIST OF REFERENCES ................................ ................................ ............................. 141 BIOGRAPHICAL SKETCH ................................ ................................ .......................... 159
8 LIST OF TABLES Table page 3 1 Temporal chemical characterization of natural water samples ........................... 61 3 2 Particle characterization of nanosilver (nAg) suspensions ................................ 62 3 3 Kinetic of dissolution of nAg particles and production of ionic Ag (Ag + ). ............. 63 4 1 Characterization of natural waters. ................................ ................................ ..... 90 4 2 Characterization of nAg suspended in different waters. ................................ ..... 90 A 1 Chemical composition of moderately hard water (MHW)) ................................ 139 B 1 Chemical composition of the algal culture medium (CM) ................................ 140
9 LIST OF FIGURES Figure page 1 1 Temporal trends of production of engineered nanomaterials ............................ 28 3 1 Excitation emission fluorescence spectra of water samples (Time 1) ............... 64 3 2 Excitation emission fluorescence spectra of water samples (Time 2) ............... 65 3 3 Physical analysis of raw nanosilver (nAg)). ................................ ........................ 66 3 4 Particle size distribution of nAg particles ................................ ............................ 67 3 5 Scanning electron microscopy images of nAg particles ................................ ..... 68 3 6 Zeta potential of stock nAg suspensions. ................................ .......................... 69 3 7 Particle size distributions of nAg suspension in ACT water ............................... 70 3 8 Particle size distribution of nAg suspension in SPG water ................................ 71 3 9 Particle size distribution of nAg suspension in NP water ................................ ... 72 3 10 Mean particle size of nAg over time ................................ ................................ ... 73 3 11 Temporal trends of ionic Ag (Ag+) concentration. ................................ ............. 74 4 1 Excitation emission fluorescence spectra of natural waters. ............................. 91 4 2 Characterization of raw nanosilver (nAg).. ................................ ......................... 92 4 3 Particle size distributions of nAg in tested waters ................................ .............. 93 4 4 Scanning Electron Microscopy images of nAg. ................................ .................. 94 4 5 Dose response curves for the toxicity of nAg particles to C. dubia ..................... 95 4 6 Dose response curves for the toxicity of nAg particles to P. subcapitata .......... 96 5 1 Excitation emission fluorescence spectra of ACT water samples. ................... 116 5 2 Excitation emission fluorescence spectra of SPG water samples ................... 117 5 3 Excitation emission fluorescence spectra of model organic compounds .......... 118 5 4 Dose response curves for the toxicity of Ag to C.dubia in MHW ....................... 119 5 5 Dose response curves for the toxicity of Ag to C. dubia in SPG water ............. 120
10 5 6 Dose response curves for the toxicity of Ag to C. dubia in ACT water .............. 121 5 7 ................................ ............ 122 5 8 Dose response curves for the toxicity to C. dubia exposed to aged nAg ......... 123 5 9 Dose response curves for the toxicity of Ag to P. subcapitata in CM ............... 124 5 10 Dose response curves for the toxicity of Ag to P. subcapitata in SPG ............. 125 5 11 Dose response curves for the toxicity of Ag to P. subcapitata in ACT ............. 126 5 12 ................................ ... 127 5 13 Dose response curves for C. dubia exposed to AgNO3 with model organic compounds ................................ ................................ ................................ ...... 128 5 14 Dose response curves for C. dubia exposed to nAg with humic acid. ............. 1 29 5 15 Dose response curves for P. subcapitata exposed to nAg with model organic compounds ................................ ................................ ................................ ...... 130 5 16 Dose response curve for P. subcapitata exposed to AgNO3 with humic acid. 131 5 17 Dissolution of nAg over time in nAg ACT, nAg SPG, and nAg NP. .................. 132 6 1 Summary of experimental approach used ................................ ....................... 137 6 2 Conceptual diagram of toxicity pathways and trends ................................ ........ 138
11 LIST OF ABBREVIATIONS ACT Alachua Conservation Trust freshwater marsh A G Silver A G NO 3 Silver n itrate BD Below detection limit BET Brunauer, Emmett and Teller method CI Confidence interval CM Culture medium CNT Carbon nanotubes CVD Chemical vapor deposition DI Deionized water DIC Dissolved inorganic carbon DLS Dynamic light scattering DLVO Derjaguin, Landau, Verwey and Overbeek Theory DOC Dissolved organic carbon EDS Energy dispersive spectroscopy EEM Excitation emission fluorescence ICP AES Inductively coupled plasma atomic emission spectroscopy MHW Moderately hard water MWNT Mutli walled carbon nanotubes N A G Nanosilver NM Nanomaterial NP Nanoparticle PSD Particle size distribution QD S Quantum dots
12 ROS Reactive oxygen species SEM EDS Scanning electron microscopy coupled with energy dispersive spectroscopy SSA Specific surface area SUVA S pecific ultraviolet absorbance at UV 254 (m 1 ) SWNT Sin gle walled carbon nanotubes
13 Abstract of Dissertation Presented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy ON THE TRANSFORMATION AND BIO INTERACTION OF NANOSILVER PARTICLES IN NATURAL WATERS: TOXICITY IMPLICATIONS FOR AQUATIC ORGANISMS By Julianne McLaughlin August 2013 Chair: Jean Claude Bonzongo Major: Environmental Engineering Sciences The fast growth of nanotechnology has stimulated research on the potential impacts of its products on both human health and the environment. Results from studies on nanosilver ( nAg ) toxicity to aquatic organisms have raised questions as to whether or not the observed toxicity was due to the dissolution of toxic silver ions (Ag + ) Most studies to date use simple synthetic waters and do not take in to account the complexity of natural waters. The goal of this dissertation research was to comparatively investigate the behavior and toxicity of nAg in natural and synthetic water s A series of batch experiments were conducted to determine the effect of different waters on nAg dispersion, stability, and toxicity in comparison with ionic Ag used as AgNO 3 Waters were fully characterized and v arious techniques were employed t o assess the stability and dissolution of nAg suspensions. Data from the above studies were used to help explain results obtained in two toxicity assays using freshwater invertebrate s and green algae, whic h were per formed in natural waters, traditional growth media and synthetic waters amended with model organic compounds
14 The findings presented in this dissertation help establish needed correlations between water matrix dependent nAg particle properties and toxicity implications us ing new methods that allow for good representation of natural aquatic systems. D issolved organic carbon (DOC) was a clear driver in reduction of negative impacts of nAg to aquatic organisms. Conversely, natural water with moderate ionic strength and low DOC and depending on the test organism used, showe d nAg toxicity trends similar to growth media and in others showed greater toxicity than nAg in growth media. Dissolution results presented here paired with stability results suggest that the toxic impacts observed were not completely due to silver ions released from nAg particles. The use of natural waters to investigate the fate of nAg remains very limited in peer reviewed literature. Similarly, almost no studies have been published using natural waters as both suspension and growth media for toxicity assays. This research shows that behavior and toxicity of nAg in complex natural waters be easily predicted by simple alterations of synthetic waters.
15 CHAPTER 1 NANOTECHNOLOGY AND THE ENVIRONMENT: APPLICATIONS AND IMPLICATIONS 1.1 Problem Statement Nanotechnology is at the forefront of research guiding scientists and engineers to develop more efficient energy production and storage methods; novel biomedical applications; stronger and lighter materials; and smaller computer chips. This emerging and fa st growing technology focuses on the development of techniques with the ability to individually address, control, and modify structures, materials, and devices with nm precision. The products of nanotechnology are now finding use in a wide variety of human activities and their environmental impact could include the release of new classes of toxins with largely unknown environmental and health hazar ds  nanotechnology on both the environment and human health. The growing success of nanotechnology will undoubtedly lead to the increased introduction of nanoparticles (NPs) int o natural systems, and aquatic systems are likely to act as sinks for these new pollutants  Amongs t the wide variety of engineered NPs nanosilver (nAg) is becoming extremely prevalent in consumer products, so much so that the Woodrow Wilson Project on Emerging Nanotechnologies developed a database solely for silver nanotechnology and its applications  The primary goal of this dissertation research was to investigate the effects of natural wate r s chemical composition on both fate and toxicity of nAg. The overarching hypothesis of this study is that the introduction of nAg into natural aquatic systems will result in physicochemical transformations dictated by solution chemistries, with implicat ions on fate and toxicity which differ from those observed in current DI water based laboratory studies. The
16 characteristics of nAg at the point of production and/or use could be vastly different after introduction, transport, and storage in natural water s. Therefore, laboratory experiments have been conducted to investigate the effects of natural water solution chemistry (e.g. effect of ionic strength, DOC type and concentration ) on nAg aggregation and dissolution patterns and the biological effects resulting from such transformations. 1.2 Production and Applications of NPs: General Overview Nanotechnology is an emerging powerhouse in the material science world. According to a recent inventory by The Project on E merging Technologies (2013) the number of nanomaterial based products has grown by nearly 521 % from March 2006 to March 2011 In the above inventory, products containing nanosilver are the most abundant (i.e. 313 products), followed by carbon based nanomaterial products, with 9 1 products (Figure 1 1 ). In this review, NPs will be separated into five major groups based on a combination of chemical composition and geometry (i.e. fullerenes, carbon nanotubes, metal oxides, metals, and quantum dots). Eac h of these groups are presented below to emphasize both differences and similarities in production methods as the latter can have toxicity implications 1.2.1 Buckminsterfullerene (C 60 ) C 60 is a caged molecule comprised of 60 carbon atoms that make 20 hexag onal and 12 pentagonal rings  Other, higher mass, fullerenes have been synthesized, but C 60 is the most widely investigated. For these carbonaceous NPs, C 60 is the most common and is produced by the Krtshmer Huffman method  Briefly, in a quenching atmosphere of approximately 200 torr of helium, graphite rods are evaporated to
17 produce what is called fullerene soo t [4, 5] In general the soot is approximately 15% fullerenes which can then be extracted by various methods based on liquid chromatography (LC) [6 8] Explored applications of fullerenes thus far have been in biomedicine  optics  and electronics  In particular, fullerenes posses promising biological activities including DNA photocleavage, HIV Protease (HIV P) inhibition, neuroprotection, and antibacterial activities [12, 13] As well as the applications above, fullerenes and their derivatives have been shown to be good radical scavengers giving them the ability to quench free radicals more effectively than other antioxidants [9, 14] There is great promise for fullerenes in these fields but there are significan t limitations mainly due to their extremely low solubility in water  Methods that have been employed to overcome this problem include non covalent encapsulation in soluble molecules (e.g. cyclodextrins) [16 18] and covalent functionalization via attachment of water soluble groups [19 21] For a large amount of applications, unmodified fullerenes are the desired product. Colloidal dispersal of these fullerenes in aqueous media is accomplished either by solvent (e.g. tetrahydrafuran (THF)) addition and removal [15, 22] excessive stirring  or ultrasonication [23, 24] Because of the various methods in which fullerenes can be dispersed and/or functionalized, the environmental implications of these materials may also vary greatly. For example, THF/nC 60 has shown various levels of toxicity to a ran ge of organisms including Daphnia magna ( D. magna )  Eisenia veneta earthworms  and bacteria  1.2.2 Carbon N anotubes (CNT) CNT are hollow cylinders made of carbon with diameters ranging between 0.4 and 2.5 nm  The length of these tubes could be from nano to several micro meters  and two types of CNT, the single walled (SWNT) and multi walled (MWNT) carbon
18 nanotubes exist  SWNT consist of a single layer graphene sheet wi th dimensions similar to those described above  Multi walled carbon nanotubes (MWNT) on the other hand are made of two or mor e concentric SWNT with different diameters and lengths. Diameters can be up to hundreds of nanometers  These CNT are generated by use of laser furnace, arc evaporation, and chemical vapor deposition (CVD) methods, with CVD being the most practical synthetic method [30, 31] HiPCO is a type of CVD process that can be scaled up to the industrial level producing a large amount of usable SWNT  Metal catalysts are employed in all of these methods (e.g. iron, nickel, cobalt, and molybdenum)  and are usually abundant in unpurified final products  and can cause increases in toxicity  As well as fullerenes, CNT may be functionalized at the end of production. CNT have properties that are very desi rable in materials including: conducting composite, field emission (FE), electrochemical, chemical, and electronic properties  CNT can improve existing products, making them stronger, more lightweight, and more wear resistant  1 2 3 Metal Oxide Nanoparticles A wide range of metal oxide nanoparticles are currently being investigated including iron oxide, zinc oxide, titania, and ceria NPs. These materials in the bulk form have been used for a variety of applications for many years. The top down approach of grinding bulk materials is the most common synthetic method for industrial production of metal oxide NPs  However many methods are being studied and used including hydrothermal synthesis [36, 37] sol gel chemistry [38, 39] CVD  and solid state pyrolysis [29, 41] Metal oxide NPs, as a whole, are already being used in various applications. For example, titanium dioxide (TiO 2 ) and zinc oxide (ZnO) NPs are being used in sunscreens and cosmetics  because of their ability absorb, scatter, and
19 reflect UVA/UVB rays  These metal oxides are implemented in sunscreens as sizes greater than the nano range, but to be used in cosmetics the metal oxides need to be in the range of 20 50 n m  The smaller size of the particles allows for cosmetic use because of the change from an opaque/white to more transparent forms. TiO 2 NPs have a lso been shown to photocatalytically degrade organic and inorganic compounds as a treatment step for waste water  Iron oxide NPs are being employed in the biomedical field and Ceria NPs have been used as an oxygen sensor and fuel additive  1.2. 4 Nanometals Nanosilver (nAg), nanogold, nanoaluminum (nAl), nanocopper (nCu), and nanozero valent iron (nZVI) are among the most produced and researched nanomet als [44 48] As with metal oxide nanoparticles the methods in which are available to produce metal NPs are extensive. The most frequently used method for production of nAg is chemical reduction in the presence of a stabilizer with reductants such as borohydride, citr ate, and elemental hydrogen [49, 50] Other metal nanoparticles synthetic methods include photochemical reduction  th ermal decomposition in organic solvents [52, 53 ] and radiolytic reduction  As stated earlier nAg has the largest production and applications of all NPs. Nanosilver has antimicrobial activity, which justifies its incorporation in commercial produc ts such as socks, cutting boards, and air filters  Colloidal Au particles have already been implemented in several medical applications  and now nAu are surfacing and showing potential in various applications including electronics and catalysts  The most popular me thod of producing nZVI is the reduction of ferric (Fe3+) or ferrous (Fe2+) salts with sodium borohydride [55 ] nZVI are mainly used for remediation of contaminated ground waters.
20 They have been shown to treat perchloroethene (PCE), carbon tetrachloride   and nitrate pollution  Overall, nAg will most likely have the greatest impact on industry and the environment due to its rapid growth in production and use and its known toxic levels 1.2. 5 Quantum dots (QDs) QDs are semiconductor nanocrystals that have very unique optical properties (e.g. size tunable photoluminescence), which allow them to function as labeling materials  Basically, these NPs have broad absorption spectra and narrow emission spectra which allow for the emission of light at various precise wavelen gths  They have been shown to be useful in biomedical and bi ological applications  QDs range in diameter from 2 10 nm  QDs typically consist of a core, shell, and coating. The reactive semiconductor core of QDs is primarily compose d of a combination of metals and semiconductors such as cadmium selenide (CdSe), cadmium telluride (CdTe), or zinc selenide (ZnSe)  More specifically, these nanocrystals are composed of groups II VI or III V elements [ 58] Once the cores are synthesized they are usually purified and then undergo a shell growth reaction  This is sometimes called capping which refers to the core being capped by an inert mate rial such as silica or ZnS to improve their stabilities (e.g. protecting core from oxidation) and performances [58, 59, 61, 62] The QDs can then be coated with organic layers (e.g. polyethylene glycol (PEG)) allowing them to be used for certain applications  In order to be useful in most biological applications the QDs need to be water soluble. This is usually accomplished by either excha nging the hydrophobic surfactant layer with a hydrophilic layer or adding a second amphiphilic or hydrophilic layer (e.g. cyclodextrin)
21  The largest threat these QDs may pose to the environment is the toxicity of the core materials if their shells and coatings were broken down. 1.3 Environmental Implicatio ns of Manufactured Nanoparticles : General Overview There is a growing concern with regard to the potential implications of NPs on living organisms. NPs are well on their way to mass production, and therefore, entering the environment from different stages in their life cycle. This observation calls for studies on the implications of NPs. As a result, a fair amount of research involving the toxicity of NPs to aquatic organisms has been published in the past decade Unfortunately, the lack of uniformity in e xperimental protocols and standard analytical procedures has led to rather conflicting results. The following section summarizes some of the toxicity results and emphasizes NP toxicity on aquatic organisms. 1.3.1 Toxicity of C 60 and CNT Various negative e ffects on aquatic organisms exposed to C 60 and CNTs have been reported including: respiratory effects  immobilization/aggregation of cells  antimicrobial effects [27, 66, 67] generation of reactive oxygen species (ROS) and associated oxidative stress [68, 69] and ROS independent oxi dative stress [70 72] Carbon based NPs exhibit poor suspension in water due to their high hydrophobicity. They are therefore routinely suspended in organic solvents (e.g. THF, DMSO toluene, etc.) or surfactants (e.g. SDS, SDBS, Triton X, etc) to obtain high degrees of particle dispersion [22, 73 76] Aqueous suspensions have also been prepared either by simple long term stirring of NPs in water to produce hydrolyzed products such as nC 60  or by the back extraction methods where NP initially suspended in organic solvent is back extracted in water [22, 23] In thi s latter case, the NP is then dispersed in water and the solvent evaporated off. Unfortunately, some solvents and surfactants are known toxic
22 chemicals and recent studies on the toxicity of C 60 and CNTs have in some cases attributed the observed biologica l responses to the residual solvents or surfactants [77, 78] However, reports do exist where C 60 /aqueous exhibits toxicity as well solvent suspended C 60 [23, 25] With regard to the toxicity mechanisms, it has been shown that oxidatio butyrolactone and 2 hydroxytetrahydrofuranol) are the substances causing toxicity [78, 79] This was confirmed through laboratory experiments using Daphnia magna ( D. magna ) and A549 lung cells exposed comparatively to extensively pre w ashed and non washed C 60 suspensions following preparation by the THF method. The pre washed C 60 had no negative biological effects  It is clear, even from the small amount of research done, that C 60 greatly influenced by preparation and suspension methods. The exact triggers and mechanisms are unclear, but it is evide nt that size/structure [23, 80] and solvents [78, 79] can play a role in toxic effects. The toxicity mechanism of C 60 has not been fully elucidated but the g eneral consensus is that toxic effects are due to oxidative stress. It is controversial whether or not this oxidative stress is caused by ROS or not [70 72, 81] In reg ards to CNTs, toxicity has been attributed to their metal impurities  as well as residual solvents  and the CNTs themselves [82, 83] 1.3.2 Toxicity of N anometal oxides Compared to other NPs, published studies on the toxicity of nanometal oxides are rather limited. Nano ZnO exhibits toxicity towards a rang e of aquatic organisms including bacteria and has been documented as the most toxic nanometal oxide to aquatic organisms [84, 85] It is worth noting that bulk ZnO can be just as toxic as nZnO to certain organisms. Organisms reported to be negatively impacted by nZnO include Thamnocephalus platyurus ( T. platyurus ) and D. magna  rainbow trout
23 ( Oncorhynchus mykiss )  Bacillus subtilis ( B. subtilis ) and Gram negative Escherichia coli ( E. coli ) [85, 87] and zebrafish embryos and larvae  Nano CuO, nTiO 2 and nSiO 2 are also toxic to aquatic organisms with nCuO with being the most toxic [84, 89, 90] Cytotoxicity of nanometal oxides has been attributed to surface net charge; the higher the cation charge the lower the toxicity observed to E. coli  It is possible tha t more of the smaller charged NPs can be attracted to the negatively charged bacteria causing greater inhibition. An increase in toxicity of nTiO 2 is seen when exposed to light most likely due to the production of reactive oxygen species (ROS) which caus e oxidative stress  However light does not seem to affect the toxicity of nZnO and nSiO 2 and inhib ition is still seen with nTiO 2 in the dark  This is evidence that oxidative stress via photocata lytic ROS production cannot be the sole toxicity mechanism for nanometal oxides. Another possible mechanism is direct NP cell membrane contact and resultant damage to the membrane wall  1.3.3 Toxicity of N anometals In regards to toxicity nAg is the most extensively studied nanometal and is toxic to several microbial groups [92 95] algal species, daphnids, and zebrafish [96 98] and nematodes  It has been debated as to whether nAg toxicity is a result of Ag ions (Ag + ) being released, as with bulk Ag, or if the toxicity is a function of the NPs properties (e.g. small size, large surface area, reactivity, etc.)  Nanocopper and nAu have also been shown to exhibit toxicity to some organisms [96, 101 103] 1.3.4 Toxicity of Q uantum Dot s It has been reported that CdTe QDs induce oxidative stress in freshwater mussels Elliptio complanata  inhibit the growth of the unicellular algae Chlamydomonas Reinhartii  lower hatch rate and increase mortality in Danio rerio
24  Mahendra et al.  showed that the toxicity of various QDs was a function of weathering, where weathering means the breakdown of the coating and/or shell surrounding the core material. Growth was shown to be inhibited when bacteria ( Gram positive Bacillus subtilis and Gram negative Escherichia coli ) were exposed to the QDs, but their effect was only bactericidal wh en the QDs were weathered beforehand (either QDs, it is thought that the toxicity is due to free Cd 2+ ions as a result of the breakdown o er study attributed CdTe QD bacterial toxicity to the formation of TeO 2 and CdO via oxidation  They expressed that toxicity is dependent on surface chemistry, exposure concentration, and the coating. For their specific QDs (water soluble and functionalized with polar ligands) it was seen that release of Cd 2+ ions was not the source of toxicity but instead oxidation of the surface was thought to be the culprit. Whether toxicity is caused by the release of Cd 2+ ions or the oxidation of the core is not clear, but it is obvious that some level of weathering must take place in order for the core of the QDs to be exposed. 1.4 Rationale for Selection of Nanosilver Based on the Woodrow Wilson Project on Emerging Nanotechnologies ( 2013 ) it is indisputable that products containing nAg are significantly outcompeting the other nano based products (Figure 1 1 ) Additionally, the project h as compiled a growing data base solely for nAg based products indicating that out of all of the nanomaterials nAg is the most utilized The variety of applications and uses of nAg and i ts increasing prevalence in our consumer, industrial, and research arenas is one of the main motivations for the selection of nAg for this research. N anosilver is anticipated to enter
25 environmental compartments through various pathways with the main compar tment of concern being aquatic ecosystems. In addition to the rapid increase of nAg production and use, t he body of data on the potential aquatic toxicity of nAg is fast growing. Several studies have investigated the toxicity of nAg on a wide variety of o rganisms including bacteria [2, 109] algal species, fish [96, 98, 110, 111] and invertebrates [112 114] Results from these studies have raised questions as to whether the observed toxicity of nAg was due to its oxidation followed by dissolution of toxic silver ions (Ag + ), or to the direct effects of specific physicochemical properties of the actual nanoparticle  In a study using the algal species Chlamydomonas reinhardtii it has been concluded that nAg particles do undergo solubilization to produce Ag +  There are multiple theories on the toxicity route and mechanism of nAg [99, 112, 115, 116] however it is highly challenging to draw sound conclusions. Although the use of synthetic waters (e.g. altered DI water) can help account for the specific effects of each of the relevant water chemistry parameters (e.g. pH, quantity and types of bot h ions and dissolved organic carbon (DOC)), it does not take into account the complexity of natural water matrices, which could result in transformation and dissolution patterns which are different from those observed with synthetic waters. With increasin g production, nAg will undoubtedly enter aquatic ecosystems. More and more research is pointing to silver ions (Ag + ) being the main cause of toxicity greatly needed. Silver ions may be what are inducing the toxic responses, but the mechanisms controlling the release of Ag + are still unfolding and will
26 Ag + bioavailability, toxicity, and fate is controlled by environmental parameters but is complicated by its small size, large surface area, and multitude of capping agents and stabilizers. 1.5 Scope of Research The research presented in this dissertation addresses the environmental implications of nanosilver (nAg) as an emerging pollutant and its fate and exposure in natural and synthetic waters. Natural waters from two locations, Alachua Conservation Trust freshwater marsh wetland (AC T) and a riparian wetland fed by Itchetucknee Springs (SPG), were used throughout the entire research project to compare nAg transformations and nAg induced biological responses in these waters as well as in synthetic waters. This introductory chapter (Chapter 1) outlined the state of nanotechnology and the potential environmental implications while presenting a rationale for studying nAg. The rest of the dissertation is structured as follows. Chapter 2 presents a review of peer reviewed literature on n Ag in the context of implications on aquatic ecosystems. Nanosilver synthesis, applications, toxicity to three trophic levels of aquatic organisms, fate in aquatic systems, and overall considerations are discussed. Chapter 3 focuses on the environmental t ransforma tions of nAg in natural waters as compared to Nanopure water and growth media using water chemistry, zeta potential, particle size distributions, scanning electron microscopy, and dissolution kinetics to describe nAg behavior under various circum stances. Chapter 4 is published in Environmental Toxicology and Chemistry and presents the initial findings of nAg impacts on Ceriodaphnia dubia ( C. dubia ) and
27 Pseudokirchneriella subcapitata ( P. subcapitata ) when exposed in natural waters as compared to in traditional growth media. Chapter 5 is an extension of Chapter 4 and it focuses on the identification of water parameters controlling the biological responses of used test organisms. M odel organic compounds are used in synthetic waters to validate con clusions reached with waters of different chemical composition using C. dubia and P. subcapitata as model organisms At the same time Chapter 5 presents the effects of ionic Ag as compared to nAg in natural and synthetic waters. Chapter 6 presents conclud ing remarks and states the different avenues for future research.
