Environmental Behavior of Silver Nanoparticles: Emissions from Consumer Products and Toxicty in Waste Treatment

A dissertation submitted to the Division of Research and Advanced Studies of the University of Cincinnati in partial fulfillment of the requirements for the degree of

Doctor of Philosophy

Department of Biomedical, Chemical, and Environmental Engineering of the College of Engineering and Applied Science

July 2016

By Alireza Gitipour B.Sc. Chemical Engineering Southern Tehran University, Iran

Thesis Committee:

Stephen W. Thiel, Ph.D., Chair Thabet Tolaymat, Ph.D. Vadim Guliants, Ph.D. George Sorial, Ph.D. Vesselin Shanov, Ph.D.

ABSTRACT

Nanotechnology has undergone a dramatic increase in popularity in the last decade due to the unique physicochemical characteristics of engineered nanomaterials (ENMs). Currently, approximately a quarter of all nano-enabled consumer products contain silver nanoparticles

(AgNPs). AgNPs are incorporated into a wide range of consumer products (e. g., textiles, filters, disinfectants, and washing machines) and have a wide range of medical, industrial and scientific applications.

The increased application of AgNPs will inevitably lead to their release into environmental systems. Since the presumed mechanisms governing the fate, transport and toxicity of matter at the bulk scale may not directly apply to nanomaterials, the potential environmental impacts associated with the release of AgNPs must be evaluated. Furthermore, AgNPs are manufactured with a wide range of physicochemical properties that impact their fate, transport and toxicity in the environment.

To this end, the impact of silver nanoparticles on the composting of municipal solid waste was evaluated. Neither the presence of AgNPs nor the presence of Ag+ had a statistically significant influence on leachate, gas and solid quality parameters, and therefore, on overall composting performance. However, AgNPs and Ag+ both changed the overall structure of the bacterial communities within the compost. Nevertheless, the functional performance of the composting process was not significantly affected due to the abundance and functional redundancy of the bacterial communities within the compost samples. While surface transformations of AgNPs to

AgCl and Ag2S reduce toxicity, complexation with organic matter may also play a role. The results

i of this study further suggest that at relatively low concentrations of AgNPs, these organically rich waste management systems can withstand the presence of AgNPs.

The microbial toxicity of silver nanoparticles stabilized with different capping agents were evaluated under anaerobic conditions. The AgNPs investigated were similar in size and shape but varied in surface charge. At lower AgNPs concentrations, the anaerobic decomposition process was not affected although the diversity of the microbial community was impacted. Interestingly, at higher concentrations only the cationic AgNPs demonstrated toxicity, while, the neutral and negatively charged AgNPs did not exhibit toxicity. These findings indicate that there are multiple mechanisms for nanoparticles toxicity.

In addition to the disposal studies, a study of a commonly-utilized nanosilver solution using a simulated dental unit water delivery system assessed the fate, mode of interaction and physicochemical transformations of AgNPs under a realistic usage scenario. The disinfection process led to the disappearance of the capping agents and consequently transformations of

AgNPs. In addition to further understanding the transformations that occur in the process, adsorption of the AgNPs onto the biofilms surface was demonstrated which may assist in further understanding the toxicity mechanisms of AgNPs to biofilms.

Finally, a comprehensive review was conducted to identify the key issues and knowledge gaps concerning the environmental impact of AgNPs in consumer products. This review summarizes the existing data related to characterization techniques, routes of environmental exposure and potential ecological risks of AgNPs and provides potential directions for future research.

iii

ACKNOWLEDGEMENTS

I would like express my deepest gratitude and undying love to my mother (Parvaneh) who has been my inspiration, and support over the years. My father (Saeid) who relentlessly encouraged me to achieve my dreams. To my confidant and sister (Behnaz) for her friendship, patience and for providing a distraction from the monotony of writing a dissertation. I would also like to thank my grandmother (maman-mali), grandfather (Babajoon) and great uncle (Ghorban amu).

Conducting research and developing a Ph.D. dissertation is never an individual effort. There are a number of individuals without whom my research would have been impossible this. I would like to acknowledge my advisors Dr. Thiel, and Dr. Tolaymat for their guidance, time, and ecouragment. Furthermore, I am grateful to other members of my committee Drs. Guliants, Sorial and Shanov. I also would like to thank Dr. Al-Abed. While he was not a member of my committee, he was instrumental in the development of many aspects of my research.

Additionally, I would like to further recognize Dr. Guliants for assistance with the admissions to the Ph.D. program at the University of Cincinnati and Mr. David Carson for introducing me to researchers who were pivotal in accomplishing my research goals. I am also grateful to Dr.

Arjmand, my undergraduate advisor, for his unconditional support and kindness over the years.

A special word of appreciation to the Pegasus Technical Services Inc. staff members: Dr.

Venkatapathy, Dr. El-Badawi, Mr. Tegenaw, Mrs. Yang, Dr. Huang and Dr. Arambewela. Finally,

I would like to thank the Center Hill Facility staff members: Miss. Goetz, Dr. Luxton, Dr.

Scheckel, Dr. Zimmer and Mr. Voit for their contributions.

Dedicated to my mother

Table of Contents

ABSTRACT…………………………………………………………………………………….………… i

ACKNOWLEDGEMENTS………………………………………………………………………..…….. iv

List of Tables…………………………………………………………………………………...... x List of Figures……………………………………………………………………………………….……. xi List of Abbreviations……………………………….……………………………………………..…...... xiv CHAPTER 1…………………………………………………………………………………...... 1 Introduction………………………………………………………………………………………………... 1 1.1 Nanotechnology …………………………..…………………………………………...... 2 1.2 Silver Nanomaterials (AgNPs)………………………………………………………...... 5 1.3 Disposal of Nanoparticles …………………………………………………………………….……... 9 1.4 Objectives…………………………………………………………………………………………… 11 1.5 Overview……………………………………………………………………………………………. 12 Literature Cited……………………………………………………………………………..…………..... 15 CHAPTER 2……………………………………………………………………………..……………..... 20 The Impact of Silver Nanoparticles on the Composting of Municipal Solid Waste………..…………… 20 Abstract…………………………………………………………………………………………………... 21 2.1 Introduction.…………………………………………...... 22 2.2 Experimental Section……………………………………………………………...…..…………..... 25 2.2.1 Nanoparticles, Selection, Synthesis, and Purification…………………………………………... 25 2.2.2 The Compost Reactors…………………………….………………………………...... 26 2.2.3 Composting Sampling and Analysis…………………………………………….…...... 26 2.2.4 DNA Sequencing and Bacterial Composition and Diversity Analyses……...……...... 27 2.2.5 Species Richness, Diversity, and Statistical Analysis of Microbial Communities……………... 29 2.3 Results and Discussion……………………………………………………………….……………... 29 2.3.1 AgNPs Characterization……………………………………………………………...... 29 2.3.2 Gas Quantity and Quality…………………………………………………………...... 30 2.3.3 Leachate Quality………………………………………………………………………………... 31 2.3.4 Solids Quality………………………………………………………………………...... 32 2.3.5 Bacterial Diversity Analyses……………………………………………………………………. 34 2.4 Environmental Implications……………………………………………………………...... 36 Literature Cited……………………………………………………………………………...... 37 APPENDIX A……………….…………………………………………………………………………… 49

A1. Purification of AgNPs suspensions……………………………………...…………………………... 50 A2. C: N Ratio Measurement…………………………………………………...………………………... 50 A3. PCR Methodology………………………………………………………….…………...... 51 A4. XPS Analysis………………………………………………………………….………...... 51 A5. XAS Analysis………………………………………………………………...…………………….... 51 CHAPTER 3………………………………………………………………………..…………………..... 66 Anaerobic Toxicity of Cationic Silver Nanoparticles………………………………...……...... 66 Abstract……………………………………………………………………………….……...... 67 3.1 Introduction………………………………………………………………………..…...... 68 3.2 Experimental Section………………………………………………………………...……………... 70 3.2.1 Nanoparticles Synthesis, Purification and Characterization………………………...... 70 3.2.2 Anaerobic Digesters Setup…………………………………………………………...………..... 70 3.2.3 Sampling and Analysis………………………………………………………………...……...... 71 3.2.3.1 Biogas Analysis…………………………………………………………………...……..... 71 3.2.3.2 Taxonomical Analysis……………………………………………………………...……... 71 3.2.3.3 Silver Speciation Analysis by X-ray Absorption Spectroscopy (XAS)…...... 72 3.2.3.4 Modelling Speciation of Dissolved Ag (MINTEQ)………………………...... 72 3.2.3.5 Statistical Analysis………………………………………………………...... 73 3.3 Results and Discussion…………………………………………………………………………...... 73 3.3.1 AgNPs Characterization……………………………………………………………………...... 73 3.3.2 Biogas Volume………………………………………………………………………………...... 73 3.3.3 Taxonomical Analysis……………………………………………………………...... 75 3.3.4 XAS (X-ray Absorption Spectroscopy)……………………………………………………...... 76 3.3.5 Speciation of Dissolved Metals in Anaerobic Sludge (MINTEQ)………………...... 77 3.4 Implications……………………………………………………………………………………...... 77 Literature Cited……………………………………………………………………………...... 79 APPENDIX B……………………………………………………………………………...... 87 B1. Synthesis of AgNPs…………………………………………………………………...... 88 CHAPTER 4………………………………………………………………………………...... 97 Nanosilver as a Disinfectant in Dental Unit Waterlines: Assessment of the Physicochemical Transformations of the AgNPs……………………………………………………………...... 97 Abstract……………………………………………………………………………………...... 98 4.1 Introduction…………………………………………………………………………...... 100 4.2 Experimental Methodology…………………………………………………………...... 102