28 Figure 1 1 Temporal trends of production of engineered nanomaterials commonly used in commercial products. (Adapted from Emerging Nanotechnologies, 2013 ).
29 CHAPTER 2 NANOSILVER AND THE ENVIRONMENT : CURRENT STATE OF KNOWLEDGE 2.1 Applications and Current Uses of Nanosilver Nanosilver (nAg) is being rapidly incorporated in numerous consumer products as well as implemented in scientific and industrial applications. It is being used as antibacterial agents in wound dressings, detergents, antimicrobial coatings, surgical instruments, hand sanitizers and soap toothpaste, personal groomers, and textiles (sportswear, military clothing, underclothes); in appliances (washing machines, refrigerators, water purification, HVAC filters); as conductive inks in the electronic industry, glass coatings for thermal insulation and silver inks for solar panels for se nsing applications, as reaction catalysts, and for medical applications  The chemical, phy sical, biological, and optical properties possessed by nAg are the driving forces for its wide range of applications. One of the properties taken advantage of the most is antimicrobial activity which can be size, shape, and c oncentration dependent  Others include antiviral, anti inflamatory, anti biofilm, surface plasmon resonance, plasmonic heating, and metal enhanced fluorescence properties  It is important to note that c olloidal silver has been used for over a hundred years and primarily at the nanoscale, although it was never called by nanosilver until recent decades  It is thought that the first report of the synthesis of nAg was in 1889  For decades, nAg has been synthesized and used for medical and biocidal applications  In fact the first registered biocidal product in the U.S. using nAg was registered in 1954 under the Federal Insecticide, fungicide, and Rodenticide Act (FIFRA)  Although these Ag collo ids have been produced and used for such a
30 long period, environmental regulations and studies have always been implemented and conducted for the ionic form of Ag. Industrial production of nAg materials has been roughly estimated by only a few. In 2008 Mueller and Nowack  estimated the annual production of n Ag worldwide to be 500 tones  For the U.S. an estimated 2.8 20 tons of nAg is produced annually [ 122] whereas for Europe productions was proposed to be between 110 and 230 tons per year by the year 2010  Switzerland was estimated to have a nAg production between 0.026 and 4.03 tons per year  This widespread use of n Ag will likely result in the introduction of both nAg and its degradation products into wastewaters and subsequently aquatic ecosystems. 2 .2 Synthesis Methods for Nanosilver A wide range of synthesis methods for nAg and NPs in general are currently being employed includes p hysical methods such as photolithography, laser beam processing, and mechanical techniques (grinding and polishing). silver is nanosi zed. assembly) strategy which includes wet chemical met hods such as organic synthesis chemical reduction, biosynthesis, and photochemical reduction  For nAg the most widely used route of synthesis is the reduction of metallic salts (e.g. AgNO 3 ) in solution (dissolved in solvent) to zero valent silver (Ag 0 ) [53, 125, 126] The process is straightforward, whereby a source of Ag ions (e.g. AgNO 3 ) are reduced with some sort of reductant (e.g. sodium borohydride, ferrous citrate, ascorbate, formaldehyde, dextrose). Zero valent silver (Ag 0 ) is formed as a result of this reaction followed by their agglomeration into
31 oligomeric clusters (i.e. colloidal Ag NPs). Stabilizing agents or capping agents, are used if needed to prevent excessive agglomeration. The s ilver salt precursors solvents, reductants, and stabilizers used to synthesize nanosilver particles all have impacts on the end result. The morphology and surface chemistry o f synthesized nAg particles are greatly influe nced by experimental conditions, kinetics, and adsorption of the stabilizing agent. Specifically, important parameters include the molar ra tio of the stabilizing agent to the silver salt, the redox potential o f the reductant, the chemical nature of the stabilizing agent, and the mixing speed and temperature of the reaction [ 126 ] Capping agents (also termed shells) including inorganic and organic molecules, are especially important in the overall behavior of nAg. They stabilize nAg via either electrostatic, steric, or electrosteric repulsive forces [127, 128] If capping agents are applied during NP synthesis they will provide control over size and shape of the particles  The Ag 0 core can vary in size and shape including rods, cubes, triangular plates, and quasi spherical NPs [126, 129, 130] Organic capping agents are typically used, especially for environmental studies, and vary in molecular weight and types of functional groups where they stabilize through adsorption to the particle surface or through covalent attachment  Either way, the Ag 0 core usually has some an impermeable barrier. The most implemented capping agents or coatings used to stabilize nAg include carboxylic acids; mainly citrate, vari ous polymers, polysaccharides (e.g. gum arabic), surfactants (e.g. PVP) and biological macromolecules [131, 132] From the production side t he behavior of nAg in the environment and interaction with
32 biota will vary depending on the type of coating or lack of one size of nAg particles, and shape of nAg particles 2.3 Toxicity of Nanosilver to Aquatic Organisms A great challenge is present among the research community, in that it is almost impossible to effectively compare and contrast the growing body of aquatic toxicity res ults for nAg as well as other NPs. Because there are no standards for NP toxicity testing and nanoecotoxicology is still in its infancy the literature is quite varied in regard to method ology and results. Nanosilver can be produced by several methods which will result in nAg with different by products. Moreover, nAg is synthesized with a wide range of capping and stabilizing agents as well as varied morphologies and sizes, oute dictate toxic implications, but also the sample preparation for the toxicity study, the dosing regimen, test conditions (e.g. lighting) and even the organism tested. 2.3.1 Bacteria Nanosilver toxicity to bacteria is more widely investigated than ot her organisms because of interest shown by the medical community; however the mechanisms and dose responses are not comprehensible  Several mechanisms have been investigated and proposed in regard to the toxicity of nAg, however it is difficult to generalize the effects to bacteria due to the huge variations in physiology, anatomy, a nd behavior among different strains [133, 134] For nitrifying bacteria Choi et al  showed that PVA capped nAg toxicity was size dependent and suggested that the smaller nAg particles exhibited more toxicity because they could transport across the cell membrane more readily. Others have also indicated that size plays a pivotal role in nAg toxicity to bacteria where by smaller particles are more toxic [94, 95, 112, 116, 135,
33 136] Some indicate that the small particle toxicity is due primarily to attachment to and/or penetration of the cell membrane [94, 95] On the other hand, some studies have suggested that nAg toxicity is caused by silver ions (Ag + ) which are released from the NPs [98, 135, 137] When compared to nAg, in some cases, ionic silver (Ag + ) has shown greater toxic effects to bacteria [135, 138, 139] Xiu et al.,  confirm ed for E. coli that nAg toxicity is fully attributed to the release of Ag + Based on anaerobic experiments where the n Ag particles undergo zero dissolution, the toxicity to E. coli was insignificant with concentrations up to 195 mg/L nAg The authors also observed size dependent toxicity, where the EC 50 exposed nAg increased with increasing particle size, illustra ting that smaller nAg particles release more Ag + A growing interest among researchers is assessing how ligands ( e.g Cl cysteine, S 2 ) may impact toxicity. For bacteria, some studies suggest that addition of ligands (i.e. Cl cysteine, S 2 ) mitigate the toxicity of Ag + to a much greater extent than for nAg [140, 141] Another factor that confounds nAg toxicity results is the capping agent; coatings, like ligands, can and will have different effects on observed tox i city [136, 142, 143] This has to be considered when analyzing toxicity results for any type of organism. Complex microbial communities are not as thoroughly studied as pure cultured microorganisms and there are conflicting results as is seen throughout the research studies in envi ronmental nano toxicology The importance of studying the effects of nAg and other NPs on natural microbial communities and wastewater sludge is equal to, if not greater than studying pure cultures. Fabrega et al.  showed nAg in sodium citrate to have a negative impact on marine biofilm volume and biomass at
34 concentrations of 200 g L 1 a nd greater. In a study on the impact of polyvinylpyrrolidone (PVP) coated nAg on wastewater activated sludge it was found that nAg was toxic but was strain dependent  Conversely, Bradford et al.  concluded that nAg presented virtually no impact on the studied estuarine sediment bacterial diversity. Given the fact that nAg is largely used as an antimicrobial agent, it is reasonable to expect that if enough is released into the environment there could be negative impacts on natural microbial communities. 2.3.2 Algae Exploring the implications of nAg on different algal species (Eukaryot es) is essential for understanding the potential effects nAg co uld have on aquatic ecosystems. Algae are autotrophic organisms (i.e. primary producers ) whereby they consume sunlight to produce chemical energy i.e. (organic compounds) for organisms in other trophic levels to consume They are the base of most aquatic food chains as well as important in biogeochemical cycling making them extremely important indicators for the health of aquatic ecosystems. Correspondingly, t he health o f most aquatic ecosystems is dependent on the health of the autotrophs (e.g. plants, algae, many bacteria) within the system. The debate on toxicity pathways for nAg and algae is far from over; there are still few studies in comparison to bacteria and i nvertebrates. M ultitudes of algal species exist and some uptake Ag + at different rates which affects the overall observed toxicity [145, 146] among research to see that Ag in the ionic form is more toxic than nAg when exposed to algae [116, 147] but if nAg is s dissolved it becomes as toxic or more toxic than the ionic Ag  For e xample, Navarro et al.,  exposed Chlamydomonas reinhardtii (a
35 fast Ag uptake species) to carbonate coated nAg and found AgNO 3 to be 17 times more toxic, but when they expressed the concentration of nAg as Ag + present the EC 50 was much lower (i.e. more toxic) for nAg than Ag + Regardless of how it is expressed th e bulk Ag ( AgNO 3 ) was more toxic than nAg but the above results indicate that the mechanism of toxicity for nAg may not be due solely to the ionic fraction, at least for algae. The bi nding Ag and therefore making it unavailable to the algae T hey also conclude d that more Ag + was formed once the algae was exposed to nAg suggesting nAg will persist in the system As with bacteria toxicity studies for algae have shown that coatings (i .e. capping t consistent and studies to make good correlations  Some report that the addition of cysteine amelerioates toxicity due the scavenging of Ag + suggesting the main toxic pathway is through the release of Ag ions [98, 147] Miao et al.  also found that release of Ag + was the main cause of toxicity to a marine diatom Thalssiosira weissflogii Barely any research has been conducted investigating the impact of natural waters on nAg and its toxicity to algae. McLaughlin and Bonzongo [ 159 ] used uncapped nAg and showed toxicity to be much higher ( IC 50 = 1600 ppb) in natural waters with high organic matter than in natural waters with low organic matter and moderate ionic strength ( IC 50 = 22.6 ppb). This study provided evidence for the need of further research in the area of natural waters. important to understand the mechanism for Ag + toxicity as it obviously plays a role. The
36 bulk of research for algae has been on uptake rates and how other ligands (i.e. Cl ) effect the uptake rate of Ag + [145, 149, 150] but not necessarily the exact mechanism of toxicity. It is thought that accidental Ag + transport through a Cu + transport system is how Ag + accumulates in the algal cell since these two metals possess similar properties and other modes of uptake have been ruled out  It has also been shown that silver uptake is enhanced by the presence of silver chlo ro complexes and by the presence of thiosulfate however mechanisms of transport were found to be different [149 151] It should be noted that different algal species exhibit varying rates of uptake of Ag +  In regard to actual mode of toxicity, Szivak et al  showed that Ag + was a strong ROS inducer in Chlamydomonas reinhar dtii which implies that oxidative stress is one cause of toxicity to algae 2.3.3 Invertebrates Unlike algae, invertebrates and nAg toxicity are more thoroughly researched, almost all of which deals with freshwater cladocerans (i.e. water fleas). I n aquatic ecosystems, cladocerans are an important source of food for fish and other higher trophic organisms. They are filter feeders who mainly eat algae and are important contributors in freshwater food webs  They are extremely sensitive to water quality conditions making them good test species. Most cladoceran nAg toxicity studies to date have been performed with Daphnia magna ( D. magna ) as their test organism, a lthough some use Ceriodaphnia dubia ( C. dubia ) and Daphnia pulex ( D. pulex ). A definite correlation between toxicity and nAg particle s ize has been established where smaller sizes are more toxic in theory due to their greater dissolution [154, 155] Additionally, when compared to Ag + nAg has shown lower tox icities in a number of studies [154, 156 158] Conversely, for C. dubia nAg has
37 been shown to enhance toxicity when compared to AgNO 3  The role of capping agents in toxicity has been explored with some indicating bare nAg is more toxic than various capped nAg and some indication of the opposite [113, 143, 155, 157] It is obvious that the type of capping agent will play a role and this area should be investigated further. As with the study by Navarro  there has been talk of expressing the toxicity of nAg as the dissolved fraction (Ag + ) to compare to AgNO 3 toxicity results. Kennedy et al. 2012 exposed citrate cap p ed nAg to C. dubia in the presence of SRHA where they found the toxicity to decrease with increasing SRHA concentrations as others have reported with organic matter [2, 114, 156, 159] AgNO 3 was found to be more toxic but when LC 50 s were calculated based on the dissolved fraction nAg and AgNO 3 toxicities were comparable. They also found that t oxicity increased with storing the NP s for 3 0 days which correlated to the dissolved Ag released in th e stored suspension Uptake of nAg and Ag + for invertebrates is going to be different than for lower trophic level organisms such as bacteria and algae as will the toxicity mechanism. In one study by Zhao and Wang  they showed that 70% of the accumulated carbonate coated nAg was through the ingestion of algae which were exposed to nAg. Others have shown uptake via imaging which suggests to some degree that nAg is ingested possibly causing a very different mode of toxicity from Ag ions  Once in the gut nAg could undergo more dissolution so there would be Ag + toxicity on the outside and the inside of the organism. The toxicity mechanism for ionic Ag is also likely different than that for nAg (without ions) for freshwater invertebrates. For freshwater animals the gills are the main
38 site of Cl and Na + active transport from the water to the extracellular fluid. This Cl and Na + uptake, which is needed to counteract the loss of these ions through the gills and excretory organs, is directly related to branchial Na + K + dependent adenosine triphosphatase (Na + K + ATPase) activity  It is thought that Ag + binds to this biotic ligand and that th e binding is directly related to Ag + toxicity A direct correlation between Na + K + ATPase activity and whole body accumulation of Ag + in D. magna was found  It is unl ikely nAg itself would induce toxicity via this route however any Ag ions released should cause negative impacts in this way. There are only a few studies using other invertebrates besides cladocerans that are worth noting. One being a toxicity study done on Caenorhabditis elegans ( soil nematode ) where Ag NPs were found to be slightly more toxic than ionic silver (AgNO 3 )  In this study they attributed increases in certain gene expressions to oxida tive stress, and they noted that the toxicity mechanisms for Ag NPs and Ag ions appeared to be different because of differences in gene expression patterns. 2.4. Nanosilver Interactions in the Environment 2.4.1 Mechanisms of Nanosilver Dissolution Nanomet al and especially nanosilver dissolution is increasingly the focal point of research dealing with the environmental implications of nanotechnology Nanosilver undergoes dissolution at varying rates depending on water chemistry, temperature, particle coati ngs, and particle size. In aerated aquatic conditions nAg should react as follows. ( 2 1) (2 2 )
39 Equation 2 1 is the initial oxidation step whereby a Ag 2 O layer is created on the surface of the NP. The formation of Ag 2 O is thermodynamically favored at room 0 298 = 11.25 kJ/mol) and has been observed under many conditions including the presence of capping agents [131, 162 164] From Cai et al .,  and based on thermodynamics a decrease in particle size increases th e free energy of formation of Ag 2 O with (kJ/mol) (2 3) Thus it is understandable that nAg undergoes oxidation faster as the particle size decreases. It has also been shown that the redox potential decreases with decreasing nAg particle size which lends them to be more susceptible to oxidation [165, 166] It has been shown that in order for nAg to become oxi dized and create an Ag 2 O layer, O 2 or another oxidant must be present in solution [109, 167] Following oxidation, the oxide layer dissolves and releases ionic silver (Equation 2 2) The dissolution of nAg is widely acknowledged as the most relevant pathway of toxicity for aquatic organisms, with a significant portion of studies now tracking the dissolution alongside toxicity [111, 135, 140, 155] Several are looking at the dissolution in various abiotic media [168 171] and a couple research groups are looking muc h deeper at the mechanism of oxidation [167, 172 174] Although it is not definitive that nAg must first undergo oxidation at the surface creating a Ag 2 O layer before Ag + can be released, this is the likely pathway. S otiriou et al.  showed that in simple aqueous solutions, nAg develops 1 to 2 Ag 2 O surface mono layers in which the Ag ion released corresponds to the mass dissolved from the monolayers. They were able to deduce this by quantitatively tracking Ag + by mass
40 balance through comparing nAg particle size distributions to the equilibrium Ag + concentration  Based on their results the authors hypothesize that after the initial Ag 2 O layer dissolution the nAg surface is not further oxidized however this may not be true for complex environmental matrices Interestingly, peroxide intermediates have been d etected when citrate stabilized nAg is suspended in aerated DI water and the suggested pathway is presented below  ( 2 4 ) They successfully showed that in aerated DI water, nAg produced H 2 O 2 while AgClO 4 and nAg plus bovine liver catalase produced no H 2 O 2 This could be evidence for the ability of nAg to induce the formation of ROS (i.e. H 2 O 2 ) Another group looked at the potential of ROS generation from another angle. They showed that at low pH and in the presence of H 2 O 2 nAg decomposes H 2 O 2 and forms hydroxyl radicals accompanied by nAg dissolution. Conversely in basic conditions O 2 increased significantly at higher pHs as did the dissolution of nAg  They proposed the following reaction in acidic conditions whereby nAg acts as a Fenton like reagent. ( 2 5 ) Hydrogen peroxide is generated constantly by cells because it is part of their redox hydroxide radicals would form and generate oxidative stress. The stud y above indicate s that the oxidation and subsequent dissolution of nAg may be complicated even more once the nAg particles come in contact with biota.
41 For environmental waters (pH range 4 9) which are aerated, it is hypothesized that nAg forms an Ag 2 O surface layer and Ag + is then released. In general, d issolution is enhanced as pH decreases, as particle size decreases, and as concentration of nAg decreases [131, 167, 175] Environmental water chemistry will be a dictator of dissolution rates and pathways as well as nAg coating types. All of these factors will contribute to potential environmental implications. 2.4.2 Effects of Organic Matter Natural organic matter (NOM) h as been shown to reduce the toxicity of nAg [114, 159] and as expected it has been shown to reduce the dissolution of nAg [155, 167] In one study, b oth Suwannee River humic acid ( SRHA ) and Suwanee River fulvic acid ( SRFA ) showed the same dose dependent decrease in release of dis solved Ag from citrate stabilized nAg Starting with 0.05 ppm nAg, at 0 ppm NOM there was a release of approximately 0.03 ppm dissolved Ag and by 50 ppm NOM there was no release of dissolved Ag  Gao et al.,  demonstrated that bare nAg SRHA adsorbs to the surface of the particles following a Langmuir adsorption pattern. They also showed that when exceeding 10 ppm SRHA the stability of the nAg particles significantly decreases suggesting the organic matter, at this point, was bridging the nAg and causing it to settle out of solution. The presence of NOM most likely reduces initial oxidation by adsorbing or coat ing the particles while it also will scavenge the majority of Ag ions that are released from the particles. 2.4.3 Effects of Inorganic Ligands, Ionic Strength, and pH Other water chemistry parameters that can significantly control the fate of nAg in the environment are the presence of inorganic ligands (e.g. chloride, sulfide ) changes in ionic strength and pH and possibly the presence of light Many have explored the
42 effects of these parameters individually on nAg fate ; some have looked at a couple par ameters at the same time, and very few have used natural waters to investigate the effects of water chemistry. Only a small amount of research has been conducted on the impacts of inorganic ligands on nAg, however it is acknowledged that i norganic ligand s will potentially play a significant role in the fate of nAg in aquatic systems. Chloride (Cl ) and sulfide (S 2 ) are known for their strong affinity for Ag + and ability to precipitate very insoluble AgCl and Ag 2 S colloids [146, 176] Inorganic ligands such as these may interact with the particle surface [174, 177, 178] and will definitely interact with any Ag + released from the particles either forming precipitates or soluble complexes [131, 140, 176] If ligands and Ag + are present at relatively low concentrations in the environment (i.e. below their solubility product (K sp )) then they will form soluble complexes which would have different fates and impacts than precipitated colloids. Thermodynamically speaking Ag + released from nAg will react with inorganic HS (anaerobic conditions) or organic thiols and Cl preferentially to other inorganic ligands  The effect of ionic strength on nAg stability and fate is well studied as it is an easy parameter to control in the lab and very relevant in terms of natural water chemistry [178 181] Waters with high ionic strengths, especially containing divalent cations, cause nAg particles to become unstable and aggregate. This behavior is supported by the Derjaguin Landau Verwey Overbeek (DLVO) theory of colloidal stability. A s the concentration of electrolytes increases there is an increase in screening of particle surface charge, which reduces the energy barrier and tilts the system towards
43 aggregati on [182 184] Aggregation of nAg will affect toxicity in different ways depending on the organism tested and should also a ffect Ag + release rates. One study has linked the aggregation patterns of nAg to its dissolution in a natural river water. Li and Lenhart,  used natural river water to look at three different types of nAg (bare nAg, Tween coated, and citrate coated) and the effects of light on their aggregation and Ag ion release. Silver ion release was not affected by presence of light but was a ffected by capping agent. Tween nAg rapidly released Ag + and then leveled off w here citrate nAg and bare nAg gradually released Ag + over time (days) and seemingly plateaued at relatively the same concentration Ag + as Tween nAg by day 16. They attributed the differences in release rates to the fact that Tween nAg was well dispersed i n the natural water while citrate nAg and bare nAg aggregated due to high concentrations of divalent cations (i.e. Ca 2+ Mg 2+ and SO 4 2 ) If cations exceed the critical coagulation concentration (CCC) the system will undergo aggregation. The Tween nAg sho wed a decr ease in particle size over time, but could not fully be explained. Released Ag + for all types of nAg was approximately 3% of the total Ag. The pH of the system is also linked to particle stability and Ag + release. When there is no DOC present, nAg is stabilized primarily via electrostatic repulsion forces where highly charged particles (negative or positive) are more stable. It should be noted that nAg particles can be stabilized with a variety of compou nds (i.e. sterically, electrostatically, and electrosterically) that will result in various reactions to pH changes as well as other environmental parameters. El Badaw y et al.,  showed that a change in pH could have a big impact on nAg stability but it strongly depended on the type of particle coating. For example bare nAg was fairly unstable at low pH until it
44 reached a neutral pH, whereas PVP nAg exhibited no changes over the pH range (2 10). Another study showed that for citrate stabilized nAg decreasing pH correlated with increasing release of Ag + from nAg surfaces  The study by El Badaw y et al.,  demonstrated that citrate stabilized nAg particles were less stable at lower pHs. There is som e discrepancy in the literature when linking aggregation patterns to dissolution patterns based on water chemistry therefore it a general trend cannot be stated 2.5 Considerations and Suggestions As evidenced here, research on the environmental impl ications of nanosilver continues to grow However, there is need for a much deeper understanding of nAg interactions with natural waters and the potential implications on aquatic organisms. Environmental regulations of Ag are designed for dissolved Ag not particulate Ag toxicity of nAg, in a lot of cases, seems to be driven by ionic Ag released from the particles, it is apparent that this oxidation process and the interact ions of nAg particles with complex environmental matrixes need to be investigated more th o roughly Nanosilver has the potential to undergo numerous types of transformations when introduc ed to natural aquatic systems and ultimately those transformations will dictate transformations are dissolved oxygen, electrolytes or pH, Ag complexing ligands such as Cl and S 2 light, and DOC ; all of which have mainly been studied separa tely or in pairs Interestingly it has also been shown that DOC can induce formation of nAg from ionic Ag [186 188] Even with ionic Ag it is sometimes hard to predict behavior and toxicity in complex matrixes making nAg much more difficult to predict It is therefore nece ssary for researchers to investigate the fate and impacts of nAg in different types of
45 natural waters alongside synthetic waters in order to obtain a better understanding of what may happen once nAg reaches high enough concentrations in our waterways. Other issues that should be addressed more thoroughly are the effects of capping agents on nAg interactions and toxicity in natural waters. Some have begun to explore this area and it is very obvious that nAg with different capping agents will react varia bly when introduced to aquatic systems [169, 189] Additionally, when researchers report toxicity data they should consider the chemistry of their growth media and how that may impact nAg behavior and toxicity.
46 CHAPTER 3 TRANSFORMATION AND FATE OF NANOSILVER PARTICLES IN WATERS OF DIFFERENT CHEMICAL COMPOSITION 3.1. Introduction The growth of nanotechnology and the anticipated widespread use of products containing engineered nanomaterials continue to raise concerns as to the fate and exposu re of these emerging pollutants in aquatic systems. With regard to nanosilver (nAg) particles, the well known antibacterial and toxic properties to fungi and algae of silver has led to a dramatic revival of its use by taking advantage of nanotechnology. As discussed earlier in Chapter 2, nAg is now used in an ever increasing range of products, including washing machines, dyes/paints, medical applications, and various consumer products such as disinfectants, cosmetics, cleaning agents, and baby bottles to name a few. However, current research tends to suggest a strong potential for significant adverse biological effects. In fact, there is clear evidence that nanosilver is toxic to aquatic and terrestrial organisms as well as a variety of mammalian cells tested in vitro Unfortunately, knowledge gaps persist with regard to the determination of the potential environmental fate and implication of nAg, primarily in aquatic systems. This stems, at least partly, from the fact that most disciplines dealing with the different aspects of engineered nanomaterials are not mature. Therefore, a good number of challenging rese arch questions remain unanswered, particularly in terms of accurate assessments of the en vironmental fate and exposure, as well as in the identification of clear detrimental effects of engineered nanomaterials in natural systems. A number of published studies have investigated the effects of solution chemistry on surface properties of nanopar ticles including nAg. While most of these studies have used synthetic waters to control solution chemistry [167, 178 180, 190] a rather small
47 number of papers have focused on the fate of nAg in natural waters [185, 191, 192] and the subsequent biological implications [2, 139, 159, 192] Overall, studi es on fate of nAg in aqueous solutions have pointed out the importance of three major water parameters, namely, pH, ionic strength and the concentration of dissolved organic carbon or DOC [133, 177, 180] A study by Chen and Zhang (2012) showed that nAg aggregation rate in water increases with increasing concentration of cations, while the critical coagulation concentration (CCC) of nAg in water tends to follow the order CCC Na+ (DOC) > CCC Na+ (no DOC) > CCC Ca2+ (DOC ) > CCC Ca2+ ( no DOC ) In the present study, ionic strength and dissolved organic carbon were used as end member parameters in the selection of natural water samples to be tested. The effect s of the complexity of solution chemistry on the transformation of nAg pa rticles was therefore investigated using natural water samples exhibiting the prevalence of one of the above two parameters. Experiments using synthetic waters (e.g. de ionized water and organism growth culture media) were also run in parallel for compari son purposes. Prepared nAg suspensions were characterized and analyzed using a combination of physical and chemical methods. The results show that the fate of nAg in aqueous solutions is tightly coupled to solution chemistry, resulting in differences in th e reactivity of the nAg particles. 3.2. Materials and Methods 3.2.1. Collection and Characterization of the Different Waters Used in this Study 126.96.36.199. Sample collection and handling N atural water samples were collected in pre cleaned high density polye thylene (HDPE) bottles from two different sites near Gainesville in North Central Florida, USA. The first sampling site is a freshwater marsh wetland within the A lachua C onservation
48 Trust lands on the Prairie Creek Preserve in Gainesville, Florida. This s ampling site will be referred to as ACT from here on. The second site is a riparian wetland on a spring fed river, the Ichetucknee River, located near Ft. White, Florida. This site will be referred to as SPG from here on. To account for temporal variabili ty in solution chemistry, water samples were collected from the above two sites on two different occasions. The ACT site was sampled on 5/11/2010 and 8/13/2010, while the SPG water samples were collected on 1/22/2010 and 3/15/ 2011. After collection, wate r samples were transported to our laboratory where they were first left to decant to help eliminate large debris through sedimentation by gravity. The water samples were then filtered through a 1.6 m membrane filter. Next, an additional filtration through 0.45 m membrane filter was undertaken. Filtered waters were stored refrigerated at 4 o C pending analysis and use in different laboratory experiments. 188.8.131.52. Sample analysis T he chemical characterization of the collected water samples included the dete rmination of dissolved organic carbon (DOC) and inorganic carbon (DIC) using a Shimadzu TOC V CPH total organic carbon analyzer. The pH was determined using an Accumet AB 15 pH meter and concentrations of major ions by ion chromatography (Dionex ICS 3000 IC ). Additionally, ultraviolet visible (UV Vis) absorbance and fluorescence excitation emission (EEM) techniques were used (Hitachi U 2900 UV visible spectrometer and Hitachi F 2500 fluorescence spectrophotometer), to qualitatively gain insight into the dif ferent types of organic compounds present in these water samples  UV absorbance at 254 nm (UV 254 ) and DOC data were sought to allow the determination of specific UV absorbance or SUVA, which is used to estimate
49 the degree of aromaticity of dissolved organic molecules  Finally, the excitation emission matrices (EEM) of tested waters were obtained by first diluting the water samples, placing them in a 1 cm quartz cell and simultaneously scanning at 5 nm increments over an excitation wavelength range of 200 to 500 nm and an emission wavelength range of 200 to 600 nm. Obtained raw EEM data were then processed using MATLAB (Mathworks)  For these analyses, deionized water EEMs were run as controls and subtr acted from actual sample EEMs and the intensities normalized by Raman water area  Contour plots were generated and compared to a reference plot  to identify the dominant groups of DOC compounds present. 3.2.2. Preparation and Characterization of nAg Stock Suspensions Nanosilver particles in a powder form were purchased from Quantum Sphere, Inc. (Santa Ana, CA), and the following was report ed by the vendor. Nanosilver particles were produced by use of a patented vapor condensation process resulting in a >99.9% purity of face centered cubic crystal structures with a bulk density of 0.25 g/cm 3 a melting temperature of 962C, a mean particle s ize of 25.4 nm and an average specific surface area (SSA) of 22.5 0.25 m 2 /g. Multipoint Brunauer Emmett Teller (BET) analysis (Quantachrome NOVA 1200) was used to verify the reported SSA. Following receipt of these nAg particles, they were analyzed for particle size distribution (PSD) Particle Engineering Research Center (PERC) and Major Analytical Instrumentation Center (MAIC). For the analysis of raw nAg by SEM EDS, appr oximately one gram of as received nAg powder was mounted on a double sided carbon tab, carbon coated, and then imaged at various magnifications.
50 To prepare nAg suspensions, in addition to the above described two natural water samples, nanopure water (NP ) and two growth culture media (algal culture medium (CM) and moderately hard water or MHW) were also used. For all these waters, suspensions of nAg were p repared by first dispersing 200 300 mg of nAg powder into 150 mL of water. The obtained mixtures were s haken on a table shaker for one week and then filtered through a1.6 m membrane filter to remove large nAg aggregates as described by Gao et al.  Next, aliquot volumes of obtained filtrates were used in different analyses. The total concentration of Ag (Ag T ) in the filtrate was determined by inductively coupled plasma atomic emission spectroscopy (ICP AES, Pe rkin Elmer Optima 3200 RL) following a nitric acid digestion procedure adapted from U.S. EPA Method 3005A  The particle size distribution (PSD) and average particle diameter was determined via dynamic light scattering ((DLS) The first set of PSD data was obtained with a Brookhaven ZetaPlus and the second with a Microtrac Nanotrac Ultra particle sizer. nsion was determined using a Brookhaven ZetaPlus and the phase analysis light scattering (PALS) technique. The electrophoretic mobility was converted to zeta potential via the Smoluchowski equation within the PALS Zeta Potential Analyzer Sofware  Results were the average of 10 runs each of which were comprised of at least 15 cycles. 3.2.3. Imaging of the different nAg suspensions Scanning Electron Microscopy (SEM, JEOL JSM6330F) coupled with energy dispersive X ray spectroscopy (EDS) housed in MAIC at the University of Florida was used in the analysis of nAg suspensions prepared as described above. A droplet of sample was placed on an aluminum mount, allowed to air dry, carbon coated, and then imaged at various magnifications.
51 3.2.4. Dissolution Kinetics Recently, it has been established that nAg suspended in water undergoes dissolution, especially in oxygen saturated waters [110, 139, 167, 185] Dissolution kinetics experiments were conducted in this study using only three nAg suspensions in ACT, SPG, and NP waters. Before initiating kinetics experiments c entrifugal filter units were pre equilibrated with respective sample matrices free of Ag additions. At time zero, 200 mg of nAg powder was added to 400 mL of each of the tested waters in a 500 mL Erlenmeyer flask. Suspensions were next placed on the shak er table in the dark for 20 minutes and then allowed to settle for 10 minutes for sedimentation of large aggregates. Silver release and dissolution from the suspended nAg particles was quantified in aliquots of samples collected from the 500 mL Erlenmeyer as a function of time. Amicon centrifugal ultra filter devices (Amicon Ultra 15, Millipore) containing cellulose membranes with a NMWL cut off of 3 kDa were used to separate nAg particles from dissolved ionic Ag (Ag + ). This molecular weight cut off is e stimated to be in the range of a 1 to 2 nm pore size  For the separation process, 10 mL of the prepared suspension was centrifuged in a pre equilibrated filter unit at 5,000 rpm for 75 min with a Beckman JA 14 rotor. Our preliminary tests showed that the totality (i.e. 100%) of the initial Ag + concentration was recovered in the filtrate when a solution of Ag prepared with AgNO 3 was used. This indicated that these filters would not retain dissolved Ag + Other research groups have also reported successful uses of these membranes [167, 175, 185] Finally, after centrifugation, filtrate portions of the samples were acidified to a final concentration of 3% (v/v) with HNO 3 and stored in the dark at 4C till analysis. Ag concentrations were then determined by ICP AES.
52 3.3. Re sults and Discussion 3.3.1. Chemical Analysis of Used Waters The chemical composition of the different waters used as solvent in this study is given in Table 3.1. ACT water samples are characterized by low pH (~4), high DOC (from 57.4 to 64.9 mgC/L), and low concentrations of all analyzed major ions. In contrast, SPG water samples exhibited pH values >7, low DOC concentrations (<2 mgC/L), and high concentrations of major ions, mostly Ca 2+ and SO 4 2 concentrations which were an order of magnitude higher tha n those determined in ACT samples. Overall, no significant differences were observed in the absolute values of parameters analyzed between the two sample collection trips. The excitation emission fluorescence spectra of these samples showed different domin ant groups of DOC compounds, as shown in Figures 3 1 and 3 2, which were obtained to help link the types of organic matter present in water samples to nAg behavior in solution. The contour lines in these graphs represent relative intensities and the contou r intervals are presented on a grayscale from white (highest intensity) to black (no intensity). Based on the approach proposed by Chen et al.  peaks at longer excitation wavelengths (>280 nm) and longer emission wavelengt hs (> 380 nm) are indicative of humic acids, whereas peaks at shorter excitation wavelengths (<250 nm) and longer emission wavelengths (>380 nm) are indicative of fulvic acids. Figures 3 1a and 3 2a show that the ACT water was dominated by humic and most l ikely some fulvic acid like materials which are considered allochthonous or terrestrially derived organic carbon. On the other hand, the SPG waters (Figures 3 1b and 2b) were dominated by aromatic protein like materials, indicated by peaks at shorter exci tation wavelengths (<250 nm) and shorter emission
53 wavelengths (< 350 nm), and soluble microbial by product like materials, indicated by peaks at intermediate excitation wavelengths (250 280 nm) and shorter emission wavelengths (< 380 nm). However, despite the presence of aromatic proteins in SPG waters, the specific UV absorbance (SUVA) value was much higher for the ACT waters than the SPG waters (Table 3 1), indicating a higher percentage of aromaticity in the make up of humic and fulvic acids present in ACT waters  3.3.2. Characterization of Prepared nAg Suspensions 184.108.40.206. Initial particle characterization A s receive d nAg particles were analyzed for PSD and SEM imaging (Figure 3 3). When suspended in NP water the obtained nAg suspension exhibited an average particle size of 145 2.9 nm (Figure 3 3a and Table 3 2) and a zeta potential of 36.8 0.56 mV. The PSD is based on an intensity weighted distribution which is skewed towards the larger particles in the sample because it relates to the intensity of light scattered by the particles. It is worth to note that the second nAg suspension prepared in nanopure water ( NP) and used in experiments discussed below had an average size of 59.06.9 nm and a zeta potential of 26.20.81 mV. For these suspensions, the PSD and average particle size were measured with a Nanotrac Microtrac and the results are based on a number wei ghted distribution which tends to be skewed to the smaller size particles in the sample. 220.127.116.11 Characterization of nAg suspensions Five types of nAg aqueous suspensions were prepared and are referred to as ACT, SPG and NP (nanopure or DI water), MHW (m oderately hard water ; see Appendix A 1 for full chemical composition) and CM (algal culture medium; see Appendix A 2 for full chemical composition). Characterization data for freshly prepared stock nAg
54 suspensions are given in Table 3 2. When comparing the two natural waters (ACT and SPG), it is apparent that nAg are better dispersed in ACT waters than in SPG suspensions. The large average size and small negative net charges of nAg particles in SPG particles are clear indication of the impact of high io nic strength, especially in water containing high calcium concentrations and not enough DOC. Nanosilver suspensions in such waters would become less stable due to increased screening of particle surface charge and ultimately aggregation caused by higher io nic strength as well as potential Ca 2+ bridging with organic molecules an nAg [180 182] Nanosilver particles exhibited rather high average diameter in all synthetic waters, and MHW suspended total Ag as the latter becomes removed from solution via sedimentation. F igures 3 4 gives the PSD of nAg particles suspended in natural waters from the first sampling trip as well as the PSD in MHW and CM waters. The distribution curves fall into roughly 3 regions of the diagram with overlapping portions. Overall, nAg particles suspended in ACT plotted on the lowest end, MHW on the highest end, and both SPG and CM in the middle. SEM images of ACT and SPG suspensions (Figure 3 5) visually illustrate the difference s in the size of nAg aggregates formed in these two different water s as well as suggests DOC coating of the nAg particles in the ACT water 18.104.22.168. Effect of storage on nAg particle surface charge and PSD Nanosilver suspensions prepared in ACT, SPG, and NP waters were stored and analyzed over time for surface net cha r ge and PSD. Zeta potential results (Figure 3 6) for freshly prepared and 1 month old nAg suspensions show the impact of solution chemistry on nAg surface physical properties. The decrease in net negative charge of the particles in the different waters shows the importance of the two end member
55 parameters selected for this study. The organic rich ACT water shows a rather slight decrease in value, while nAg suspended in SPG water were at or near the point of zero charge. Interestingly, nAg suspended in NP wat er seem to parallel nAg ACT suspension. PSDs are provided for freshly prepared suspensions, 2 day old suspensions, and 2 week old suspensions. The three tested plain water types (without nAg) did not give a signal (i.e. loading index) when analyzed via DL S. It should be noted that DLS is not a perfect technique especially for multimodal distributions and for natural waters and thus the following data must be interpreted with caution. Figure 3 7 gives the PSD of nAg suspended in ACT waters. Although slight shifts are observed over time, the PSD appears to be consistent and peaking at approximately 20 nm. Figure 3 8 gives the PSD of nAg NP suspensions, revealing much broader distributions than those in nAg ACT suspensions, but still narrower than the PSD in nAg SPG suspensions (Figure 3 8). For the latter, the change in mean particle size is more drastic over time. The particle size distribution and mean particle size can be presented based on number, area, volume and intensity weighed approaches. The PSDs discussed above are number weighted distributions and the mean particle sizes are calculated from number and vol ume weighted distributions (Figure 3 10). Volume weighted distributions are strongly influenced by large particles as it expresses the size of the particles that comprise the bulk of the sample volume. Conversely, number weighted distributions are weighted toward the smaller particles in the distribution because each particle in the sample is equally weighted regardless of its size. Figure 3 8 shows the mean particle
56 size of volume weighted distributions (MV) and of number weighted distributions (MN). Thes e values were calculated in the Microtrac software. The mean diameter in the nAg ACT suspension was fairly consistent over the time range for both the volume weighted distribution and the number weighted distributions. This data assists in drawing the co nclusion that nAg ACT was stable in comparison to the other tested waters. However, and although expected, there was a difference in sizes calculated by the two statistical methods. For instance, the fresh suspension had an MN of 18.17 nm but a MV of 37.2 9 nm. The MNs for nAg SPG and nAg NP followed a similar trend to each other whereby the initial MN decreased in size by day two and then increased to approximately the same size as in the fresh suspension by day 14. The initial decrease in size can be a ttributed to larger particles or aggregates settling to the bottom of the cuvette. The increase by week 2 is likely due to growth of particle aggregates with time, which ultimately would lead to instable suspensions. The MV values tell a different story for nAg sPG and nAg NP suspensions whereby the SPG suspension particle size increases drastically from day 2 to day 14 and the Ag NP suspension has an initial decrease at day 2 and then stays about the same at day 14. The drastic increase in MV for nAg SP G must be due to large aggregates still in suspension that are not weighted as heavily in number based distributions. Assuming that nAg could undergo dissolution to some extent in these waters, it is worth noting that small AgCl colloids could still be in suspension and therefore included in the sizing data for nAg SPG.
57 Overall, the three suspensions exhibit contrasting stabilities and particle sizes. It is apparent that nAg ACT is the most stable suspension followed by nAg NP and nAg SPG in that order. 3.3.3. Dissolution Kinetics of nAg Suspended in ACT, SPG, and NP Waters Recent findings show that nAg undergoes dissolution when suspended in well aerated waters [110, 167, 170, 173, 175] such as those used in this study. Generally, reports show that the percentage of Ag + released is dependent on the size of the NPs and concentration of Ag. The smaller the NPs the higher percentage of Ag + is released and the l ower the concentration nAg the higher percentage of Ag + is released. However, when dealing with natural waters, there are obviously going to be other factors that could contribute to the dissolution of nAg. Several methods have been explored to separate and measure ionic silver from particulate nAg. Unfortunately, all have advantages and disadvantages. Centrifugal ultrafiltration (Amicon Ultra 15, Millipore) was chosen for the separation of Ag + from nAg particles in this study. In attempts to more thoro ughly decipher the toxicity response of organisms exposed to nAg, dissolution kinetic experiments were conducted as described earlier. Figure 3 11 shows the concentration of Ag + released over time from three nAg suspensions (ACT, SPG, and NP). nAg ACT a nd nAg SPG suspensions released very low concentrations of Ag + over the 9 day period, whereas the nAg NP suspension comparatively released over 25 times more Ag + at any given time point. All suspensions started with the same high concentration of nAg (500 ppm nAg) and Day 1 is 30 minutes after suspensions were placed on the shaker table. In regard to the nAg ACT suspension low release is due to the nAg particles being coated by organic molecules, which can also complex soluble Ag + The filtrate
58 was almost clear with just a hint of yellow indicating that most of the organic molecules were too large to pass the 3 kDa membrane and that any form of Ag was tightly bound/coated by the organic compounds. As for the nAg SPG suspension, the combination of salt sti mulated particle size growth which decrease s rates of nAg dissolution and the high affinity of dissolved Ag+ for Cl explains its low release of Ag + Formatio n of AgCl colloids could result in particles which are too big for the cutoff pore size of used filters. The SPG water had a measured Cl concentration of 7.57 ppm (or 2.14E 4 M), which would react with Ag + and precipitate as AgCl(s) until the concentration of Ag + in solution reaches a saturation point, 2.14 E 4 M (23 ppm Ag). The ACT water also has a significant Cl concentration (6.23 ppm), but it is likely that the large concentration of DOC brings about a comp etition between these two ligands, and in this case, it appears that DOC plays a more pronounced role in interacting with nAg than Cl Table 3 3 depicts the total concentration of Ag in solution and the percent dissolved of the total concentration for three selected time points. The nAg NP suspension released high percentag es of Ag + at each time point (62.6 70.5 % Ag + ). Although nAg SPG and nAg ACT both released very low amounts of Ag + nAg SPG released a higher percentage (5.05 5.08 % Ag + ) of the total Ag in suspension. This is due to the fact that the total amount of Ag in suspension is much lower for the SPG water than the ACT water. Even though it appears that all nAg suspensions have reached a plateau in terms of dissolved Ag concentrations, it is expected that the dissolved fraction and total amount of silver in susp ension will increase significantly for the SPG water over an extended period of time. Likewise, the nAg NP suspension is
59 expected to increase in dissolved Ag and total suspended Ag although to a lesser extent than the SPG suspension. 3.4. Conclusion The r esults presented here are necessary to link nAg particle behavior to potential ecological implications. We have shown that synthetic waters are not suitable proxies for natural waters. For example, Gao et al.  showed a dramatic decrease in stability of bare nAg particles from 10 to 20 ppm SRHA. They postulated that at high concentrations of SRHA (>10ppm) the organic molecules in synthetic water bridged the nAg and caused the particles to destabilize and aggregate. This is an interesting comparison to the ACT water used here to suspend ba re nAg particles in that ACT had a much higher DOC concentration than 20 ppm and it remained a stable suspension for over a month. The differences shown between these two studies is enough to suggest that synthetic waters designed to mimic natural waters with significant DOC concentratio ns may not be appropriate to use for toxicity assays. Additionally the CM used for algal toxicity studies provided similar particle stability as the SPG water suggesting that it may accurately predict toxicity in natural waters similar to the SPG water. However, the M HW produced much less stable suspensions most likely due to its high concentrations of divalent cations which are known to destabilize nanoparticle systems. Therefore this growth medium will likely give different toxicity results than the SPG and ACT wate rs. Based on the dissolution results it is apparent that all three waters (NP, ACT, and SPG) react differently with nAg which in turn will affect the toxicity seen in these waters. The high DOC in ACT overpowers all other ligands and electrolytes by steric aly stabilizing nAg particles and likely scavenging most Ag +
60 surfaces. As for the SPG water nAg dissolution is complicated mainly by the potential interactions with Cl higher pH and increased aggregation It is also possi ble that the small amount of DOC could provide some protection from particle dissolution. Some have shown that nAg dissolution rates are decreased by aggregation in the system [169, 190] and that dissolution decreases as a function of increasing pH  The nAg in the SPG water also likely interacted with Cl present in the system producing AgCl colloids from relased Ag + and producing AgCl complexes on the surface of the nAg particles [174, 177] The nAg NP suspension, meant to represent typical laboratory suspensions, behaved completely different from the natural waters giving high percentages of dissolved Ag + This alone is reason to suggest that synthetic media would also give different results in regard to ecological impacts.