vii

4.3 Results and Discussions………………………………………………………………………...…. 105 4.3.1 Microbial Analysis………………………………………………………………...... 105 4.3.2 Scanning Electron Microscopy (SEM)……………………………………………………...… 106 4.3.3 Transmission Electron Microscopy (TEM) and EDX……………………………………….... 107 4.3.4 Silver Speciation by XPS and XAS………………………………………………...... 109 4.4 Conclusion………………………………………………………………………………………… 110 Literature Cited……………………………………………………………………………...... 112 APPENDIX C………………………………………………………………………………………...… 126 C1. XPS Analysis…………………………………………………………………………...... 127 CHAPTER 5………………………………………………………………………………...... 135 Fate and Transportation of Silver Nanoparticles Released from Consumer Products: Ecological Risk Assessments..……………………………………………………………………………………… 135 Abstract……………………………………………………………………………………...... 136 5.1 Introduction…………………………………………………………………………...... 137 5.2 Characterization of AgNPs in Consumer Products….………………………………...... 138 5.2.1 Sample Collection and Characterization Techniques………………...... 139 5.2.1.1 Solid Materials………………………………………………………………………...… 139 5.2.1.2 Aqueous Materials…………………………………………………………………...….. 141 5.2.1.3 Aerosols…………………………………………………………………………...…...... 142 5.3 Past Research on AgNP-containing Consumer Products……………………………………….… 144 5.3.1 Solid Phase……………………………………………………………………………………. 144 5.3.2 Liquid/Aerosol Phase…………………………………………………………………………. 146 5.4 Potential Ecological Risks Associated with AgNPs………………………………………………. 148 5.4.1 Transformations of AgNPs in Aquatic Environments…………………………………...……. 148 5.4.1.1 Usage and Disposal Scenarios ………………………………………………………...... 149 5.4.1.2 Air……………………………………………………………………………………...... 156 5.4.1.3 Soils……………………………………………………………………………………… 157 5.5 Ecotoxicity Analyses………………………………………………………………………………. 157 5.5.1 Tools for Ecotoxicity Impact Evaluation…………...……....…………………………………. 157 5.6 Research Recommendations and Gaps…………...……………………………………………….. 158 5.6.1 Consideration of All Possible Interactions in a given Scenario ……....………………………. 159 5.6.2 Effects of Capping Agents…………………………..……....……………………………….... 159 5.6.3 Effects of Aging and Disposal Scenario……………..……....……………………………...... 159 5.6.4 Potential Transformations as a Function of Natural Environmental Factors………………..... 160

viii

5.6.5 Wastewater Treatment Conditions ………………….……………………………………….... 160 5.6.6 Aerosol and Atmospheric Conditions……………….……....……………………………….... 160 5.6.7 Ecotoxicity Analyses……………………………………....………………………………...... 161 Literature Cited…………………………………………………………………………………………. 162 CHAPTER 6………………………………………………………………………………………….… 179 Conclusions and Implications………………………………………………………………………...... 179 CHAPTER 7……………………………………………………………………………………....….… 191 Recommended Future Work……………………………………………………….………...…….…… 191 Literature Cited……………………………………………………………………………...... 196

List of Tables

Table 2.1 Results of Two-Way ANOSIM Test Based on Bray−Curtis Similarity Matrix Derived from the Distribution of Microbial Communities……………………………………………………...... 47

Table A.1 XPS analysis results……………………………………………………………………….…. 59

Table A.2 XAS speciation results…………………………………………………………………….…. 59

Table A.3 SIMPER analysis representing the top OTUs responsible for 50% of the differences between samples (i.e. treatments-week)…………………………………………………..… 63

Table 3.1 Ag speciation as identified by LCF of k-edge XANES spectra. Silver species proportions are presented as percentages. The R-factor indicates the goodness of fit………………...... 83

Table B.1 Biosolids Characteristics...... 89

Table B.2 Hydrodynamic diameter (HDD) and Zeta potential of particles...... 90

Table B.3 Most abundant bacterial species over 28 day period……………………………………….... 93