61 Table 3 1 Temporal chemical characterizati on of natural water samples coll ected from Alachua Conservation T rust (ACT) wetland and a spring fed riparian wetland (SPG) along the Ichetucknee River, in North Central Florida, USA. Relevant water chemistry parameters of synthetic waters tested in this study are also included (CM = algal growth media and MHW = moderat ely hard water used as medium for toxicity testing of freshwater invertebrates such as C. dubia ). NA: not applicable, ND: not determined; DOC: dissolved organic carbon; DIC: dissolved inorganic carbon; SUVA: specific UV absorbance at 254nm (m 1 ), and BD: below analytical detection limit Water samples (sampling dates) DOC* (mg/L) DIC* (mg/L) pH Na + (mg/L) Mg 2+ (mg/L) Ca 2+ (mg/L) Cl (mg/L) SO 4 2 (mg/L) SUVA* (Lmg 1 m 1 ) EEM Dominant DOC Type ACT (5/11/2010) 64.9 10.5 4.1 6.04 1.63 4.03 7.76 BD* 4.94 Humic & fulvic acids ACT (8/13/2010) 57.4 ND* 4.3 2.32 1.17 2.32 6.23 0.68 3.31 Humic & fulvic acids SPG (1/22/2010) 1.15 34.7 7.9 4.73 8.75 54.0 6.75 19.0 1.74 Aromatic proteins & microbial by products SPG (3/11/201 1 ) 1.67 ND* 8.1 4.50 7.85 53.2 7.57 17.0 0.96 Aromatic proteins & microbial byproducts CM NA* 2.14 7.5 11.0 2.95 1.20 6.72 5.73 NA* NA* MHW NA* 14.1 7.6 27.0 12.1 14.0 1.90 81.4 NA* NA*
62 Table 3 2 Particle size, total Ag concentrations (Ag T ), electrophoretic mobility, zeta potential ( ), and pH values measured in the different natural and synthetic waters containing nAg particles. nAg suspension media (sample collection dates) Average diameter (nm) Ag T (ppm) Electrophoretic mobility (10 8 m 2 /V*s) (mV) pH ACT (5/11/2010) 76.8 0.4 ** 52.8 1.27 2.17 0.05 28.8 0.58 ND ACT (8/13/2010) 18.2 3.71 32.3 0.55 2.36 0.05 30.2 0.59 5.62 SPG (1/22/2010) 192 5.1 ** 5.03 0.45 0.96 0.07 12.7 0.94 ND SPG (3/11/2011 ) 81.7 11.5 14.0 0.27 0.54 0.1 6.93 1.28 8.42 CM* 174 1.6 ** 2.87 0.61 1.86 0.04 24.8 0.58 ND MHW* 395 50.9 ** 10.1 0.15 0.33 0.04 4.38 0.47 ND NP* 145 2.9 ** 3.71 0.04 2.88 0.04 36.8 0.56 ND NP 59.0 6.85 30.1 0.37 2.05 0.06 26.2 0.81 7.44 *ACT = Alachua Conservation Trust wetland; SPG = Spring fed riparian wetland along the Ichetucknee River; CM = algal growth media; MHW= moderately hard water; NP = Nanopure water ; and ND = Not determined **Average diameter calculated from intensity weighted particle size distributions.
63 Table 3 3 Kinetic of dissolution of nAg particles and production of ionic Ag (Ag + ) in ACT, SPGs, and NP waters. Results are presented as percent dissolved Ag of total Ag concentra tion in the suspensions. Water samples (sampling dates) --------------------------t = 24 hours Contact time (t) ---------------------------t = 72 hours ---------------------------t = 144 hours Ag T (ppm) Ag diss (%) Ag T (ppm) Ag diss (%) Ag T (ppm) Ag diss (%) ACT (8/13/2010) 15.10 1.34 36.6 0.526 36.10 0.67 SPG (3/11/2011 ) 2.36 5.08 ND ND 3.87 5.05 NP 15.40 66.30 17.9 62.60 18.00 70.50 *ACT = Alachua Conservation Trust wetland; SPG = Spring fed riparian wetland along the Ichetucknee River; NP = Nanopure water and ND= Not determined
6 4 Figure 3 1 Excitation emission fluorescence spectra of water samples collected: ( A ) on 5/11/2010 from the Alachua Conservation Trust (ACT) wetland, and ( B ) on 1/22/2010 from a riparian wetland along the spring fed Ichetucknee River (SPG). ( A ) ( B )
65 Figure 3 2. Excitation emission fluorescence spectra of water samples collected: ( A ) from the Alachua Conservation Trust (ACT) wetland on 8/13/2010 ; and ( B ) from a riparian wetland along the Ichetucknee spr ing fed river (SPG) on 3/11/2011 (A) (B)
66 Figure 3 3 Physical analysis of raw nanosilver (nAg) as received from Quantum Sphere, Inc. (Santa Ana, CA). ( A ) Particle size distribution of nAg suspended in Nanopure water determined by dynamic light scattering. The average diameter and zeta potential were 145 2.9 nm and 36.8 0.56 mV, respectively. The PSD data is presented based on intensity. ( B ) Scanning Electron Microscopy image of dry nAg powder. The presence of Ag was verified by energy dispersive X ray spectroscopy (EDS). (B) ( A )
67 Figure 3 4. Example particle size distribution of nanosilver particles (nAg) suspended in:(i) riparian wetland along the Ichetucknee spring fed river water (nAg+SPG) collected on 1/22/2010 (ii) Alachua Conservation Trust (nAg+ACT) wetland collected on 5/11/2010 (iii ) moderately hard water (nAg+MHW), and (iv) algal growth culture medium (nAg+CM). Data were determined by dynamic light scattering and distributions are weighted based on intensity
68 Figure 3 5. Scanning electron microscopy images of nAg suspended in: ( A ) Alachua Conservation Trust (ACT) wetland water collected on 5/11/2010 and ( B ) riparian wetland along the Ichetucknee spring fed river (SPG) water collected on 1/22/2010 The presence of elemental Ag was verified by energy dispe rsive X ray spectroscopy. ( A ) (B)
69 Figure 3 6. Zeta potential of stock nAg suspensions prepared in Alachua Conservation Trust (ACT) wetland water (collected on 8/13/2010 ), nanopure water (NP); and water from the sprin g fed wetland along the Ichetucknee River (SP G; sample collected on 3/11/2011 ). Data illustrated by black bars are from freshly prepared nAg suspensions, and gray bars correspond to data obtained from nAg suspensions left to equilibrate for a month. Error bars represent standard error of the mean.
70 Figure 3 7. Particle size distributions of nanosilver suspension in Alachua Conservation Trust (ACT) wetland water collected on 8/13/2010 ( A ) Fresh suspension (i.e. right off shaker table) ( B ) 2 days stored suspension and ( C ) 2 weeks stored suspension. Size distributions are number based distributions and percent channel represents the percent of particles between two given sizes (A) (B) (C)
71 Figure 3 8. Particle size distribution (PSD) of nAg particles suspended in water samples collected from the Ichetucknee River wetland (SPG) collected on 3/11/2011 ( A ) Fr eshly prepared suspensions, ( B ) PSD after 2 days of equilibration, and ( C ) PSD after a 2 week equilibration period. Size distributions are number based distributions and percent channel represents the percent of particles between two given sizes (C) (B) (A)
72 Figure 3 9. Particle size distribution (PSD) of nAg suspension in nanopure water (NP). ( A ) Freshly prepared suspensions; ( B ) after 2 days of equilibration; and ( C ) after 2 weeks of equilibration. Size distributions are number based distributions and percent channel represents the percent of particles between two given sizes. (C) ( B ) (A)
73 Figure 3 10. Mean particle size of nAg over time in: Alachua Conservation Trust (ACT) wetland water collected on 8/13/20 10 (nAg ACT suspension), Ichetucknee river water collected on 3/11/2011 (nAg SPG suspension); and nanopure water (nAg NP suspension). ( A ) Volume weighed particle size distribution and ( B ) number weighed particle size distribution. Error bars represent th e standard deviation calculated from 3 measurements of the same sample. (A) (B)
74 Figure 3 11. Temporal trends of ionic Ag (Ag + ) concentration resulting from the dissolution of nAg particles suspended in 3 of the tested water types. ACT = Alachua Conservation Trust wetland water colle cted on 8/13/2010 SPG=spring fed wetland along the Ichetucknee River collected on 3/11/2011 NP = Nanopure water. Error bars represent 1 standard deviations of the mean. The inset graph shows the separation of Ag+ concentration trends for ACT and SPG.
75 CHAPTER 4 EFFECTS OF NATURAL WATER CHEMISTRY ON NANOSILVER BEHAVIOR AND TOXICITY TO CERIODAPHNIA DUBIA AND PSEUDOKIRCHNERIELLA SUBCAPITATA 4 .1 Introduction Nanotechnology is at the forefront of research, guiding scientists and engineers to develop efficient energy production and storage methods, novel biomedical applications, smaller computer chips, and strong but light materials, to name a few. This emerging discipline provides the ability to individually address, control, and modify structures, materials, and devices with nanometer precision. The products of nanotechnology are now finding use in a wide variety of human activities, and their environmental imp act could include the release of new classes of toxins with largely unknown environmental and health hazards  Amongst the wide variety of engineered n anoparticles (NPs), nanosilver (nAg) is becoming extremely prevalent in consumer products, so much so that the Woodrow Wilson Project on Emerging Nanotechnologies (http://www.nanotechproject.org  ) developed a database solely for silver nanotechnology and its applications. Over 50% of commercial products that contain NPs are nAg based products  Nanosilver has proven useful in a variety of applications, where the rapid growth in the marketpl ace is due primarily to its antimicrobial activity and cost effectiveness  The fast growth of nanotechnology and the use of nanoproducts have stimulated research on the potential impacts of NPs on both human health and the environment. Research on the potential toxicity of nAg is fast growing. In fact, the toxicity of bulk silver This cha pter is published in Environmental Toxicology and Chemistry, Vol., 31, No. 1, pp. 168 175, 2012
76 com pounds (e.g. the monovalent silver ion) to aquatic organisms is well established Ag is one of the most toxic heavy metals [146, 200] It appears on the U.S. ty pollutant list under the Clean Water Act  variety of organisms including bacteria  algal species and fish [2, 96, 98] and invertebrates  However, r esults from these studies have raised questions as to disso lution of toxic silver ions (Ag + ) in aqueous solutions, or to the direct effects of specific physicochemical properties of the actual nanoparticle  In a study using the algal species Chlamydomona s reinhardtii Navarro et al.  concluded that nAg particles do undergo dissolution to produce Ag + which then causes the observed adverse biological effects in tested model organisms. It is however clear that this is not the sole toxicity pathway and multiple theories on nAg toxicity routes and mechanisms can be found in published studies [92, 99, 112, 115, 116] Laboratory investigation s of the environmental implications of nAg have primarily used synthetic waters (e.g. altered deionized (DI) water), which help control for the specific effects of individual water chemistry parameters (e.g quantity and types of both ions and dissolved or ganic carbon (DOC)). However, altered DI water based experiments fail to take into account the complexity and chemical diversity of natural water matrices. Accordingly, nAg transformation and dissolution patterns observed in synthetic waters may not accura tely reflect its interactions with natural water components, resulting in potentially different biological implications.
77 In the present study, the interaction of nAg with natural waters and the resulting biological effects on selected model aquatic organisms were investigated and compared to trends obtained from synthetic water (e.g. moderately hard water (MHW) and algal growth culture medium (CM)) based nAg suspensions. The ultimate goals for this chapter were to successfully suspend and characterize nAg in natural waters, and then comparatively conduct reproducible toxicity assays using both natural water based nAg suspensions. 4 .2 Materials and Methods 4 .2.1 Collection and Characterization of Test Waters Natural water samples were c ollected in pre cleaned high density polyethylene (HDPE) bottles from two different sites near Gainesville in North Central Florida USA The first sampling site was a freshwater marsh wetland within the Alachua Conservation Trust lands on the Prairie Cre ek Preserve in Gainesville, FL which will be referred to as ACT from here on. The second was a riparian wetland on a spring fed river located in Ft. White, FL (called the Ichetucknee River) which will be referred to as SPG from here on. Water samples wer e first decanted to eliminate large debris through simple sedimentation and subsequently filtered at different cutoffs (e.g. 1.6 m for storage and most experiments and 0.45 m for analytical analyses and Ceriodaphnia dubia (C. dubia experiments)) dependin g on usage. The different water aliquots were then stored at 4 o C pending analysis and use in laboratory experiments. Analyses included, dissolved organic (DOC) and inorganic (DIC) carbon using a Shimadzu TOC V CPH total organic carbon analyzer, pH with an A ccumet AB 15 meter and major ions by ion
78 chromatography (Dionex ICS 3000 IC). All analyses were done within a week of collection and pH was measured throughout experimentation. To investigate the potential effects of the types of organic compounds present in ACT and SPG waters on nAg dispersion and toxicity, ultraviolet visible (UV Vis) absorbance and fluorescence excitation emission (EEM) techniques were used (Hitachi U 2900 UV visible spectrometer and Hitachi F 2500 fluorescence spectrophotometer)  Ultraviolet absorbance at 254 nm (UV 254 ) was measured using a 1 cm quartz cell in order to give an indication of carbon carbon double bonds in the samples. The UV 254 and DOC data were used to calculate the specific UV absorbance (SUVA) va lues where the UV 254 (m 1 ) is divided by the DOC concentration (mg L 1 ). The SUVA values are typically used to estimate the percent DOC aromaticity of a given water sample  To obtain a given EEM spectrum, a diluted water sample was placed in a 1 cm quartz cell and simultaneously scanned at 5 nm increments over an excitation (EX) wavelength range of 200 to 500 nm and at 5 nm increments over an emission (EM) wavelength range of 200 to 600 nm. Raw EEMs were processed in MATLAB (Mathworks)  Deionized water EEMs were subtracted from all sample EEMs and intensities normalized by Raman water area  Contour plots were obtained in MATLAB and compared to a reference plot  to identify the different types of DOC compounds present in each sample. 4 .2. 2 Preparation and C haracterization of nAg S tock S uspensions Nanosilver powder was purchased from Quantum Sphere, Inc. (Santa Ana, CA). Relevant Information on nAg characterization as provided by the vendor were as follows. Nanosilver particles were produc ed by use of a patented vapor condensation process resulting in a >99.9% purity of face centered cubic crystal structures with a bulk
79 density of 0.25 g/cm 3 a melting temperature of 962C, a mean particle size of 25.4 nm and an average specific surface are a (SSA) of 22.5 0.25 m 2 /g. Multipoint Brunauer Emmett Teller (BET) analysis (Quantachrome NOVA 1200) was used to verify the reported SSA. Suspensions of nAg were prepared by dispersing 200 mg of nAg powder into 150 ml of each of the tested waters (i.e. natural and synthetic). Mixtures were shaken for one week and then filtered (1.6 m) to remove the large aggregates  Obtained filtrates were subsequently analyzed to determine the total concentration of Ag (Ag T ) by inductively coupled plasma atomic emission spectroscopy (ICP AES, Perkin Elmer Optima 3200 RL) following a nitric acid digestion procedure adap ted from U.S. EPA Method 3005A  The particle size distribution (PSD), average particle mobility were determined using a Brookhaven ZetaPlus Average particle diameters and PSDs were evaluated via dynamic light scattering (DLS) in which measurements were made at a 90 scattering angle utilizing a HeNe laser at 633 nm PSDs reported are the average of ten runs, each for a duration of 1 min and are intensity weighted distributions Zeta potential and electrophoretic mobility were determined with the same instrument where the electrophoretic mobility was converted to zeta potential via the Sm oluchowski equation within the PALS (phase analysis light scattering) program  Results were the average of 10 runs each of which were comprised of at least 15 cycles. Fresh stock suspensions were pr epared prior to the start of each set of experiments. 4 .2.3 Microscopy Scanning Electron Microscopy (SEM, JEOL JSM6330F) coupled with energy dispersive X ray spectroscopy (EDS) housed in the Major Analytical Instrumentation Center (MAIC) at the Universi ty of Florida were used in the analysis of raw nAg and nAg
80 suspended in tested natural waters. For the analysis of raw nAg approximately one gram of as received nAg powder was mounted on a double sided carbon tab, carbon coated, and then imaged at various magnifications For the water suspensions, a droplet of sample was placed on an aluminum mount, allowed to air dry, carbon coated, and then imaged as above. 4 .2.4 Toxicity Assays To comparatively assess the biological impacts of nAg suspended in different types of wat ers, well established toxicity tests were conducted including the Ceriodaphnia dubia ( C. dubia ) acute toxicity assay and the Pseudokirchneriella subcapitata ( P. subcapitata ) bioassay  Nanosilver suspensions in tested natural waters and relevant culture media (i.e. CM and MHW) were comparatively used in toxicity tests. Prior to nAg testing, preliminary experiments were performed in order to confirm that the two model orga nisms could survive in the natural waters followed by necessary pH adjustments, and then tested again. Preliminary studies were also done to assure that nAg suspended in adjusted natural waters would provide approximately the same particle characterizatio n (i.e. PSD and zeta potential) as nAg suspended in unadjusted natural waters. Nanosilver concentration range finding experiments were conducted to determine toxic levels for each organism in each water type. 4 .2.4.1 Ceriodaphnia dubia bioassays C eriodap hnia dubia cultures were obtained from Hydrosphere Research (Alachua, FL), maintained in 1 L beakers containing approximately 500 ml MHW and kept in an environmental chamber at 25C with constant aeration under a photo period of 16 :8 h light :dark Cultu res were fed every other day with a combination of concentrated algal cells ( P. subcapitata ) and a mixture of yeast, cereal leaves and trout
81 chow (YCT) extract. The C. dubia toxicity assays were performed as 48 hr static bioassays where neonates (< 24 h) were separated from adults and fed (YCT and algae) 2 h prior to toxicity testing. Five neonates were transferred using a small, wide mouth, plastic pipette into a 30 ml plastic cup containing 20 ml of the nAg suspension and were not fed for the duration o f the 48 hr exposure. The end point was mortality and/or immobilization and was determined visually. Controls and six solutions with increasing nAg concentrations were run in triplicates. The LC 50 (i.e. concentration resulting in the death/immobilization of 50% of the population) values and associated 95% confidence intervals (CI) were calculated by probit analysis  The LC 50 s were considered divergent when CIs were not overlapping. 4 .2.4.2 Pseudokirchneriella subcapitata bioassays P suedokirchneriella subcapitata cultures were obtained from Carolina Biological Supply Company (Burlington, NC) and stock cultures initiated as n eeded in culture medium (CM) prepared in accordance with U.S. EPA methods  Stock algal cultures were maintained via aseptic techniques at 25 1 C, under continuous f luorescent lighting of 86 8.6 E/m 2 /s (i.e. Cool White), and shaken twice daily by hand. P. subcapitata was used to evaluate the chronic toxicity of nAg suspended in various types of waters based on the U.S. EPA 96 hr P. subcapitata growth inhibition assay protocol  Following the dilution of the nAg suspensions into respective test waters to produce a nAg concentration gradient, triplicates of each treatment, inclu ding negative controls, were inoculated with the same volume (1 ml) of algal culture. Each test was incubated for 96 h at room temperature under controlled light and shaken twice daily. Algal growth was assessed by chlorophyll a (chl a ) measurement via c hl a
82 filter fluorometer. Chlorophyll a was extracted from glass fiber filters (0.7 m) by first cutting the filters into small pieces, submerging in methanol, and sonicating for 5 min. The supernatant was then analyzed for chl a  Percent inhibition was determined from Equation 4 1. ( 4 1) where C s is the sample chl a concentration and C 0 is the control chl a concentration. Subsequently the IC 50 was determined by plotting nAg sample concentrations versus percent inhibition and performing a regression analysis in the linear portion of the obtained graph using Equation 4 2. ( 4 2) IC 50 is the concentration of nAg which inhibits chl a by 50%, y i is the Y intercept from the linear regression, and S is the slope of the linear regression. 4 .3 Results and Discussion 4 .3.1 Characterization of Test Waters Chemical compositions of the two natural waters used in the present study are given in Table 4 1 The two waters show significant differences in DOC concentrations, pH, and in several ion concentrations (Table 4 1). As typical for a marsh wetland, the or ganic rich ACT water had a lower pH and higher DOC concentration, where as the SPG had circum neutral pH and a much lower DOC concentration (Table 4 1). The SPG water also showed higher values in DIC, Ca 2+ Mg 2+ and SO 4 2 The specific UV absorbance (SUV A) value is much higher for the ACT water than the SPG water indicating that ACT has a higher percentage aromaticity 
83 Flu orescence EEM spectra were obtained for each natural water to help link the types of organic matter present in water samples to nAg behavior. The results are presented in Figure 4 1 Fig ure 4 1a is an EEM fluorescence spectrum of a water sample collected from the ACT wetland. Based on the approach proposed by Chen et al.  peaks at longer excitation wavelengths (>280 nm) and longer emission wavelengths (> 380 nm) are indicative of humic acids, wh ereas peaks at shorter excitation wavelengths (<250 nm) and longer emission wavelengths (>350 nm) are indicative of fulvic acids. Figure 4 1a shows that the ACT water was dominated by humic and fulvic acid like materials (Fig ure 4 1a) which are considered alloc hthonous or terrestrially derived organic carbon. The SPG water was dominated by aromatic protein like materials, indicated by peaks at shorter excitation wavelengths (<250 nm) and shorter emission wavelengths (< 350 nm), and soluble microbial by product like materials, indicated by peaks at intermediate excitation wavelengths (250 280 nm) and shorter emission wavelengths (< 380 nm) (Fig ure 3 1b). Humic acids have the largest molecular weight with the greatest aromaticity of the organic molecules discu ssed above and are known for their ability to reduce metal toxicity [205 207] Terrestrially derived (allochthonous) fulvic acids have a lower molecular weight and have not been investigated to the same extent as humic acids in regards to interactions with metals, but tend to be grouped with humic acids in the literature dealing with dissolved organic matter. Conversely, microbially derived (autochthonous) organic matter has a much lower molecular weight along with a lower percentage of aromaticity. The degree at which type/quality of organic matter effects nanosilver, or any nanomaterial, bioavailability and toxicity to aquatic organisms is
84 unknown. The characterization of the two natural waters (ACT and SPG) in the present study show two clearly different waters in terms of types of organic matter which is likely responsible for some of the differences in toxicity and particle behavior presented below. It should be noted that the concentration of DOC in water dominated by allochthonous sources of organic carbon (ACT water) is considerably higher than that of water dominated by autochthonous organic carbon (SPG water). However, in regards to biological implications, some studies report a decrease in ionic Ag tox icity to aquatic organisms as a result of increasing DOC concentration from only 0 to 2 ppm [176, 208] 4 .3.2 Characterization of P repared nAg S uspensions 4 .3.2.1 Initial particle characterization The SSA of raw nAg determined in our laboratory by multipoint BET analysis prior to use averaged 22.3 m 2 /g, a value that was similar to the 22.5 m 2 /g reported by the manufacturer. A SEM image of raw nAg powder is shown in Figure 4 2a and confirms silver particles on the nanoscale with obvious aggregation. The average or effective diameter of the nAg powder suspended in Nanopure water w as 145 2.9 nm with a polydispersity of 0.233 0.004 and a zeta potential of 36.8 0.56 mV indicating a particle size slightly outside the nanoscale, a fairly large size distribution, and moderately high stability The particle size distribution determ ined by DLS is given in Figure 4 2b Adequate SEM images of the suspension could not be obtained. 4 .3.2.2 Characterization of nAg suspensions Four types of nAg suspensions were prepared and are referred to as ACT, SPG, MHW (moderately hard water), and CM (algal culture medium). The chemical compositions of ACT and SPG waters as determined by analytical techniques listed above are given in Table 4 1 along with the calculated values of used MHW and CM.
85 The full compositions of used MHW and CM have been pub lished elsewhere by the U.S. EPA  The PSD and zeta potentials in prepared nAg suspensions were measured by DLS and the results are presented in Figure 4 3 and Table 4 2 respectively. It is apparent that nAg in ACT water produced the most stable suspension of the 4 water types tested (i.e. lowest PSD range and most negative zeta potential). This can be attributed to the high concentration of DOC present in ACT water. In the suspensions due to steric repulsion forces  Whereas suspensions of to aggregation and their behavior can be predicted by the Derjaguin Landau Verwey Overbeek (DLVO) theory of c olloidal stability. Briefly, as the concentration of electrolytes increases, the attachment efficiency increases due to increased screening of particle surface charge, which in turn reduces the energy barrier and tilts the system towards aggregation [182 184] This tre nd does also impact the actual concentration of Ag in the suspension as measured by ICP AES. The highest silver concentration was measured in the ACT suspension while the other water matrices exhib ited much lower concentrations (Table 4 2) Scanning elect ron microscopy coupled with energy dispersive X ray spectroscopy (SEM EDS) was used to image nAg suspended in ACT and SPG waters. Figure 4 4a and 4 4b are SEM images of nAg suspended in ACT and SPG waters respectively. Nanosilver suspended in the ACT wate r seemingly shows organi c matter (i.e. humic and fulvic acids) coated particles (Fig ure 4 4a). Nanosilver suspended in the SPG water shows large and small aggregates with no evidence of organic matter
86 coating (Fig ure 4 4b). Scanning electron microscopy im ages of MHW and CM suspensions were unsatisfactory most likely due to disturbances during the sample preparation process. 4 .3.3 Biological I mpacts of nAg Ceriodaphnia dubia survived and reproduced in the SPG and pH modified ACT (brought to ~7.5 using 0.1M NaOH) waters and proved to be a good test organism for these experiments. Figure 4 5 presents the concentration response curves for the toxicity of nAg suspended in each of the three waters. It was found that nAg suspended in MHW and SPG water resulted in similar levels of toxicity response, with calculated LC 50 0.523 ppb) and 0.433 ppb (0.424 0.440 ppb), respecti vely. This is interesting because the MHW suspension was less stable and resulted in a larger PSD than in the SPG water, suggesting that the MHW suspension should be more toxic. However, this difference in stability/aggregation did not affect the respons e of the test organism, which may be attributed to C. dubia to better protect against nAg toxicity t han in MHW. When nAg was suspended in ACT water the LC 50 was found to be almost 3 orders of magnitude higher (i.e. 3 times less toxic) than all the other nAg suspensions. The LC 50 and 95% CI determined from ACT based nAg suspensions were 221ppb (209 235 ppb). Based on chemical composition and SEM images, it is apparent that the humic substances in the ACT water were responsible for the observed toxicity reduction. It is difficult to compare nAg toxicity data from individual publications to the results p resented here due to the use of natural waters as the test medium. The closest
87 comparison would be to results previously published by Gao et al.  a study in which similar methods of nAg suspension in natural waters were used. However, no toxicity assays in natural waters were performed; instead growth media were used for the different assays. Using a NP suspen sion water with a similar chemical composition to 50 of 6.18 ppb for C. dubia which is more than an order of magnitude mor e toxic than what was found in the present study. The other waters they tested were more comparable to the SPG water and produced LC 50 s of approximately 0.7 ppb which is slightly higher than the result s of the present study of 0.433 ppb. Conversely, a study using sonicated nAg suspensions in Mill iQ water reported a LC 50 of 67 ppb Ag  a value that is significantly higher than the present suspension. P seudokirchneriella subcapitata survi ved and reproduced in the SPG and pH modified ACT (brought to ~7.5 using 0.1M NaOH). Figure 4 6 presents the biological response of P. subcapitata to increasing concentrations of nAg suspended in the different tested waters. Observed toxicity trends were similar to that of C. dubia however this organism seemed to be less sensitive than C. dubia to nAg exposure. The CM suspensions showed the highest toxicity with an IC50 of 4.61 0.325 ppb. The IC 50 s for nAg SPG and nAg ACT suspensions were 22.6 3.13 ppb and 1600 0.034 ppb respectively. The CM suspension and SPG suspension had a similar PSD but CM showed a higher stability. The higher stability of the CM suspension may be attributed to the lower Ca 2+ concentration which would allow for more Ag parti cles to be suspended and thus a higher toxicity result to P. subcapitata In turn, the lower toxicity seen in the
88 SPG suspension could be a result of the higher concentration of Ca 2+ causing greater aggregation and thus a lower chance of coming into conta ct with suspended algae. Another possibility would be that the small amount of organic matter in the SPG water was enough to protect against nAg toxicity to P. subcapitata but not to C. dubia Even fewer studies exist testing the effects of nAg on P. subc apitata and algal species in general. Using nAg suspended in MilliQ water through sonication, Griffitt et al.  measured an IC 50 of 190 ppb. This value is lower than the one determined for ACT water but higher than IC 50 values obtained with CM and SPG waters. 4.4 Conclusion Results obtained in the present study show the critical importance of water chemistry on the fate and biological impact of nAg Experimental results indicate that natural dissolved organic matter pl ays a major role in the toxicity of nAg. Dissolved organic matter, especially that which is comprised of humic and/or fulvic acids ( as determined by EEM spectra ) significantly reduces the toxicity of nAg, potentially through particle coating as qualified by SEM observations. Currently most published toxicity results on the effect of N Ps on aquatic organisms are obtained by use of altered synthetic waters. Results from this study suggest that accurate predictions of the biological impacts would be obtaine d by the use of natural water containing NPs as growth media to account for the complexity and diversity of water chemical compositions. It was found that when C. dubia was used as the test organism, it could be acceptable to use traditional growth media (i.e. MHW) in place of the water of interest when the chemical compositions are similar. In contrast, toxicity results obtained from the P. subcapitata assay show that this paradigm is not a panacea. For instance, results show that C dubia toxicity in th e SPG water (low DOC and moderate
89 ionic strength) were very similar to that of the traditional growth media (i.e. MHW). Though, with the P. subcapitata even the SPG water produced a significantly lower toxicity impact than the CM. When studying water with a high DOC and low ionic strength, none of the toxicity results were comparable to that of synthetic waters. Because the mechanism of toxicity is of nAg is still not well understood, it is important to explore alternate methods such as the ones presented here in toxicity testing as well as fate and transport studies Results obtained in th is study should help stimulate research and discussion on the actual fate and implications of nAg and NPs in general on aquatic systems. Future research avenues should focu s on using natural waters with clear gradients of key water parameters known to impact the surface properties of nAg and comparatively assessing ionic Ag and nAg toxic effects in natural waters Results presented in this chapter do not only initiate the process of establish ing needed correlations between water matrix dependent nAg particle properties and toxicity implications, but suggest that the use of traditional growth media in toxicity assays concerning N P s may not be appropriate.
90 Table 4 1. Characterization of natural waters collected in North Central Florida, USA from Alachua Conservation Trust (ACT) wetland and a riparian wetland along the Ichetucknee springs (SPGs). Relevant water chemistry parameters used in synthetic growth media (algal culture medium or CM and moderately hard water or MHW for Ceriodaphnia dubi a ) are also presented. BD: belo w analytical detection limit. NA: not applicable DOC: dissolved organic carbon. DIC: dissolved inorganic carbon. SUVA: specific UV absorbance at UV 254 (m 1 ). Table 4 2. Characterization of nAg suspended in different waters: Alachua Conservation Trust (ACT) wetland water and a riparian wetland along the Ichetucknee springs (SPGs) water (both collected in North Central Florida, USA) culture medium (CM), moderately hard water (MHW) and nanopure water Average diameter, mobility, and zeta potential were determined by dynamic light scattering and total Ag concentration (Ag T ) determined by inductively coupled plasma atomic emission spectroscopy nAg suspension medium Average diameter (nm) Ag T (ppm) Electrophoretic mobility(10 8 m 2 /V*s) Zeta potential (mV) ACT 76.8 0.4 52.8 1.27 2.17 0.05 28.8 0.58 SPG 192 5.1 5.03 0.45 0.96 0.07 12.7 0.94 CM 174 1.6 2.87 0.61 1.86 0.04 24.8 0.58 MHW 395 50.9 10.1 0.15 0.33 0.04 4.38 0.47 Nanopure 145 2.9 3.71 0.04 2.88 0.04 36.8 0.56 Sampled Locations DOC (ppm) DIC (ppm) pH Na + (ppm) Mg 2+ (ppm) Ca 2+ (ppm) Cl (ppm) SO 4 2 (ppm) SUVA (Lmg 1 m 1 ) Dominant DOC Type ACT 64.9 10.5 4.1 6.04 1.63 4.03 7.76 BD 4.94 Humic a cids & f ulvic a cids SPG 1.15 34.7 7.9 4.73 8.75 54.0 6.75 19.0 1.74 Aromatic p roteins & m icrobial by product like organic compounds CM NA 2.14 7.5 11.0 2.95 1.20 6.72 5.73 NA NA MHW NA 14.1 7.6 27.0 12.1 14.0 1.90 81.4 NA NA
91 Figure 4 1. Excitation emission fluorescence spectra of water samples collected from the ( A ) Alachua Conservation Trust (ACT) wetland and ( B ) a riparian wetland along the Ichetucknee spring fed river (SPG). (A ) (B)
92 Figure 4 2. Characterization of raw nanosilver (nAg). ( A ) Particle size distribution (PSD) of nAg suspended in Nanopure water using dynamic light scattering (DLS ). The average diameter was 145 2.9 nm and zeta potential was 36.8 0.56 mV. The PSD data is presented based on intensity. ( B ) Scanning Electron Microscopy (SEM) image of dry nAg powder as received from manufacturer. The presence of Ag was verified by energy dispersive X ray spectroscopy (EDS). Scale bar = 100 nm. (B) (A)
93 Figure 4 3. Dynamic light scattering (DLS) determination o f p article size distributions (PSD) of nanosilver (nAg) particles su spended in four different water types used in this study : (a) riparian wetland along the Ichetucknee spring fed river water (SPG), (b) Alachua Conservation Trust (ACT) wetland, (c) moderat ely hard water (MHW), and (d) culture medium (CM). PSDs are intensity weighted distributions.
94 Figure 4 4. Scanning Electron Microscopy  images of nAg suspended in ( A ) Alachua Conservation Trust (ACT) wetland water and ( B ) riparian wetland along the Ichetucknee spring fed river (SPG) water. The presence of Ag was verified by energy dis persive X ray spectroscopy (EDS). Scale bars are 1 m and 100 nm for ( A ) and ( B ) respectively (A) (B)
95 Figure 4 5. Dose response curves for the toxicity of nanosilver (nAg) particles to Ceriodaphnia dubia suspended in ( A ) Alachua Conservation Trust (ACT) wetland water, ( B ) riparian wetland along the Ichetucknee spring fed river (SPG) water, and ( C ) moderately h ard water (MHW). Percent survival values are presented as the mean standard deviation The short dashed lines represent the LC 50 s calculated by probit analysis and are ( A ) 221 ppb (209 235 ppb) ( B ) 0.433 ppb (0.424 0.440 ppb) and ( C ) 0.482 ppb (0.455 0.523 ppb). Values in parentheses are 95% confidence intervals. (A) (B) (C)
96 Figure 4 6. Dose response curves for the toxicity of nanosilver (nAg) particles to Pseudokirchneriella subcapitata suspended in ( A ) Alachua Conservation Trust (ACT) wetland water, ( B ) riparian wetland along the Ichetucknee spring fed river (SPG) water, and ( C ) culture medium (CM) water. A lgal vitality is presented as chlorophyll a concentration mean standard deviation Calculated IC 50 values are as follows: ( A ) 1600 0.034 ppb ( B ) 22.6 3.13 ppb and ( C ) 4.61 0.325 ppb, and are represented by the dashed line (C) (A ) (B)
97 CHAPTER 5 INTERACTION OF WATER TRANSFORMED NANOSILVER PARTICLES WITH AQUATIC ORGANISMS: LINKING SOLUTION CHEMISTRY TO OBSERVED BIOLOGICAL RESPONSES 5.1 Introduction The applications and implications of n anosilver (nAg) particles are hot conversation topic s among re searchers, industry, and regulatory agencies. In terms of ecosystem functions and even human health there are two main questions that swarm nAg : (i) is nAg more toxic than its ionic counterpart (Ag + ) ? A nd (ii) is the toxicity of nAg solely due to Ag + dis s olution? These two questions have been addressed in a variety of ways by researchers [92, 98, 109, 111, 137, 147, 154 158] A review of the literature shows that most published findings poin t to the oxidation of nAg particles and dissolution of Ag + as a main pathway leading to the observed adverse effects on exposed aquatic organisms However, it is still unclear as to exactly how nAg particles would affect the biology when the chemical compo sition of receiving waters varies from one aquatic system to another. There have been studies using synthetic waters including spiking de ionized water with monovalent and divalent sa lt compounds or organic compounds such as Suwanee River humic acids (SRHA ) to mimic natural waters [114, 155, 162, 178, 189] Overall, these studies tend to focus on the dissolution, aggregation, and toxicity of nAg particles and present the advantage of pinpointing in very simple systems the role of each of the key water parameters. On the other hand, the experimental approach common to these studies fails to take into account the complexity and variability of natural water chemistry. Unfortunately, a much less number of published papers deal with the fate and exposure of nAg particles in actual
98 natural wa ters [139, 159, 192] Additionally, dose exposure studies mimic king in situ conditions are still lacking. As a continuation of research effort presented in Chapter 4, the study presented in this chapter compares the toxicity of nAg particles and AgNO 3 on model organisms grown in two different types of natural waters, described in earlier chapters as ACT and SPG. Additionally, synthetic water spiked or not spiked wit h model organic compounds are used in parallel experiments to determine differences/similarities, if any, on how water transformed nAg particles would interact with aquatic organisms. Finally, the role of water chemical composition on nAg dissolution and s ubsequent toxicity is also examined. 5 .2 Materials and Methods Most of the steps described in this section are similar to those already described in Chapter 3, and to some extent in Chapter 4. This approach is used simply to guarantee the clarity of pres entation. 5 .2.1 Collection and Characterization of Test Waters 22.214.171.124 Sample collection and handling Natural water samples were collected in pre cleaned high density polyethylene (HDPE) bottles from two different sites near Gainesville in North Central Florida, USA The first sampling site i s a freshwater marsh wetland within the A lachua C onservation Trust lands on the Prairie Creek Preserve in Gainesville, Florida. This sampling site will be referred to as ACT from here on. The second site is a ripari an wetland on a spring fed river, the Ichetucknee River, located near Ft. White, Florida. This site will be referred to as SPG from here on. The ACT water was collected on September 13, 2010 and the
99 SPG water was collected on March 15, 2011. In comparison to the waters used for Chapter 4 : ACT was sampled on May 11 2010 and SPG on January 22 2010. After collection, w ater samples were decanted to eliminate large debris through simple sedimentation and subsequently filtered at different cutoffs (e.g. 1.6 m for storage and most experiments and 0.45 m for analytical analyses and Ceriodaphnia dubia (C. dubia experiments ) depending on usage. The different water aliquots were then stored at 4 o C pending analysis and use in laboratory experiments. 126.96.36.199. Sample analysis The chemical composition of the collected water samples included the determination of dissolved organic (DOC) using a Shimadzu TOC V CPH total organic carbon analyzer, pH with an Accumet AB 15 meter and major ions by ion chromatography (Dion ex ICS 3000 IC). All analyses were done within a week of collection and pH was measured throughout experimentation. Additionally, ultraviolet visible (UV Vis) absorbance and fluorescence excitation emission (EEM) techniques were used (Hitachi U 2900 UV visible spectrometer and Hitachi F 2500 fluorescence spectrophotometer) to qualitatively gain insight into the different types of organic compounds present in these water s amples  UV absorbance at 254 nm (UV 254 ) and DOC data were soug ht to allow the determination of specific UV absorbance or SUVA which is used to estimate the degree of aromaticity of dissolved organic molecules  Finally, the EEMs of tested waters were obtained first by diluting the water sample s, placing them in a 1 cm quartz cell and simultaneously scanned at 5 nm increments over an excitation (EX) wavelength range of 2 20 to 450 nm and at 5 nm increments over an emission (EM) wavelength range of 300 to 600 nm. Obtained r aw EEMs were processed using MATLAB (Mathworks)
100  For these analyses, d eionized water EEMs were subtracted from all sample EEMs and the intensiti es normalized by Raman water area  Contour plots were obtained in MATLAB and compared to a reference plot  to identify the different ty pes of DOC compounds present in each sample. EEMs were obtained for pH adjusted ACT, non adjusted ACT, pH adjusted SPG, non adjusted SPG, humic acid sodium salt, and L tyrosine waters and solutions. Synthetic waters including humic acid sodium salt (S igma Aldrich) and L tyrosine ( Sigma Aldrich ) were used in several experiments to compare toxic responses to those of tested natural waters Stock solutions were prepared with thorough mixing followed by filtration through 0.45m membrane filters and then stored at 4C. Stock solutions were analyzed for DOC using methods detailed previously. 5 .2.2 Preparation and Characterization of nAg a nd AgNO 3 Stock Suspensions Nanosilver powder was purchased from Quantum Sphere, Inc. (Santa Ana, CA). Information provided by the vendor were as follows. Nanosilver particles were produced by use of a patented vapor condensation process resulting in a >99 .9% purity of face centered cubic crystal structures with a bulk density of 0.25 g/cm 3 a melting temperature of 962C, a mean particle size of 25.4 nm and an average specific surface area (SSA) of 22.5 0.25 m 2 /g. Multipoint Brunauer Emmett Teller (BET) analysis (Quantachrome NOVA 1200) was used to verify the reported SSA. Suspensions of nAg were prepared by dispersing 300 mg of nAg powder into 150 ml of each of the tested waters (i.e. natural and synthetic). Mixtures were shaken for one week in the dark and then filtered (1.6 m) to remove the large aggregates  Obtained filtrates were subsequently analyzed to determine the total concentration of Ag (Ag T ) by inductively coupled plasma atomic emission spectroscopy (ICP AES, Perkin Elmer Optima 3200
101 RL) following a nitric acid digestion procedure adapted from U.S. EPA Method 3005A  The particle size distribution (PSD) and average particle diameter were determined via dynamic light scattering ((DLS) Microtrac Nanotrac ). Average particle diameters and PSDs were measured with a Microtra c Nanotrac Ultra particle sizer w h ere measurements were made at a 180 scattering angle utilizing a semiconductor laser at 780 nm directed to the sample through a fiber optic cable The PSDs reported are the average of three runs. The average PSDs are generated automatically from the individual runs by the software. Individual runs collected data for one minute each. using a Bro okhaven ZetaPlus and the phase analysis light scattering (PALS) technique T he electrophoretic mobility was converted to via the Smoluchowski equation within the PALS Zeta Potential Analyzer Sofware  Results were the average of 10 runs each of which were comprised of at least 15 cycles. In aims to compare the effects of bulk silver versus nanosilver, similar experiments were run for both nAg and AgNO 3 A 50 ppm AgNO 3 (3% nitric acid (HNO 3 )) stock solution was prepared from crystalline silver nitrate (Fisher Scientific) If stock solutions were needed at lower concentrations of Ag + they were prepared fresh from the 50 ppm stock for each experiment. The concentration of the 50 ppm AgNO 3 st ock was verified several times via ICP AES as were several lower concentrations used in experiments. 5 .2. 3 Toxicity Assays To comparatively assess the biological impacts of nAg suspended in different types of waters and AgNO 3 dissolved in those waters well established toxicity tests were conducted including the Ceriodaphnia dubia ( C. dubia ) acute toxicity assay and
102 the Pseudokirchneriella subcapitata ( P. subcapitata ) bioassay  Nanosilver suspensions in t ested natural waters, synthetic waters, and relevant culture media (i.e. CM and MHW) were comparatively used in toxicity tests. Prior to nAg testing, preliminary experiments were performed in order to confirm that the two model organisms could survive in the natural waters (ACT and SPG) followed by necessary pH adju stments, and then tested again. Likewise, organism survival was tested for a range of concentrations of humic acid and tyrosine. Dilutions of humic acid and tyrosine were made with necessary culture media. The organisms survived adequately in waters prepared with these two model organic compounds however they did not survive in tannic acid or tryptophan solutions, therefore these two compounds were not select ed for experiments. Nanosilver and AgNO 3 concentration range finding experiments were conducted to determine toxic levels for each organism in each water type. 5 .2. 3 .1 Ceriodaphnia dubia bioassays Ceriodaphnia dubia cultures were obtained from Hydrosphere Research (Alachua, FL), maintained in 1 L beakers containing approximately 500 ml MHW, and kept in an environmental chamber at 25C with constant aeration under a photo period of 16:8 h light:dark. Cultures were fed every other day with a combination of concentrated algal cells ( P. subcapitata ) and a mixture of yeast, cereal leaves and trout chow (YCT) extract. The C. dubia toxicity assays were performed as 48 hr static bioassays where neonates (< 24 h) were separated from adults and fed (YCT and algae) 2 h prior to toxicity testing. Five neonates were transferred using a small, wide mouth, plastic pipette into a 30 ml plastic cup containing 20 ml of the nAg suspension or AgNO 3 solution and were not fed for the duration of the 48 hr exposure. The end po int was mortality and/or immobilization and was determined visually. Controls and five
103 solutions with increasing nAg or AgNO 3 concentrations were run in quintuplicates The LC 50 (i.e. concentration resulting in the death/immobilization of 50% of the popu lation) values and assoc iated 95% confidence intervals (CI) were calculated by probit analysis  The LC 50 s were considered divergent when CIs were not overlapping. Model organic compounds (humic acid and tyrosine) were used to explore the change in organic carbon concentration at a set Ag (nAg and AgNO 3 ) concentration and the effects on C. dubia survival. Controls and five solutions with increasing humic acid or tyr osine concentrations were run in quintuplicates The control had neither treatment (i.e. humic acid or tyrosine) nor silver (i.e. nAg or AgNO 3 ). The concentration ranges of humic acid and tyrosine went from 0 10 ppm carbon, in which all were spiked to a concentration of 1 ppb Ag, either as nAg or AgNO 3 ). The nAg treatments were spiked from the NP nAg suspension. C. dubia were exposed as described above. 5 .2. 3 .2 Pseudokirchneriella subcapitata bioassays Psuedokirchneriella subcapitata cultures were obta ined from Hydrosphere Research (Alachua, FL) and stock cultures initiated as needed in culture medium (CM) prepared in accordance with U.S. EPA methods  Stock algal cu ltures were maintained via aseptic techniques at 25 1 C, under continuous fluorescent lighting of 86 8.6 E/m 2 /s (i.e. Cool White) aerated, and shaken twice daily by hand. P. subcapitata was used to evaluate the chronic toxicity of nAg suspended in various types of waters based on the U.S. EPA 96 hr P. subcapitata growth inhibition assay protocol  Fol lowing the dilution of the nAg suspensions and AgNO 3 stock solutions into respective test waters to produce a Ag concentration gradient, quadruplicates of each treatment, including negative controls, were inoculated with the same volume (1 ml) of algal cul ture. Each test was incubated for 96 h at room temperature under
104 controlled light and shaken twice daily. Algal growth was typically assessed by direct chlorophyll a (chl a ) measurement via digital filter fl uorometer. In previous experiments a chlorophyll extraction technique was developed and implemented. However, for most water types the chl a results for extracted vs. non extracted samples had a 1:1 correlation. Circumstances that did not yield a 1:1 corre lation were ones where the concentration of organic matter was changed, thus the extraction procedure was used in these experiments. Extraction procedure is as follows. Chlorophyll a was extracted from glass fiber filters (0.7 m) by first cutting the filt ers into small pieces, submerging in methanol, and sonicating for 5 min. The supernatant was then analyzed for chl a  Percent inhibition was determined from Equation 5 1. ( 5 1) where C s is the sample chl a concentration and C 0 is the control chl a concentration. Subsequently the IC 50 was determined by plotting nAg or AgNO 3 sample concentrations versus percent inhibition and performing a regression analysis in the linear portion of the obtained graph using Equation 5 2. ( 5 2) IC 50 is the concentration of nAg which inhibits chl a by 50%, y i is the Y intercept from the linear regression, and S is the slope of the linear regression. Model organic compounds (humic acid and tyrosine) were used to explore the change in organic carbon con centration at a set Ag (nAg and AgNO 3 ) concentration and
105 the effects on P. subcapitata survival. Controls and five solutions with increasing humic acid or tyrosine concentrations were run in quadruplicates The control had neither treatment (i.e. humic acid or tyrosine) nor silver (i.e. nAg or AgNO 3 ). The concentration ranges of humic acid and tyrosine went from 0 10 ppm carbon, in which all were spiked to a concentration of 1 0 ppb Ag, either as nAg or AgNO 3 ). The nAg treatments were spiked from the NP nAg suspension. P. subcapitata were exposed as described above. 5 .3 Results and Discussion 5 .3.1 Characterization of Test Waters Chemical compositions of the two natural waters used in the present study are given in Table 3 1 and discussed previously i n Chapter 3 The ACT water was collected on September 13, 2010 and the SPG water was collected on March 15, 2011. Fluorescence EEM spectra (Figures 5 1, 5 2, and 5 3) were obtained for each natural water and model organic compounds to help link the types of organic matter present in water samples to nAg behavior. The contour lines in these plots represent relative intensities and the contour intervals are presented on a grayscale from white (highest intensity) to black (no intensity). Figures 5 1a and b a re EEM fluorescence spectrums of a water sample collected from the ACT wetland. Figure 5 1a is the unadjusted ACT water and figure 5 1b is the ACT water adjusted to a pH of 7.5 which is the pH at which toxicity assays were run. Figure 5 1 shows that the A CT water was dominated by humic and most likely some fulvic acid like materials which are considered allochthonous or terrestrially derived organic carbon. There is no functional difference between the unadjusted and adjusted ACT waters except for a smal l change
106 matter makeup. The SPG water was dominated by aromatic protein like materials and soluble microbial by product like materials (Figure 5 2 ) with the most inten se peaks being in those regions. Again there was some variation in the EEM intensities expressed by the unadjusted and pH adjusted SPG water but these differences do not imply a shift in organic matter composition. Figure 5 3 gives EEMs of the two model organic compounds that were implemented in various experiments: humic acid (Figure 5 3a) and tyrosine (Figure 5 3b). The fluorescence peaks for humic acid are well within the defined range for humic acids in general and wh en compared to the EEM for ACT, the highest intensity peak is roughly at the same location (EM ~ 450nm, EX~260 nm). The fluorescence intensities expressed in the tyrosine water (Figure 5 3b) were extremely high even after many dilutions. It is hypothesized that the peaks seen in the right hand corner of the spectrum (EM 600nm) are simply interferences from the high intensity expressed on the far left of the spectrum (EM 300). Because the fluorescence intensity was so high for the tyrosine, the contour lines appear to stop or be cut off. This is due to the slope being too steep from the intensity readings. Regardless of this high intensity and that the scanned range for emission is not large enough, the tyrosine peaks are in fact where aromatic amino acid peak s would be expected. Generally protein peaks should be in an emission range of 280 380 nm and excitation range of 200 320 nm. The tyrosine spectrum shown here is within this range, giving two major peaks within EM 300 350 nm and EX 200 290 nm. Natural organic matter (NOM) can be comprised of many different types and sizes of organic carbon molecules which will all affect the behavior of contaminants via different pathways. In regards to nAg, research has been picking up momentum on the
107 ef fects of organic matter on behavior and toxicity. Most published studies use synthetic waters with the addition of a purchased NOM such as Suwanee river humic acid (SRHA) to investigate the effects of NOM on nAg. A few reports have surfaced recently usin g natural waters to perform nAg fate and toxicity studies; however none have attempted to characterize the type of organic matter in their system as we have done here [139 185, 192] The two natural waters under consideration in this study ( i.e., ACT and SPG) are comprised of two clearly different types of organic matter, allochthonous and autochthonous derived organic matter respectively. It should be noted that the ACT water has a much higher DOC concentration than the SPG water which will obviously impact the toxicity and particle behavior presented below. However, as d iscussed previously amelioration of Ag + toxicity has been reported at DOC concentrations as low as 2ppm [176, 208] The model organic compounds, humic acid and tyrosine, used in this study showed similar EEMs to the respective natural waters ACT and SPG. Based on these results more toxicity assays were carried out with nAg and AgNO 3 using the model organic compounds to see if they would give the same results as the tested natural waters. 5.3.2. Characterization of nAg S uspensions U sed in T oxicity A ssays Three types of nAg suspensions were prepared and are referred to as ACT, SPG and NP. Moderately hard water (MHW) and algal culture medium (CM) described in Chapter 3 were also used (See Tables 3 1 and 3 2 in Chapter 3). Characterization of initial stock suspensions is given in Table 3 2. Zeta potentials for each suspension prepared in ACT, SPG and NP water samples are given in Figure 3 6, where data is presented for freshly prepared suspensions and 1 month old suspensions. There is an
108 obvious trend in stability of nAg ACT>nAg NP>>nAg SPG. When comparing the two natural waters (ACT and SPG), it is apparent that the nAg ACT suspension is more stable due to organic matter coating resulting in steri c repulsion whereas the nAg SP G suspension is much less stable due to increased screening of particle surface charge and ultimately aggregation caused by higher ionic strength as well as aggregation caused by the likely formation of AgCl colloids. All suspensions showed a decrease in stability (i.e. less negative zeta potentials) after one month of sitting undisturbed. The stored nAg SPG suspension had a zeta potential of almost 0 mV indicating that all particles aggregated and settled to the bottom o f the cuvette. The PSD data determined for all prepared nAg suspensions in different waters are presented an d discussed in Chapter 3. 5 .3.3 Biological I mpacts of nAg and AgNO 3 Toxicity assays were conducted using P. subcapitata and C. dubia using water s amples dosed with silver added as either nAg or AgNO 3 For clarity of presentation, t he results are split into two sections In the first section the biological response of tested organisms grown directly into the natural waters and growth media to the a bove two toxicants are presented. Second, the responses of organisms to toxicants in synthetic water samples containing increasing concentrations of the two model organic compounds (i.e. tyrosine or humic acids) are discussed. 5. 3.3.1 Exposure to silver a nd biological responses of organisms grown in MHW growth medi um and in the two n atural waters Using C. dubia the i mpacts of nAg suspended in MHW growth medium on one hand (Figure 5 4 ) and in SPG and ACT waters on the other (Figures 5 5 and 5 6 ) showed similar toxicity response trends.
109 In waters spiked with nAg particles, the synthetic growth medium MHW and the natural SPG water produced similar toxicit y responses from C. dubia with LC 50 values of 0.482 ppb Ag and 0.519 ppb Ag respectively In contrast, suspension s prepared in the ACT water exhibited a much lower toxicity and therefore, a much higher LC 50 of 37.9 ppb Ag. Comparatively, waters spiked with AgNO 3 showed the following toxicity trends. In ACT water, AgNO 3 exhibited a higher toxi city on C. dubia (LC 50 = 5.50 ppb Ag) than the nAg suspension and in SPG water AgNO 3 toxicity was lower than that of nAg with a LC 50 of 1.24 ppb A comparison plot ( Figure 5 7) shows toxicity data in terms of LC 50 in nAg and AgNO 3 spiked waters Given the fact that C. dubia is exposed to nAg particles by both ingestion and external contact, nAg toxicity should be high in water suspensions that favor both ideal dispersion (limiting the removal of aggregated particles from solution via sedimentation) and dissolution to produce the toxic Ag+. The lower toxicity of nAg in ACT water could be due to the high affinity of silver for dissolved organic matter which is present in a rather high concentration in this specific water. It is likely that the organic coat ing of nAg particles would slow down the rate of nAg dissolution while isolating the organisms from direct contact with nAg surfaces at the same time. To some extent, the role of dissolved organic matter in mitigating Ag toxicity to C. dubia is validated by the slightly higher LC 50 measured in SPG water as compared to the fully inorganic MHW. A general look at the LC 50 values shows low values in waters with no or low DOC and high values in organic rich water. Therefore, the increase in salt concentrations could result in increased toxicity response as long as the ionic strength of the solution is not strong enough and the salt levels are below the critical coagulation concentrations (CCC). In such cases,
110 formed aggregates are likely on the lo w to mid range size of the nanoscale realm, and therefore, not quickly removed by sedimentation and prone to uptake by the tested particle feeder, C. dubia Two additional C. dubia toxicity assays were conducted using one month aged suspensions of nAg ACT and nAg SPG (Figure 5 8 ). The LC 50 for the nAg SPG suspension (0.585 ppb Ag) was virtually the same as LC 50 ly prepared suspensions, indicating that filtering out coarse nAg partic les led to stable nAg SPG suspension s for duration te sted in this study The nAg ACT suspension was slightly more toxic (LC 50 = 18.4ppb Ag) than reported earlier, although still on the same order of magnitude. Based on nAg dissolution kinetics discussed earlier in Chapter 3, it appears that the prepared su spensions reach equilibrium rather quickly in terms of Ag dissolution. Accordingly, the toxicity impact corresponding to the effect of dissolved Ag should be the same in 1 week old suspensions used to conduct dose exposure studies discussed earlier and in those conducted with 1 month old suspensions. Correspondingly the drop in LC 50 (i.e. increased toxicity) in ACT water could be an indication of an additional toxicity pathway other than Ag + induced adverse biological effects. The lack of such change in su spensions prepared in SPG water suggests that DOC is likely a key parameter in processes taking place in ACT water. As a particle feeder, C. dubia offers the advantages to study a combination of exposure pathways, namely ingestion and direct contact. In contrast, the use of the freshwater green algae P. subcapitata excludes the ingestion pathway. Also, synthetic water used as suspension media is the algal growth medium (CM) shown in Appendix B
111 from Chapter 3. Experiments similar to those discussed above but conducted using P. subcapitata led to the following observations. For algal cells growing in the culture media (CM) spiked with either nAg or Ag NO3, Ag concentration leading to 50% growth inhibition (IC 50 ) is 4.61 ppb and 6.08 ppb respectively (Figure 5 9 ). These endpoint values are an order of magnitude higher than those determined in the synthetic MHW used for growth of C. dubia In fact, this t ranslates into Ag being less toxic to the tested algal species than to C.dubia in their respective growth media. It is likely that the exposure pathways play a significant role in differences observed in these biological responses. The toxicity response o f P. subcapitata in both SPG and ACT waters are substantially lowered when compared to C. dubia resulting in IC 50 values which are comparatively very high (Figures 5 1 0 and 5 1 1 ). Figure 5 1 2 summarizes the IC 50 s determined for P. subcapitata when exposed to different Ag forms and in waters of different chemical composition. Overall, the combination of toxicity responses exhibited by the two model organisms and differences in the chemical composition of used waters can be used to help elucidate the mechan ism of Ag interaction as a function of organisms and solution chemistry. IC 50 data (algal based data) show a clear indication of toxicity migration with increasing concentrations of DOC, but with AgNO 3 being more toxic than nAg, except in CM suspensions. F or LC 50 ( C. dubia data), the toxicity trend was still more in line with DOC concentrations instead of the ionic strength. Based on these experimental results, it can be concluded that concentrations of dissolved organic matter are the main driver of toxici ty responses of P. subcapitata and C. dubia exposed to silver. Additionally, the
112 type of the two organic matter tested in this study (e.g. humic acids or protein like compounds) does not seem to make a difference Finally, no clear trend was observed between toxicity response and ionic strength. The much higher toxicity of the tested silver compounds to C. dubia as compared to P. subcapitata is likely due to differences in exposure pathways as mentioned previously. 5 .3.3.2 Effects of model organic comp ounds on nAg and AgNO 3 toxicity H umic acid and tyrosine were used as model organic compounds to verify the effect of increasing DOC concentration and DOC type on silver toxicity to both C. dubia and P. subcapitata The dose response curves for C. dubia e xposed to 1ppb AgNO 3 are given in Figure 5 1 3 showing that for both model organic compounds toxicity is mostly ameliorated after reaching a concentration of 0.34 ppm tyrosine or humic acid. Because both model organic compounds showed the same trend and be cause there was some experimental difficulty with daphnia survival in tyrosine, nAg experiments were limited to humic acid. Figure 5 1 4 gives dose response curves of C. dubia exposed to 1 ppb nAg (nAg NP suspension) in increasing concentrations of humic ac id. Two suspensions were used: a fresh nAg NP suspension and a 4 month aged nAg NP suspension. The later suspension was on the shaker table in the dark for 4 months after the initial filtration step. For the fresh suspension minimal toxicity was seen at al l concentrations of humic acid suggesting that the dissolved fraction is not high enough to induce C. dubia death. The aged suspension showed highly variable toxicity up to humic acid concentrations as high as 0.68 ppm At 3.4 ppm humic acid toxicity was reduced and by 6.8 ppm humic acid virtually no toxicity was observed. In comparison to the fresh suspension it is apparent that this toxicity was induced by Ag + released from
113 the nAg. Additionally, for the aged nAg suspension and AgNO 3 humic acid binds Ag + and protects the organisms at concentrations as low as 3.4 ppm and 0.34 ppm humic acid respectively. This reinforces the toxicity trends seen for nAg ACT suspensions which have a much higher concentration of DOC. The question then presents itself, wh y is nAg SPG toxicity higher than nAg NP (>1ppb Ag for fresh suspension) if the dissolution (section 4.3.4) of nAg is much higher for the nAg NP suspension? One plausible answer is that for the SPG water nAg particles contributed to C. dubia toxicity more than the Ag + fraction due to ingestion. Tyrosine also showed an ameliorative effect for AgNO 3 starting at 0.34 ppm. Based on the EEM results tyrosine could serve as a model organic compound with respect to the SPG water which had a DOC con centration of 1.67 ppm. The LC 50 of the AgNO 3 SPG solution was slightly above the 1ppb Ag (LC 50 = 1.24 ppb Ag) benchmark of this tyrosine experiment, so one could deduce that the organic matter had a protective effect in the SPG water for AgNO 3 When C. dubia were exposed to a concentration gradient of nAg NP suspension in 1 ppm tyrosine an LC 50 of 1.85 ppb Ag was observed (data not reported). However the LC 50 of the nAg SPG suspension was below 1 ppb (LC 50 = 0.519 ppb) suggesting that tyrosine protects from toxicity more than the organic matter in SPG water. This can be explained by the difference in the chemical complexity of tyrosine spiked nanopure water versus the natural SPG water which includes several dissolved components in addition to DOC compo unds. It is worth noting that there was some Cl (1.90 ppm) in the MHW growth medium but based
114 reached approximately 362 pp b Ag + Theoretically 1 ppb is too low of a concentration to induce formation of AgCl as a colloid. Similar experiments were carried out for P. subcapitata whereby toxicity assays were performed by expos ure to 10 ppb nAg NP while increasing the concentrations of tyrosine or humic acid ( Figure 5 1 5 ). Overall, the r esults show an initial growth inhibition which disappears progressively with increasing concentrations of tested organic compounds. Only humic acid was used to investigate the effect of increasing organic compound concentration of AgNO 3 toxicity (Figure 5 16 ). Again, this experiment confirms the ability of dissolved organic carbon to mitigate the toxicity of silver. Like with MHW the CM contains some Cl (6.72 ppm) which should not be overlooked. Thermodynamically, 102 ppb of Ag + would need to be added to the system before precipitation of AgCl colloids could begin. 5 .3.4 Significance of d issolution Kinetics in nAg induced toxicity in natural and synthetic waters Recent findings in the literature show that nAg undergoes dissolution wh en suspended in aerated waters [110, 167, 170, 173, 175] Reported data show that the percentage of Ag + released is dependent on the size of nAg particle s and concentration. The smaller the particle size the higher the percentage of Ag + released. When dealing with natural waters several parameters need to be taken into account as they might all have control on the dissolution of nAg particles Several methods have been explored to separate and measure ionic silver from nAg, all of which have advantages and disadvantages  Centrifu gal ultrafiltration (Amicon Ultra 15, Millipore) was chosen for the separation of Ag + from nAg particles in this study as described in Chapter 3
115 Figure 5 17 shows the concentration of Ag + released over time from three nAg suspensions ; two natural water samples ( ACT and SPG ) and nanopure water (NP) as synthetic water. All prepared suspensions started with the same initial concentration of 500 ppm added as nAg particles. Day 1 sample was collected after 30 minutes of equilibration The results show that nAg ACT and nAg SPG suspensions released a very low concentration of Ag + over the duration of the experiment (a 9 day time frame ), whereas the nAg in nanopure water suspension released over 25 times more Ag + at any given time point. This observation poin ts to the significance of water chemistry in controlling the dissolution of nAg. At the same time, these results tend to suggest that toxicity in complex water matrices may not be attributed solely to nAg dissolution. 5 .4 Conclusion The results presented here are compelling, showing that DOC is the main driver of both the fate of nAg and of the biological responses induced by nAg suspensions in natural freshwater s of different chemical composition These r esults were verified by use o f model organic compounds in simple synthetic waters. At the same time these findings show ed that model organic compounds added to DI water cannot always sufficiently represent the complex nature of natural organic matter. Additionally, coupling dissolut ion trends with toxicity results suggests that dissolution of nAg may not be the only cause of observed toxic responses when dealing with natural waters being that the dissolved Ag fraction was rather small Studies assessing the potential contribution of particle related toxicity mechanisms such as the presence and quantity of reactive oxygen species are necessary.
116 Figure 5 1 Excitation emission fluorescence spectra of water samples collected from the Alachua Conservation Trust (ACT) wetland ( A ) pH 4.3 ( B ) pH adjusted to 7.5. ( A ) (B)
117 Figure 5 2 Excitation emission fluorescence spectra of water samples collected from a riparian wetland along the Ichetucknee spring fed river (SPG) ( A ) pH 8.1 ( B ) pH adjusted to 7.5. (A) (B)
118 Figure 5 3 Excitation emission fluorescence spectra of water synthetic organic compounds ( A ) Tyrosine ( B ) Humic acid sodium salt. (B) (A)
119 Figure 5 4. Dose response curves for the toxicity to Ceriodaphnia dubia exposed to: ( A ) AgNO 3 and ( B ) nAg particles in MHW. Percent survival values are presented as the mean SD. The vertical dashed lines represent the LC 50s calculated by probit analysis. ( A ) LC50 = 0.695 ppb; Confidence Interval at 95% ranging from 0.648 to 0.739 ppb. ( B ) LC50= 0.482 ppb; and Confide nce Interval at 95% ranging from 0.455 to 0.523 ppb (A ) b (B)
120 Figure 5 5 Dose response curves for the toxicity to Ceriodaphnia dubia exposed to: ( A ) AgNO 3 and ( B ) nAg particles in SPG water. Percent survival values are presented as the mean standard deviation. The vertical dashed lines represent the LC50s calculated by probit analysis. ( A ) LC50 = 1.243 ppb; Confidence Interval at 95% ranging from 1.169 to 1.333 ppb ( B ) LC50= 0.519 ppb; and Confidence Interval at 95% ranging from 0.452 to 0.572 ppb (A) (B)
121 Figure 5 6 Dose response curves for the toxicity to Ceriodaphnia dubia exposed to: ( A ) AgNO 3 and ( B ) nAg particles in ACT water. Percent survival values are presented as the mean standard deviation. The vert ical dashed lines represent the LC50s calculated by probit analysis. ( A ) LC50 = 5.50 ppb ; Confidence Interval at 95% ranging from 4.743 to 6.502ppb ( B ) LC50= 37.859 ppb; and Confidence Interval at 95% ranging from 33.890 to 43.122 ppb (A) (B)
122 Figure 5 7 Comparison plot of obtained LC 50 Ceriodaphnia dubia grown in ACT, SPG, and MHW waters containing silver added as nanosilver (nAg) particles and AgNO 3
123 Figure 5 8. Dose response curves for the toxicity to Ceriodaphnia dubia exposed to 1 month old nAg suspensions prepared in : ( A ) SPG water and ( B ) ACT water. Percent survival values are presented as the mean standard deviation. The vertical dashed lines represent the LC 50 s calculated b y probit analysis. ( A ) LC 50 = 0.585 ; Confidence Interval at 95% ranging from 0.543 to 0.632 ppb. ( B ) LC 50 = 18 .4 ppb Confid ence Interval at 95% ranging from 15.6 to 20.7 ppb (A) (B)
124 Figure 5 9 Dose response curves for the toxicity to P seudokirchneriella subcapitata exposed to ( A ) AgNO 3 and ( B ) nAg particles in CM. Algal vitality is presented as chlorophyll a concentration mean standard deviation The vertical dashed lines reperesent calculated IC 50 values. ( A ) IC 50 = 6.08 ppb. ( B ) IC 50 = 4.61 ppb. (A) (B)
125 Figure 5 10. Dose response curves for the toxicity to P seudokirchneriella subcapitata exposed to ( A ) AgNO 3 and ( B ) nAg particles in SPG water. Algal vitality is presented as chlorophyll a concentration mean standard deviation The vertical dashed lines represent calculated IC 50 values. ( A ) IC 50 = 15.3 ppb. ( B ) IC 50 = 18.2 ppb. (A) (B)
126 Figure 5 11 Dose response curves for the toxicity to P seudokirchneriella subcapitata exposed to ( A ) AgNO 3 and ( B ) nAg particles in ACT water. Algal vitality is presented as chlorophyll a concentration mean standard deviation The vertical dashed lines represent calculated IC 50 values. ( A ) IC 50 = 30.6 ppb. ( B ) IC 50 = 158 ppb (A) (B)
127 Figure 5 12 Comparison plot of obtained LC 50 P seudokirchneriella subcapitata grown in ACT, SPG, and CM waters containing silver added as nanosilver (nAg) particles and AgNO 3
128 Figure 5 13. Dose response curves for C eriodaphnia dubia exposed to 1ppb AgNO 3 (reported as total Ag) at increasing concentrations of model organic compounds ( A ) tyrosine and ( B ) humic acid (HA). Percent survival values are presented as the mean standard deviation. (a) (b) (B) (A)
129 Figure 5 14 Dose response curves for Ceriodaphnia dubia exposed to 1ppb nAg (reported as total Ag) using ( A ) fresh nAg NP suspension and ( B ) 4 month aged nAg NP suspension with increasing concentrations of humic acid (HA). Percent survival values are presented as the mean standard deviation. (A) (b) (B)
130 Figure 5 15 Dose response curves for P seudokirchneriella subcapitata exposed to 10ppb nAg (reported as total Ag) at increasing concentrations of model organic compounds ( A ) tyrosine and ( B ) humic acid (HA). Percent survival values are presented as the mean standard deviation. Horizontal dashed lines are chl a concentrations of controls with no nAg or humic acid additions. (A) (B)
131 Figure 5 16 Dose response curve for P. subcapitata exposed to 10ppb AgNO 3 (reported as total Ag) at increasing concentra tions of model organic compound humic acid (HA). Percent survival values are presented as the mean standard deviation. Horizontal dashed lines are chl a concentrations of control with no AgNO 3 or humic acid additions.
132 Figure 5 17 Dissolution of nAg over time in nAg ACT, nAg SPG, and nAg NP suspensions expressed as averages of Ag + in three successive filtrates. Error bars are standard deviations of averages. The inset graph is the dissoluti on of nAg ACT and nAg SPG over time on a smaller scale.
133 CHAPTER 6 CONCLUSIONS AND RECOMMENDATIONS The development of and use of nanosilver (nAg) based products is rapidly growing as is the interest in determining its fate and potential impacts on organisms in aquatic systems. Most research to date has used synthetic waters to deduce the potential fate and impact of nAg in aquatic environments. This research was driven by the hypothesis that the complex chemistry and high variability in the chemical composition of natural waters, as well as the type of exposed organisms would dictate the biological imp acts of nAg resulting in toxicity patterns that vary from those predicted using lab synthetic waters. Based on a thorough review of current literature, this appears to be the first set of studies designed to take into account both nanoparticle transforma tion and interaction with biota in natural waters. This objective was accomplished by comparatively growing two selected model organisms (i.e. the cladoceran Ceriodaphnia dubia and the freshwater green algae Pseudokirchneriella subcapitata ) in synthetic (c ulture media) and natural waters Figure 6.1 summarizes the main steps of the experimental approach used in this study. Natural water samples used in the study were collected from a freshwater marsh wetland within the Alachua Conservation Trust lands on the Prairie Creek Preserve in Gainesville, FL (ACT) and from a riparian wetland on a spring fed river located in Ft. White, FL (SPG). The former is a low ionic strength and high DOC wa ter, while the latter exhibited a much higher ionic strength but low DOC concentration. These natural waters were used throughout the study and in parallel with commonly used synthetic
134 media such as Nanopure water (NP), algal culture medium (CM), an d moderately hard water (MHW) used as growth media for freshwater invertebrates. Results from transformation and stability studies demonstrated successfully that natural waters solution) in a way that do not follow patterns observed with nAg suspended in synthetic waters. When linked to biological responses in dose exposure studies, the results show that toxicity is dependent on water chemistry and organism type as well, but unlike findings generated from synthetic waters, ionic Ag do es not appear to be the only driver of toxicity in complex natural waters. Organisms exposed to nAg in the two natural waters showed drastically different responses. For both organisms, the ACT water (characterized by high DOC and low ionic strength) mitigated the toxic response better than the SPG water (characterized by low DOC and higher ionic strength). It should be noted that C. dubia was more sensitive than P. subcapitat a in all exposure studies. In addition to nAg suspensions, silver solutions were also prepared by dissolving AgNO 3 in different waters and used to help elucidate toxicity pathways/mechanisms (i.e. Ag + vs. nAg). First, AgNO 3 was more toxic than nAg to both organisms in ACT water suggesting that nAg coating by DOC affects both particle dissolution and toxicity through contact with cell membranes. However, in SPG water, nAg shows toxicity to C. dubia that is higher than the level observed when grown in AgNO 3 containing waters. The opposite trend is seen for P. subcapitata These trends point out the specific responses of different organisms to the same toxicant in the same water, suggesting that data extrapolation from one organism to another could be mislea ding. In fact, organism responses to nAg
135 are often generalized where bacterial responses to nAg are sometimes used to imply toxicity for higher trophic level organisms. Further studies should therefore include higher organisms such as aquatic vertebra surface with these organisms in natural waters should be explored further. In addition to the above findings, this research has demonstrated the significance of natural organic matter in terms of particle reactivity and organism response to nAg. Waters with high concentrations of DOC can ameliorate toxicity of nAg and AgNO 3 due to particle coating and scavenging of Ag + while wa ters with low DOC barely ameliorate the toxicity induced by AgNO 3 Titration experiments of synthetic suspensions of nAg with model organic compounds confirmed the role of DOC in mitigating nAg toxicity. Figure 6.2 brings together in a comparative way the role of Ag + produced from nAg dissolution and that of DOM in the toxicity of nAg. Research relying on the use of synthetic waters to assess the toxic implications of nAg including data obtained in this study has demonstrated the significance of nAg dissolu tion and production of ionic Ag as main pathway leading to observed adverse biological effects. In contrast, the use of natural waters as solvents show that dissolved organic carbon plays a significant role in mitigating the adverse biological effects of n Ag through a combination of nanoparticle coating (change of surface properties) and binding of soluble Ag + Accordingly, an accurate assessment of the implications of nAg and other engineered nanoparticles in natural aquatic systems requires experimental a pproaches that take into account the peculiar aspects of aquatic systems under consideration.
136 Finally, this research is the first of its kind and therefore many issues still need to be addressed. Similar experiments should be carried out for different sur face modified nAg particles. Here, bare nAg particles were used and their interactions with natural waters would likely be very different from that of coated particles (e.g. citrate and PVP coated nAg particles). Additionally, chronic toxicity assays alon g with long term dissolution experiments should be conducted in order to understand the implications of long term exposure in natural waters. With regard to the potential mechanisms of toxicity prevalent in different natural waters, it is advisable to inve stigate pathways such as the generation of reactive oxygen species in addition to nanoparticle dissolution
137 Figure 6 1 Experimental approach used for determining nanos ilver transformations and interactions with aquatic organisms in natural and synthetic waters.
138 Figure 6 2. Conceptual diagram linking the role of dissolved organic matter (DOM), synthetic waters and ionic Ag to nanosilver toxicity trends observed in these studies The presence of other types of ligands is p resented as a question since their contribution to toxicity has not been fully elucidated
139 APPENDIX A CHEMICAL COMPOSITION OF MODERATELY HARD WATER Table A 1 Chemical composition of moderately hard water (MHW) used as example synthetic high ionic strength water (culture growth media for freshwater invertebrates) Chemical Concentration in mg/L NaHCO 3 98.9 CaSO 4 *2H 2 O 60 MgSO 4 (anhydrous) 60 KCl 4
140 APPENDIX B CHEMICAL COMPOSITION OF THE ALGAL CULTURE MEDIUM Table B 1 Chemical composition of the algal culture medium used. Macron utrients Concentration in m g/L Micron utrients Concentration in g /L MgSO 4 7H 2 O 14.7 ZnCl 2 3.28 MgCl 2 6H 2 O 12.16 CoCl 2 6H 2 O 1.428 CaCl 2 2H 2 O 4.4 Na 2 MoO 4 2H 2 O 7.26 NaHCO 3 15 CuCl 2 2H 2 O 0.012 NaNO 3 25.5 H 3 BO 3 185.6 K 2 HPO 4 1.044 MnCl 2 4H 2 0 415 FeCl 3 6H 2 O 159.8
141 LIST OF REFERENCES  Masciangioli T, Zhang WX. 2003. Environmental technologies at the nanoscale. Environmental Science & Technology 37:102A 108A.  Gao J, Youn S, Hovsepyan A, Llaneza VnL, Wang Y, Bitton G, Bonzongo J CJ. 2009. Dispersion and Toxicity of Selected Manufactured Nanomaterials in Natural River Water Samples: Effects of Water Chemical Composition. Environmental Science & Technology 43:3322 3328.  2010. The Project on Emerging Nanotechnologi es. Woodrow Wilson International Center for Scholars and the Pew Charitable Trusts.  Kratschmer W. 1995. Fullerenes and fullerites: New forms of carbon. Nanostructured Materials 6:65 72.  Lamb LD, Huffman DR. 1993. Fullerene production. Journal of Ph ysics and Chemistry of Solids 54:1635 1643.  Jinno K, Sato Y, Nagashima H, Itoh K. 1998. Separation and identification of higher fullerenes by high performance liquid chromatography coupled with electrospray ionization mass spectrometry. Journal of Micr ocolumn Separations 10:79 88.  Treubig JM, Brown PR. 2002. Analysis of C 60 and C 70 fullerenes using high performance liquid chromatography Fourier transform infrared spectroscopy. Journal of Chromatography A 960:135 142.  Zarzycki PK, Ohta H, Saito Y, Jinno K. 2006. Chromatographic behavior of C60 and C70 fullerenes at subambient temperature with n alkanes mobile phases. Chromatographia 64:79 82.  Markovic Z, Trajkovic V. 2008. Biomedical potential of the reactive oxygen species generation and qu enching by fullerenes (C60). Biomaterials 29:3561 3573.  Singh CP, Roy S. 2004. Dynamics of all optical switching in C 60 and its application to optical logic gates. Optical Engineering 43:426 431.  Majumdar HS, Baral JK, Osterbacka R, Ikkala O, St ubb H. 2005. Fullerene based bistable devices and associated negative differential resistance effect. Organic Electronics 6:188 192.  Da Ros T, Spalluto G, Prato M. 2001. Biological applications of fullerene derivatives: A brief overview. Croatica Chem ica Acta 74:743 755.  Jensen AW, Wilson SR, Schuster DI. 1996. Biological applications of fullerenes. Bioorganic & Medicinal Chemistry 4:767 779.
142  Bisaglia M, Natalini B, Pellicciari R, Straface E, Malorni W, Monti D, Franceschi C, Schettini G. 2000. C 3 fullero tris methanodicarboxylic acid protects cerebellar granule cells from apoptosis. Journal of Neurochemistry 74:1197 1204.  Alargova RG, De guchi S, Tsujii K. 2001. Stable Colloidal Dispersions of Fullerenes in Polar Organic Solvents. Journal of the American Chemical Society 123:10460 10467.  Andersson T, Nilsson K, Sundahl M, Westman G, Wennerstrom O. 1992. C 60 Embedded in Gamma Cyclodex trin a Water Soluble Fullerene. Journal of the Chemical Society Chemical Communications :604 606.  Guldi DM, Huie RE, Neta P, Hungerbuhler H, Asmus KD. 1994. Excitation of C 60, Solubilized in Water by Triton X 100 and Gamma Cyclodextrin, and Subsequent Charge Separation Via Reductive Quenching. Chemical Physics Letters 223:511 516.  Kawaguchi M, Ik eda A, Shinkai S, Neda I. 2000. Electrochemical studies of calixarene fullerene inclusion processes. Journal of Inclusion Phenomena and Macrocyclic Chemistry 37:253 258.  Brettreich M, Hirsch A. 1998. A highly water soluble dendrofullerene. Tet rahedron Letters 39:2731 2734.  Wudl F. 2002. Fullerene materials. Journal of Materials Chemistry 12:1959 1963.  Yang J, Alemany LB, Driver J, Hartgerink JD, Barron AR. 2007. Fullerene derivatized amino acids: synthesis, characterization, antioxida nt properties, and solid phase peptide synthesis. Chemistry 13:2530 2545.  Deguchi S, Alargova RG, Tsujii K. 2001. Stable dispersions of fullerenes, C 60 and C 70, in water. Preparation and characterization. Langmuir 17:6013 6017.  Lyon DY, Adams L K, Falkner JC, Alvarez PJJ. 2006. Antibacterial activity of fullerene water suspensions: Effects of preparation method and particle size. Environmental Science & Technology 40:4360 4366.  Andrievsky GV, Klochkov VK, Karyakina EL, McHedlov Petrossyan NO 1999. Studies of aqueous colloidal solutions of fullerene C60 by electron microscopy. Chemical Physics Letters 300:392 396.  Lovern SB, Klaper R. 2006. Daphnia magna mortality when exposed to titanium dioxide and fullerene (C 60) nanoparticles. Envir onmental Toxicology and Chemistry 25:1132 1137.  Scott Fordsmand JJ, Krogh PH, Schaefer M, Johansen A. 2008. The toxicity testing of double walled nanotubes contaminated food to Eisenia veneta earthworms. Ecotoxicology and Environmental Safety 71:616 6 19.
143  Johansen A, Pedersen AL, Jensen KA, Karlson U, Hansen BM, Scott Fordsmand JJ, Winding A. 2008. Effects of C60 Fullerene Nanoparticles on Soil Bacteria and Protozoans Environmen tal Toxicology and Chemistry 27:1895 1903.  Khler AR, Som C, Helland A, Gottschalk F. 2008. Studying the potential release of carbon nanotubes throughout the application life cycle. Journal of Cleaner Production 16:927 937.  Ju Nam Y, Lead JR. 200 8. Manufactured nanoparticles: An overview of their chemistry, interactions and potential environmental implications. Science of The Total Environment 400:396 414  Hsu H L, Jehng J M, Sung Y, Wang L C, Yang S R. 2008. The synthesis, characterization of oxidized multi walled carbon nanotubes, and application to surface acoustic wave quartz crystal gas sensor. Materials Chemistry and Physics 109:148 155.  Robertson J. 2004. Realistic applications of CNTs. Materials Today 7:46 52.  Moisala A, Nasibulin AG, Kauppinen EI. 2004. The Role of Metal Nanoparticles in the Catalytic Production of Single Walled Carbon Nanotubes A Review. ChemInform 35.  Niyogi S, H amon MA, Hu H, Zhao B, Bhowmik P, Sen R, Itkis ME, Haddon RC. 2002. Chemistry of Single Walled Carbon Nanotubes. Accounts of Chemical Research 35:1105 1113.  Yu Y, Zhang Q, Mu Q, Zhang B, Yan B. 2008. Exploring the Immunotoxicity of Carbon Nanotubes. N anoscale Research Letters 3:271 277.  Klaine SJ, Alvarez PJJ, Batley GE, Fernandes TF, Handy RD, Lyon DY, Mahendra S, McLaughlin MJ, Lead JR. 2008. Nanomaterials in the environment: Behavior, fate, bioavailability, and effects. Environmental Toxicology and Chemistry 27:1825 1851.  Kasuga T. 2006. Formation of titanium oxide nanotubes using chemical treatments and their characteristic properties. Thin Solid Films 496:141 145.  Wu D, Liu J, Zhao XN, Li AD, Chen YF, Ming NB. 2006. Sequence of event s for the formation of titanate nanotubes, nanofibers, nanowires, and nanobelts. Chemistry of Materials 18:547 553.  Hung WC, Chen YC, Chu H, Tseng TK. 2008. Synthesis and characterization of TiO2 and Fe/TiO2 nanoparticles and their performance for pho tocatalytic degradation of 1,2 dichloroethane. Applied Surface Science 255:2205 2213.  Li Y, Lee N H, Hwang D S, Song JS, Lee EG, Kim S J. 2004. Synthesis and Characterization of Nano Titania Powder with High Photoactivity for Gas Phase
144 Photo oxidation of Benzene from TiOCl2 Aqueous Solution at Low Temperatures. Langmuir 20:10838 10844.  Barreca D, Gasparotto A, Maccato C, Tondello E, Comini E, Sberveglieri G. 2008. Innovative metal oxide nanosystems for gas sensing: From design to application. Nuov o Cimento Della Societa Italiana Di Fisica B General Physics Relativity Astronomy and Mathematical Physics and Methods 123:1369 1380.  Wang ZJ, Zhang HM, Zhang LG, Yuan JS, Yan SG, Wang CY. 2003. Low temperature synthesis of ZnO nanoparticles by solid state pyrolytic reaction. Nanotechnology 14:11 15.  Serpone N, Dondi D, Albini A. 2007. Inorganic and organic UV filters: Their role an d efficacy in sunscreens and suncare products. Inorganica Chimica Acta 360:794 802.  Gao L, Zhang Q. 2001. Effects of amorphous contents and particle size on the photocatalytic properties of TiO2 nanoparticles. Scripta Materialia 44:1195 1198.  Cio ffi N, Ditaranto N, Torsi L, Picca R, De Giglio E, Sabbatini L, Novello L, Tantillo G, Bleve Zacheo T, Zambonin P. 2005. Synthesis, analytical characterization and bioactivity of Ag and Cu nanoparticles embedded in poly vinyl methyl ketone films. Analytica l and Bioanalytical Chemistry 382:1912 1918.  Diegoli S, Manciulea AL, Begum S, Jones IP, Lead JR, Preece JA. 2008. Interaction between manufactured gold nanoparticles and naturally occurring organic macromolecules. Science of The Total Environment 402 :51 61.  Doshi R, Braida W, Christodoulatos C, Wazne M, O'Connor G. 2008. Nano aluminum: Transport through sand columns and environmental effects on plants and soil communities. Environmental Research 106:296 303.  Li Q, Mahendra S, Lyon DY, Brunet L, Liga MV, Li D, Alvarez PJJ. 2008 Antimicrobial nanomaterials for water disinfection and microbial control: Potential applications and implications. Water Research 42:4591 4602.  Quinn J, Geiger C, Clausen C, Brooks K, Coon C, O'Hara S, Krug T, Major D, Yoon W S, Gavaskar A, Holdsworth T. 2005. Field Demonstration of DNAPL Dehalogenation Using Emulsified Zero Valent Iron. Environmental Science & Technology 39:1309 1318.  Pyaten ko A, Yamaguchi M, Suzuki M. 2005. Laser photolysis of silver colloid prepared by citric acid reduction method. Journal of Physical Chemistry B 109:21608 21611.  Sharma VK, Yngard RA, Lin Y. 2009. Silver nanoparticles: Green synthesis and their antimic robial activities. Advances in Colloid and Interface Science 145:83 96.
145  Lu HW, Liu SH, Wang XL, Qian XF, Yin J, Zhu ZK. 2003. Silver nanocrystals by hyperbranched polyurethane assisted photochemical reduction of Ag+. Materials Chemistry and Physics 81 :104 107.  Esumi K, Tano T, Torigoe K, Meguro K. 1990. Preparation and characterization of bimetallic palladium copper colloids by thermal decomposition of their acetate compounds in organic solvents. Chemistry of Materials 2:564 567.  Kim D, Jeong S, Moon J. 2006. Synthesis of silver nanoparticles using the polyol process and the influence of precursor injection. Nanotechnology 17:4019 4024.  Henglein A. 2001. Reduction of Ag(CN)2 on Silver and Platinum Colloidal Nanoparticles. Langmuir 17:232 9 2333.  Xiao qin L, Elliott DW, Wei xian Z. 2006. Zero Valent Iron Nanoparticles for Abatement of Environmental Pollutants: Materials and Engineering Aspects. Critical Reviews in Solid State & Materials Science 31:111 122.  Barrena R, Casals E, Co ln J, Font X, Snchez A, Puntes V. 2009. Evaluation of the ecotoxicity of model nanoparticles. Chemosphere 75: 850 857  Yang Y, Jing L, Yu X, Yan D, Gao MY. 2007. Coating Aqueous Quantum Dots with Silica via Reverse Microemul sion Method: Toward Size Controllable and Robust Fluorescent Nanoparticles. Chem Mater 19:4123 4128.  Jamieson T, Bakhshi R, Petrova D, Pocock R, Imani M, Seifalian AM. 2007. Biological applications of quantum dots. Biomaterials 28:4717 4732.  Y. Y ang MYG. 2005. Preparation of Fluorescent SiO2 Particles with Single CdTe Nanocrystal Cores by the Reverse Microemulsion Method. Advanced Materials 17:2354 2357.  Wang X, Ruedas Rama MJ, Hall EAH. 2007. The Emerging Use of Quantum Dots in Analysis. Ana lytical Letters 40:1497 1520.  Reiss P, Protiere M, Li L. 2009. Core/Shell Semiconductor Nanocrystats. Small 5:154 168.  Hines MA, Guyot Sionnest P. 1996. Synthesis and Characterization of Strongly Luminescing ZnS Capped CdSe Nanocrystals. The Jo urnal of Physical Chemistry 100:468 471.  Daou TJ, Li L, Reiss P, Josserand Vr, Texier I. 2009. Effect of Poly(ethylene glycol) Length on the in Vivo Behavior of Coated Quantum Dots. Langmuir 25:3040 3044.
146  Smith CJ, Shaw BJ, Handy RD. 2007. Toxici ty of single walled carbon nanotubes to rainbow trout, (Oncorhynchus mykiss): Respiratory toxicity, organ pathologies, and other physiological effects. Aquatic Toxicology 82:94 109.  Ghafari P, St Denis CH, Power ME, Jin X, Tsou V, Mandal HS, Bols NC, Tang X. 2008. Impact of carbon nanotubes on the ingestion and digestion of bacteria by ciliated protozoa. Nat Nano 3:347 351.  Lyon DY, Fortner JD, Sayes CM, Colvin VL, Hughes JB. 200 5. Bacterial Cell Association and Antimicrobial Activity of a C60 Water Suspension Environmental Toxicology and Chemistry 24:2757 2762.  Spesia MB, Milanesio ME, Durantini EN. 2008. Synthesis, properties and photodynamic inactivation of Escherichia coli by novel cationic fullerene C60 derivatives. European Journal of Medicinal Chemistry 43:853 861.  Fang JS, Lyon DY, Wiesner MR, Dong JP, Alvarez PJJ. 2007. Effect of a f ullerene water suspension on bacterial phospholipids and membrane phase behavior. Environmental Science & Technology 41:2636 2642.  Oberdorster E. 2004. Manufactured nanomaterials (Fullerenes, C 60) induce oxidative stress in the brain of juvenile largemouth bass. Environmental Health Perspectives 112:1058 1062.  Lyon DY, Alvarez PJJ. 2008. Fullerene Water Suspension (nC(60)) Exert s Antibacterial Effects via ROS Independent Protein Oxidation. Environmental Science & Technology 42:8127 8132.  Lyon DY, Brunet L, Hinkal GW, Wiesner MR, Alvarez PJJ. 2008. Antibacterial Activity of Fullerene Water Suspensions (nC60) Is Not Due to ROS Mediated Damage. Nano Letters 8:1539 1543.  Zhu X, Zhu L, Li Y, Duan Z, Chen W, Alvarez PJJ. 2007. Developmental Toxicity in Zebrafish (Danio Rerio) Embryos After Exposure to Manufactured Nanomaterials:Buckminsterfullerence Aggregates (nC60) and Fullerol Environmental Toxicology and Chemistry 26:976 979.  Fortner JD, Lyon DY, Sayes CM, Boyd AM, Falkner JC, Hotze EM, Alemany LB, Tao YJ, Guo W, Ausman KD, Colvin VL, Hughes JB. 2005. C60 in Water: Nanocrystal Formation and Microbial Response. Environ Sci Technol 39:4307 4316.  Han ZT, Zhang FW, Lin DH, Xing BS. 2008. Clay minerals affect the sta bility of surfactant facilitated carbon nanotube suspensions. Environmental Science & Technology 42:6869 6875.  Ma X, Bouchard D. 2009. Formation of Aqueous Suspensions of Fullerenes. Environmental Science & Technology 43:330 336.
147  Wang H, Zhou W, Ho DL, Winey KI, Fischer JE, Glinka CJ, Hobbie EK. 2004. Dispersing single walled carbon nanotubes with surfactants: A small angle neutron scattering study. Nano Letters 4:1789 1793.  Dong L, Joseph KL, Witkowski CM, Craig MM. 2008. Cytotoxicity of sin gle walled carbon nanotubes suspended in various surfactants. Nanotechnology 19.  Henry TF MM, James T. Fleming, John Wilgus, Robert N. Compton, and Gary S. Sayler 2007. Attributing Effects of Aqueous C60 Nano Aggregates to Tetrahydrofuran Decomposition Products in Larval Zebrafish by Assessment of Gene Expression Environ Health Perspect 115:1059 1065.  Spohn P, Hirsch C, Hasler F, Bruinink A, Krug HF, Wick P. 2009. C60 fullerene: A powerful antioxidant or a damaging agent? The importa nce of an in depth material characterization prior to toxicity assays. Environmental Pollution In Press, Corrected Proof.  Hotze EM, Labille J, Alvarez P, Wiesner MR. 2008. Mechanisms of photochemistry and reactive oxygen production by fullerene suspen sions in water. Environmental Science & Technology 42:4175 4180.  Isakovic A, Markovic Z, Todorovic Markovic B, Nikolic N, Vranjes Djuric S, Mirkovic M, Dramicanin M, Harhaji L, Raicevic N, Nikolic Z, Trajkovic V. 2006. Distinct cytotoxic mechanisms of pristine versus hydroxylated fullerene. Toxicological Sciences 91:173 183.  Kang S, Pinault M, Pfefferle LD, Elimelech M. 2007. Single walled carbon nanotubes exhibit strong antimicrobial activity. Langmuir 23:8670 8673.  Kennedy AJ, Hull MS, Stee vens JA, Dontsova KM, Chappell MA, Gunter JC, Weiss CA. 2008. Factors influencing the partitioning and toxicity of nanotubes in the aquatic environment. Environmental Toxicology and Chemistry 27:1932 1941.  Heinlaan M, Ivask A, Blinova I, Dubourguier H C, Kahru A. 2008. Toxicity of nanosized and bulk ZnO, CuO and TiO2 to bacteria Vibrio fischeri and crustaceans Daphnia magna and Thamnocephalus platyurus. Chemosphere 71:1308 1316.  Hu X, Cook S, Wang P, Hwang H m. 2009 In vitro evaluation of cytotox icity of engineered metal oxide nanoparticles. Science of The Total Environment 407:3070 3072.  Federici G, Shaw BJ, Handy RD. 2007. Toxicity of titanium dioxide nanoparticles to rainbow trout (Oncorhynchus mykiss): Gill injury, oxidative stress, and o ther physiological effects. Aquatic Toxicology 84:415 430.
148  Adams LK, Lyon DY, Alvarez PJJ. 2006. Comparative eco toxicity of nanoscale TiO2, SiO2, and ZnO water suspensions. Water Research 40:3527 3532.  Zhu XS, Zhu L, Duan ZH, Qi RQ, Li Y, Lang Y P. 2008. Comparative toxicity of several metal oxide nanoparticle aqueous suspensions to Zebrafish (Danio rerio) early developmental stage. Journal of Environmental Science and Health Part a Toxic/Hazardous Substances & Environmental Engineering 43:278 284  Pradhan A, Seena S, Pascoal Cu, C a ssio F. 2011. Can Metal Nanoparticles Be a Threat to Microbial Decomposers of Plant Litter in Streams? Microbial Ecology 62:58 68.  Wang YG, Aker WG, Hwang HM, Yedjou CG, Yu HT, Tchounwou PB. 2011. A study of the mechanism of in vitro cytotoxicity of metal oxide nanoparticles using catfish primary hepatocytes and human HepG2 cells. Science of the Total Environment 409:4753 4762.  Z hang L, Jiang Y, Ding Y, Povey M, York D. 2007 Investigation into the antibacterial behaviour of suspensions of ZnO nanoparticles (ZnO nanofluids). Journal of Nanoparticle Research 9:479 489.  Choi O, Deng KK, Kim N J, Ross Jr L, Surampalli RY, Hu Z. 2008. The inhibitory effects of silver nanoparticles, silver ions, and silver chloride colloids on microbial growth. Water Research 42:3066 3074.  Lok CN, Ho CM, Chen R, He QY, Yu WY, Sun HZ, Tam PKH, Chiu JF, Che CM. 2006. Proteomic analysis of the m ode of antibacterial action of silver nanoparticles. Journal of Proteome Research 5:916 924.  Morones JR, Elechiguerra JL, Camacho A, Holt K, Kouri JB, Ramirez JT, Yacaman MJ. 2005. The bactericidal effect of silver nanoparticles. Nanotechnology 16:234 6 2353.  Sondi I, Salopek Sondi B. 2004. Silver nanoparticles as antimicrobial agent: a case study on E coli as a model for Gram negative bacteria. Journal of Colloid and Interface Science 275:177 182.  Griffitt RJ, Luo J, Gao J, Bonzongo JC, Barber DS. 2008. Effects of particle composition and species on toxicity of metallic nanomaterials in aquatic organisms. Environmental Toxicology and Chemistry 27:1972 1978.  Navarro E, Baun A, Behra R, Hartmann NB, Filser J, Miao AJ, Quigg A, Santschi PH, Sigg L. 2008. Environmental behavior and ecotoxicity of engineered nanoparticles to algae, plants, and fungi. Ecotoxicology 17:372 386.  Navarro E, Piccapietra F, Wagner B, Marconi F, Kaegi R, Odzak N, Sigg L, Behra R. 2008. Toxicity of Silver Nanoparticles to Chlamydomonas reinhardtii. Environ Sci Technol
149  Roh JY, Sim SJ, Yi J, Park K, Chung KH, Ryu DY, Choi J. 2009. Ecotoxicity of Silver Nanoparticles on the Soil Nematode Caenorhabditis elegans Using F unctional Ecotoxicogenomics. Environmental Science & Technology 43:3933 3940.  Lubick N. 2008. Nanosilver toxicity: ions, nanoparticies or both? Environmental Science & Technology 42:8617 8617.  Goodman CM, McCusker CD, Yilmaz T, Rotello VM. 2004 Toxicity of gold nanoparticles functionalized with cationic and anionic side chains. Bioconjugate Chemistry 15:897 900.  Griffitt RJ, Weil R, Hyndman KA, Denslow ND, Powers K, Taylor D, Barber DS. 2007. Exposure to Copper Nanoparticles Causes Gill I njury and Acute Lethality in Zebrafish (Danio rerio). Environmental Science & Technology 41:8178 8186.  Lopes I, Ribeiro R, Antunes FE, Rocha Santos TAP, Rasteiro MG, Soares A, Goncalves F, Pereira R. 2012. Toxicity and genotoxicity of organic and ino rganic nanoparticles to the bacteria Vibrio fischeri and Salmonella typhimurium. Ecotoxicology 21:637 648.  Gagne F, Auclair J, Turcotte P, Fournier M, Gagnon C, Sauve S, Blaise C. 2008. Ecotoxicity of CdTe quantum dots to freshwater mussels: Impacts on immune system, oxidative stress and genotoxicity. Aquatic Toxicology 86:333 340.  Wang J, Zhang X, Chen Y, Sommerfeld M, Hu Q. 2008. Toxicity assessment of manufactured nanomaterials using the unicellular green alga Chlamydomonas reinhardtii. Chemo sphere 73:1121 1128  Zhang W, Sun X, Chen L, Lin KF, Dong QX, Huang CJ, Fu RB, Zhu J. 2012. Toxicological effect of joint cadmium selenium quantum dots and copper ion exposure on zebrafish. Environmental Toxicology and Chemistry 31:2117 2123.  Mahendra S, Zhu HG, Colvin VL, Alvarez PJ. 2008. Quantum Dot Weathering Results in Microbial Toxicity. Environmental Science & Technology 42:9424 9430.  Schneider R, Wolpert C, Guilloteau H, Balan L, Lambert J, Merlin C. 2009. The exposure of bacteria to CdTe core quantum dots: the importance of surface chemistry on cytotoxicity. Nanotechnology 20.  Xiu Z m, Zhang Q b, Puppala HL, Colvin VL, Alvarez PJJ. 2012. Negligible Particle Specific Antibacterial Activity of Sil ver Nanoparticles. Nano Letters  Kennedy AJ, Hull MS, Bednar AJ, Goss JD, Gunter JC, Bouldin JL, Vikesland PJ, Steevens JA. 2010. Fractionating Nanosilver: Importance for Determining Toxicity to Aquatic Test Organisms. Environmental Science & Technol ogy 44:9571 9577.
150  Wang Z, Chen J, Li X, Shao J, Peijnenburg WJGM. 2012. Aquatic toxicity of nanosilver colloids to different trophic organisms: Contributions of particles and free silver ion. Environmental Toxicology and Chemistry 31:2408 2413.  Choi O, Hu Z. 2008. Size Dependent and Reactive Oxygen Species Related Nanosilver Toxicity to Nitrifying Bacteria. Environmental Science & Technology 42:4583 4588.  Allen HJ, Impellitteri CA, Macke DA, Heckman JL, Poynton HC, Lazorchak JM, Govindaswa my S, Roose DL, Nadagouda MN. 2010. Effects from filtration, capping agents, and presence/absence of food on the toxicity of silver nanoparticles to Daphnia magna. Environmental Toxicology and Chemistry 29:2742 2750.  Gao J, Powers K, Wang Y, Zhou H, Roberts SM, Moudgil BM, Koopman B, Barber DS. 2012. Influence of Suwannee River humic acid on particle properties and toxicity of silver nanoparticles. Chemosphere 89:96 101.  Hwang ET, Lee JH, Chae YJ, Kim YS, Kim BC, Sang B I, Gu MB. 2008. Analysis of the Toxic Mode of Action of Silver Nanoparticles Using Stress Specific Bioluminescent Bacteria. Small 4:746 750.  Griffitt RJ, Hyndman K, Denslow ND, Barber DS. 2009. Comparison of Molecular and Histological Changes in Zebrafish Gills Exposed to Me tallic Nanoparticles. Toxicol Sci 107:404 415.  USEPA 2010. State of the Science Literature Review: Everything Nanosilver and More.  Nowack B, Krug HF, Height M. 2011. 120 Years of Nanosilver History: Implications for Policy Makers. Enviro nmental Science & Technology 45:1177 1183.  Lea M. 1889. On allotropic forms of silver. American Journal of Science 37:476 491.  Gottschalk F, Sonderer T, Scholz RW, Nowack B. 2010. Possibilities and limitations of modeling environmental exposure to engineered nanomaterials by probabilistic material flow analysis. Environmental Toxicology and Chemistry 29:1036 1048.  Mueller NC, Nowack B. 2008. Exposure Modeling of Engineered Nanoparticles in the Environment. Environ Sci Technol 42:4447 4453.  Hendren CO, Mesnard X, Drge J, Wiesner MR. 2011. Estimating Production Data for Five Engineered Nanomaterials As a Basis for Exposure Assessment. Environmental Science & Technology 45:2562 2569.
151  Blaser SA, Scheringer M, MacLeod M, Hungerbuhl er K. 2008. Estimation of cumulative aquatic exposure and risk due to silver: Contribution of nano functionalized plastics and textiles. Science of the Total Environment 390:396 409.  Manikam VR, Cheong KY, Razak KA. 2011. Chemical reduction methods for synthesizing Ag and Al nanoparticles and their respective nanoalloys. Materials Science and Engineering: B 176:187 203.  Bauer C, Stellacci F, Perry J. 2008. Relationship Between Structure and Solubility of Thiol Protect ed Silver Nanoparticles and Assemblies. Topics in Catalysis 47:32 41.  Sun Y, Mayers B, Xia Y. 2003. Transformation of Silver Nanospheres into Nanobelts and Triangular Nanoplates through a Thermal Process. Nano Letters 3:675 679.  Hotze EM, Phenr at T, Lowry GV. 2010. Nanoparticle Aggregation: Challenges to Understanding Transport and Reactivity in the Environment J Environ Qual 39:1909 1924.  Levard C, Hotze EM, Lowry GV, Brown GE. 2012. Environmental Transformations of Silver N anoparticles: Impact on Stability and Toxicity. Environmental Science & Technology 46:6900 6914.  Wiley B, Sun Y, Mayers B, Xia Y. 2005. Shape Controlled Synthesis of Metal Nanostructures: The Case of Silver. Chemistry A European Journal 11:454 463.  Pillai ZS, Kamat PV. 2003. What Factors Control the Size and Shape of Silver Nanoparticles in the Citrate Ion Reduction Method? The Journal of Physical Chemistry B 108:945 951.  Levard C, Hotze EM, Lowry GV, Brown GE. 2012. Environmental Transf ormations of Silver Nanoparticles: Impact on Stability and Toxicity. Environmental Science & Technology 46:6900 6914.  Tolaymat TM, El Badawy AM, Genaidy A, Scheckel KG, Luxton TP, Suidan M. 2010. An evidence based environmental perspective of manufac tured silver nanoparticle in syntheses and applications: A systematic review and critical appraisal of peer reviewed scientific papers. Science of The Total Environment 408:999 1006.  Fabrega J, Luoma SN, Tyler CR, Galloway TS, Lead JR. 2011. Silver n anoparticles: Behaviour and effects in the aquatic environment. Environment International 37:517 531.
152  Xiaohui S, Zhiya S, Yang L. 2012. Effects of silver nanoparticles on microbial community structure in activated sludge. Science of The Total Environ ment 443:828 835.  Radniecki TS, Stankus DP, Neigh A, Nason JA, Semprini L. 2011. Influence of liberated silver from silver nanoparticles on nitrification inhibition of Nitrosomonas europaea. Chemosphere 85:43 49.  Yuan ZH, Li JW, Cui L, Xu B, Zh ang HW, Yu CP. 2013. Interaction of silver nanoparticles with pure nitrifying bacteria. Chemosphere 90:1404 1411.  Lok C N, Ho C M, Chen R, He Q Y, Yu W Y, Sun H, Tam P, Chiu J F, Che C M. 2007. Silver nanoparticles: partial oxidation and antibacterial activities. Journal of Biological Inorganic Chemistry 12:527 534.  Choi O, Yu C P, Esteban Fern¡ndez G, Hu Z. 20 10. Interactions of nanosilver with Escherichia coli cells in planktonic and biofilm cultures. Water Research 44:6095 6103.  Das P, Xenopoulos MA, Williams CJ, Hoque ME, Metcalfe CD. 2012. Effects of silver nanoparticles on bacterial activity in natur al waters. Environmental Toxicology and Chemistry 31:122 130.  Xiu ZM, Ma J, Alvarez PJJ. 2011. Differential Effect of Common Ligands and Molecular Oxygen on Antimicrobial Activity of Silver Nanoparticles versus Silver Ions. Environmental Science & Te chnology 45:9003 9008.  Choi O, Cleuenger TE, Deng BL, Surampalli RY, Ross L, Hu ZQ. 2009. Role of sulfide and ligand strength in controlling nanosilver toxicity. Water Research 43:1879 1886.  Pokhrel LR, Dubey B. 2012. Potential Impact of Low Co ncentration Silver Nymphs and Daphnia magna as a Prey. Environmental Science & Technology 46:7755 7762.  Yang X, Gondikas AP, Marinakos SM, Auffan M, Liu J, Hsu Kim H, Meyer JN. 2012. Mechanism of Silver Nanoparticle Toxicity Is Dependent on Dissolved Silver and Surface Coating in Caenorhabditis elegans. Environmental Science & Technology 46:1119 1127.  f Silver Nanoparticle Contamination on the Genetic Diversity of Natural Bacterial Assemblages in Estuarine Sediments. Environmental Science & Technology 43:4530 4536.
153  Lee D Y, Fortin C, Campbell PGC. 2005. Contrasting effects of chloride on the toxic ity of silver to two green algae, Pseudokirchneriella subcapitata and Chlamydomonas reinhardtii. Aquatic Toxicology 75:127 135.  Ratte HT. 1999. Bioaccumulation and toxicity of silver compounds: A review. Environmental Toxicology and Chemistry 18:89 1 08.  He D, Dorantes Aranda JJ, Waite TD. 2012. Silver Nanoparticle Algae Interactions: Oxidative Dissolution, Reactive Oxygen Species Generation and Synergistic Toxic Effects. Environmental Science & Technology 46:8731 8738.  Miao A J, Schwehr KA, Xu C, Zhang S J, Luo Z, Quigg A, Santschi PH. 2009. The algal toxicity of silver engineered nanoparticles and detoxification by exopolymeric substances. Environmental Pollution 157:3034 3041.  Fortin C, Campbell PGC. 2001. Thiosulfate Enhances Si lver Uptake by a Green Environmental Science & Technology 35:2214 2218.  Fortin C, Campbell PGC. 2000. Silver uptake by the green alga Chlamydomonas reinhardtii in relation to chemical speciation: I nfluence of chloride. Environmental Toxicology and Chemistry 19:2769 2778.  Hiriart Baer VrP, Fortin C, Lee D Y, Campbell PGC. 2006. Toxicity of silver to two freshwater algae, Chlamydomonas reinhardtii and Pseudokirchneriella subcapitata, grown under continuous culture conditions: Influence of thiosulphate. Aquatic Toxicology 78:136 148.  Szivk I, Behra R, Sigg L. 2009. Metal induced Reactive Oxygen Species Production in Chlamydomonas Reinhardtii (Chlorophyceae) Journal of Phycology 45:427 435.  Gulati R, Lammens E, Pauw N, Donk E, Haberman J, Laugaste R, N a ges T. 2007. The role of cladocerans reflecting the trophic status of two large and sha llow Estonian lakes. Shallow Lakes in a Changing World Springer Netherlands, pp 157 166.  Hoheisel SM, Diamond S, Mount D. 2012. Comparison of nanosilver and ionic silver toxicity in Daphnia magna and Pimephales promelas. Environmental Toxicology and Chemistry 31:2557 2563.  Kennedy AJ, Chappell MA, Bednar AJ, Ryan AC, Laird JG, Stanley JK, Steevens JA. 2012. Impact of Organic Carbon on the Stability and Toxicity of Fresh and Stored Silver Nanoparticles. Environmental Science & Technology 46:1077 2 10780.  Blinova I, Niskanen J, Kajankari P, Kanarbik L, K¤kinen A, Tenhu H, Penttinen O P, Kahru A. 2012. Toxicity of two types of silver nanoparticles to aquatic
154 crustaceans Daphnia magna and Thamnocephalus platyurus. Environmental Science and Pol lution Research :1 8.  Asghari S, Johari SA, Lee JH, Kim YS, Jeon YB, Choi HJ, Moon MC, Yu IJ. 2012. Toxicity of various silver nanoparticles compared to silver ions in Daphnia magna. Journal of Nanobiotechnology 10.  Zhao C M, Wang W X. 2011. Com parison of acute and chronic toxicity of silver nanoparticles and silver nitrate to Daphnia magna. Environmental Toxicology and Chemistry 30:885 892.  McLaughlin J, Bonzongo J CJ. 2012. Effects of natural water chemistry on nanosilver behavior and tox icity to Ceriodaphnia dubia and Pseudokirchneriella subcapitata. Environmental Toxicology and Chemistry 1:168 175  Zhao C M, Wang W X. 2010. Biokinetic Uptake and Efflux of Silver Nanoparticles in Daphnia magna. Environmental Science & Technology 44: 7699 7704.  Bianchini A, Wood CM. 2003. Mechanism of acute silver toxicity in Daphnia magna. Environmental Toxicology and Chemistry 22:1361 1367.  Li X, Lenhart JJ, Walker HW. 2010. Dissolution Accompanied Aggregation Kinetics of Silver Nanoparti cles. Langmuir 26:16690 16698.  Henglein A. 1998. Colloidal Silver Nanoparticles: Photochemical Preparation and Interaction with O2, CCl4, and Some Metal Ions. Chemistry of Materials 10:444 450.  Cai W, Zhong H, Zhang L. 1998. Optical measurement s of oxidation behavior of silver nanometer particle within pores of silica host. Journal of Applied Physics 83:1705 1710.  Ivanova OS, Zamborini FP. 2009. Size Dependent Electrochemical Oxidation of Silver Nanoparticles. Journal of the American Chemi cal Society 132:70 72.  Henglein A. 1993. Physicochemical properties of small metal particles in solution: "microelectrode" reactions, chemisorption, composite metal particles, and the atom to metal transition. The Journal of Physical Chemistry 97:545 7 5471.  Liu J, Hurt RH. 2010. Ion Release Kinetics and Particle Persistence in Aqueous Nano Silver Colloids. Environmental Science & Technology 44:2169 2175.  Zhang W, Yao Y, Sullivan N, Chen Y. 2011. Modeling the Primary Size Effects of Citrate Coated Silver Nanoparticles on Their Ion Release Kinetics. Environmental Science & Technology 45:4422 4428.
155  Zook JM, Halter MD, Cleveland D, Long SE. 2012. Disenta ngling the effects of polymer coatings on silver nanoparticle agglomeration, dissolution, and toxicity to determine mechanisms of nanotoxicity. Journal of Nanoparticle Research 14.  Hadioui M, Leclerc S, Wilkinson KJ. 2013. Multimethod quantification of Ag+ release from nanosilver. Talanta 105:15 19.  Zook JM, Long SE, Cleveland D, Geronimo CLA, MacCuspie RI. 2011. Measuring silver nanoparticle dissolution in complex biological and environmental matrices using UV visible absorbance. Analytical and Bioanalytical Chemistry 401:1993 2002.  He W, Zhou Y T, Wamer WG, Boudreau MD, Yin J J. 2012. Mechanisms of the pH dependent generation of hydroxyl radicals and oxygen induced by Ag nanoparticles. Biomaterials 33:7547 7555.  Sotiriou GA, Meyer A Knijnenburg JTN, Panke S, Pratsinis SE. 2012. Quantifying the Origin of Released Ag+ Ions from Nanosilver. Langmuir 28:15929 15936.  Ho C M, Yau SK W, Lok C N, So M H, Che C M. 2010. Oxidative Dissolution of Silver Nanoparticles by Biologically Rele vant Oxidants: A Kinetic and Mechanistic Study. Chemistry An Asian Journal 5:285 293.  Ma R, Levard Cm, Marinakos SM, Cheng Y, Liu J, Michel FM, Brown GE, Lowry GV. 2012. Size Controlled Dissolution of Organic Coated Silver Nanoparticles. Environmen tal Science & Technology 46:752 759.  Bury NR, Galvez F, Wood CM. 1999. Effects of chloride, calcium, and dissolved organic carbon on silver toxicity: Comparison between rainbow trout and fathead minnows. Environmental Toxicology and Chemistry 18:56 6 2.  Gebauer JS, Treuel L. 2011. Influence of individual ionic components on the agglomeration kinetics of silver nanoparticles. Journal of Colloid and Interface Science 354:546 554.  El Badawy AM, Luxton TP, Silva RG, Scheckel KG, Suidan MT, Tola ymat TM. 2010. Impact of Environmental Conditions (pH, Ionic Strength, and Electrolyte Type) on the Surface Charge and Aggregation of Silver Nanoparticles Suspensions. Environmental Science & Technology 44:1260 1266.  Cumberland SA, Lead JR. 2009. Par ticle size distributions of silver nanoparticles at environmentally relevant conditions. Journal of Chromatography A 1216:9099 9105.  Huynh KA, Chen KL. 2011. Aggregation Kinetics of Citrate and Polyvinylpyrrolidone Coated Silver Nanoparticles in Monovalent and Divalent Electrolyte Solutions. Environmental Science & Technology 45:5564 5571.
156  Chen S F, Zhang H. 2012. Aggregation kine tics of nanosilver in different water conditions. Advances in Natural Sciences: Nanoscience and Nanotechnology 3:035006.  Chen KL, Elimelech M. 2007. Influence of humic acid on the aggregation kinetics of fullerene (C60) nanoparticles in monovalent an d divalent electrolyte solutions. Journal of Colloid and Interface Science 309:126 134.  French RA, Jacobson AR, Kim B, Isley SL, Leepenn R, Baveye PC. 2008. Influence of ionic strength, pH, and cation valence on aggregation kinetics of TiO2 nanoparti cles. Geochimica Et Cosmochimica Acta 72:A283 A283.  Liu B, Xie W, Wang D, Huang W, Yu M, Yao A. 2008. Preparation and characterization of magnetic luminescent nanocomposite particles. Materials Letters 62:3014 3017.  Li X, Lenhart JJ. 2012. Aggr egation and Dissolution of Silver Nanoparticles in Natural Surface Water. Environmental Science & Technology 46:5378 5386.  Akaighe N, MacCuspie RI, Navarro DA, Aga DS, Banerjee S, Sohn M, Sharma VK. 2011. Humic Acid Induced Silver Nanoparticle Format ion Under Environmentally Relevant Conditions. Environmental Science & Technology 45:3895 3901.  Akaighe N, Depner SW, Banerjee S, Sharma VK, Sohn M. 2012. The effects of monovalent and divalent cations on the stability of silver nanoparticles formed from direct reduction of silver ions by Suwannee River humic acid/natural organic matter. Science of The Total Environment 441:277 289.  Yin Y, Liu J, Jiang G. 2012. Sunlight Induced Reduction of Ionic Ag and Au to Metallic Nanoparticles by Dissolved Organic Matter. ACS Nano 6:7910 7919.  El Badawy AM, Silva RG, Morris B, Scheckel KG, Suidan MT, Tolaymat TM. 2011. Surface Charge Depend ent Toxicity of Silver Nanoparticles. Environmental Science & Technology 45:283 287.  Gondikas AP, Morris A, Reinsch BC, Marinakos SM, Lowry GV, Hsu Kim H. 2012. Cysteine Induced Modifications of Zero valent Silver Nanomaterials: Implications for Part icle Surface Chemistry, Aggregation, Dissolution, and Silver Speciation. Environmental Science & Technology 46:7037 7045.  Chinnapongse SL, MacCuspie RI, Hackley VA. 2011. Persistence of singly dispersed silver nanoparticles in natural freshwaters, sy nthetic seawater, and simulated estuarine waters. Science of The Total Environment 409:2443 2450.  Zhang H, Smith JA, Oyanedel Craver V. 2012. The effect of natural water conditions on the anti bacterial performance and stability of silver nanoparticl es capped with different polymers. Water Research 46:691 699.
157  Cory RM, McKnight DM. 2005. Fluorescence Spectroscopy Reveals Ubiquitous Presence of Oxidized and Reduced Quinones in Dissolved Organic Matter. Environ Sci Technol 39:8142 8149.  Weis haar JL, Aiken GR, Bergamaschi BA, Fram MS, Fujii R, Mopper K. 2003. Evaluation of Specific Ultraviolet Absorbance as an Indicator of the Chemical Composition and Reactivity of Dissolved Organic Carbon. Environ Sci Technol 37:4702 4708.  Apell JN, Boy er TH. 2010. Combined ion exchange treatment for removal of dissolved organic matter and hardness. Water Research 44:2419 2430.  Chen W, Westerhoff P, Leenheer JA, Booksh K. 2003. Fluorescence Excitation Emission Matrix Regional Integration to Quantif y Spectra for Dissolved Organic Matter. Environmental Science & Technology 37:5701 5710.  Agency USEP. 1992. Acid digestion of waters for total recoverable of dissolved metals for analysis by FLAA or ICP spectroscopy.  Hu H, Yu A, Kim E, Zhao B, Itkis ME, Bekyarova E, Haddon RC. 2005. Influence of the Zeta Potential on the Dispersability and Purification of Single Walled Carbon Nanotubes. The Journal of Physical Chemistry B 109:11520 11524.  Krutyakov YA, Kudrinskiy AA, Olenin AY, Lisichkin G V. 2008. Synthesis and properties of silver nanoparticles: Achievements and prospects. Uspekhi Khimii 77:242 269.  Erickson RJ, Brooke LT, Kahl MD, Venter FV, Harting SL, Markee TP, Spehar RL. 1998. Effects of laboratory test conditions on the toxicit y of silver to aquatic organisms. Environmental Toxicology and Chemistry 17:572 578.  USEP A 1972. Clean Water Act. In USEPA, ed.  USEPA. 2002. Short term Methods for Estimating the Chronic Toxicity of Effluents and Receving Waters to Fres hwater Organisms. In (4303T) OoW, ed, 4th Edition ed, p. 350.  USEP A 2006. Probit Analyis Program. 1.5 ed, Cincinnati, OH, USA.  Simon D, Helliwell S. 1998. Extraction and quantification of chlorophyll a from freshwater green algae. Water Research 32:2220 2223.  Gheorghiu C, Smith DS, Al Reasi HA, McGeer JC, Wilkie MP. 2010. Influence of natural organic matter (NOM) quality on Cu gill binding in the rainbow trout (Oncorhynchus mykiss). Aquatic Toxicology 97:343 352.  Richards JG, Curtis PJ, Burnison BK, Playle RC. 2001. Effects of natural organic matter source on reducing metal toxicity to rainbow trout (Oncorhynchus mykiss)
158 and on metal binding to their gills. Environmental Toxicology and Chemistry 20:1159 1166.  Glover CN, Wood CM. 2005. Accumulation and elimination of silver in Daphnia magna and the effect of natural organic matter. Aquatic Toxicology 73:406 417.  Karen DJ, Ownby DR, Forsythe BL, Bills TP, La Point TW, Cobb GB, Klaine SJ. 1999. Influence of water quality on silver toxicity to rainbow trout (Oncorhynchus mykiss), fathead minnows (Pimephales promelas), and water fleas (Daphnia magna). Environmental To xicology and Chemistry 18:63 70.  Kahru A, Dubourguier HC, Blinova I, Ivask A, Kasemets K. 2008. Biotests and biosensors for ecotoxicology of metal oxide nanoparticles: A minireview. Sensors 8:5153 5170.
159 BIOGRAPHICAL SKETCH Julianne McLaughlin was born and raised in the foothills of South Carolina. She attended College of Charleston in Charleston SC and received a Bachelor of Science in chemistry. Her love for science and the lab took her to University of Florida where she began her doctorate degree in the chemistry department with the intent to concentrate on organic chemistry. Her long held dedication to environmental protection and awareness, however, inspired Julianne to transition from the chemistry department to the department of Environmental Engineering Sciences (EES). Julianne began her Ph.D. in EES under the advisement of Dr. Jean Claude Bonzongo in August of 2007. Initially her research focus dealt with the remediation of mercury contaminated soils using alumin um wastewater treatment residuals. After a year or so working with this project Julianne made another transition. Her current Ph.D. research explores the environmental implications (e.g., toxicity to bacteria, algae and aquatic invertebrates) of nanosilve r on aquatic ecosystems using waters from a diversity of wetland ecosystems. Throughout this research, she has become well versed in biogeochemistry, nanotechnology, and aquatic ecosystem health.