Table B.4 MINTEQ Speciation Output…………………………………………………………………. .95

Table 4.1 Microbial analysis of dental tubing and corresponding effluent water (CFU/ml) after 2 weeks, 1 month, 10 weeks, 3.5 months and 5.5 months of continuous tap water flow………………. 116

Table 4.2 Microbial analysis results of the dental tubing and corresponding effluent water before and after the disinfection process (re-circulation with 2 ppm nanosilver for 3 days)…………………... 117

Table 5.1 Summary of past research using AgNP-containing consumer products...... 170

List of Figures

Figure 1.1 Silver Nanoparticle Product Distribution……………………………………………………. 19

Figure 2.1 Schematic of the experimental setup of the compost reactors……………………………. 42

Figure 2.2 TEM image of PVP-coated AgNPs………………………………………………………….. 43

Figure 2.3 Average gas levels in composters’ headspace (a) O2, (b) CO2, and (c) N2O……………….... 44

Figure 2.4 The Ag Kα XAS spectra of AgCl, Ag Cystine, and Ag Humic as pure phases and

PVP-AgNPs reacted with the compost material and AgNO3 with the compost leachate……………...… 45

Figure 2.5 The relative distribution of bacterial classes. Percentages are shown for control (C), Ag+ treated (A), and AgNP treated (B) samples…………………………………………………………. 46

Figure 2.6 Richness for OTUs at 99% sequence identity with rare members included (a, b) and after rare members were removed (c, d)…………………………………………………………………. 48

Figure A.1 pH levels of composter leachate…………………………………………………………….. 54

Figure A.2 Conductivity of composter leachate………………………………………………………… 54

Figure A.3 Total Organic Carbon (TOC) concentration of composter leachate………………………… 55

Figure A.4 Chemical Oxygen Demand (COD) concentration of composter leachate………………...… 55

Figure A.5 Ammonia-Nitrogen concentration of composter leachate...... 56

Figure A.6 Total silver concentration in composter leachate...... 56

Figure A.7 Chloride concentration of composter leachate………...... 57

Figure A.8 C:N ratio of composter solid samples……………………………………………………….. 57

Figure A.9 Total silver concentration of composter solid samples……………………………………….58

Figure A.10 nMDS based on total community (120699 OTUs at 99% cutoff value), 1) A, B, C are the negative control composters, 2) D, E, F are the composters treated with Ag+, and 3) G, H, P are composters treated with AgNPs…………………………………………………………... 60

Figure A.11 nMDS based on total community (9368 OTUs at 99% cutoff value with no RARE members), 1) A, B, C are the negative control composters, 2) D, E, F are the composters treated with Ag+, and 3) G, H, P are composters treated with AgNPs………………………………………...... 60

Figure A.12 Shannon diversity for OTUs at 99% sequence identity with rare members included (a, b) and after rare members were removed (c, d)…………………………………………….. 61

Figure A.13 Evenness for OTUs at 99% sequence identity with rare members included (a, b) and after rare members were removed (c, d)…………………………………………………………………. 62

Figure A.14 Components in each composter (%w/w)…………………………………………………... 64

Figure 3.1 Cumulative gas volume production of anaerobic biosolids after exposure to different concentrations of AgNPs or Ag+ for a duration of 28 days…………………………………………...... 84

Figure 3.2 Flavobacterium species among different treatments at 5 mgL-1 of silver...... 85

Figure 3.3 Ag-Kα XAS spectra of Ag2S, Ag metallic as pure phases and BPEI-AgNPs reacted with the anaerobic biosolid material at natural (15 mgL-1) and elevated (100 mgL-1) sulfide levels over a 24 hr reaction period……………………………………………………………………..... 86

Figure B.1 TEM image of Citrate-AgNPs...... 90

Figure B.2 TEM images of PVP-AgNPs...... 91

Figure B.3 TEM images of BPEI-AgNPs...... 92

Figure B.4 Purification of AgNPs………………………………………………………………………. 96

Figure 4.1 SEM images of unused dental tubing surface. Images A) and B) comprise of the large view and detailed texture of tubing surface, respectively...... 118

Figure 4.2 SEM image of the biofilm on dental tubing surface after 19 days (A) and 4 months (B) of continuous tap water flow...... 119

xii

Figure 4.3 Backscattered SEM (A) and secondary electron SEM (B) images of the same area of biofilm. The backscattered SEM image shows the clear contrast between Ag nanoparticles and biofilm surrounding………………………………………………………………………………… 120

Figure 4.4 TEM image of original commercial ASAP-AGX-32 nanosilver (A) and corresponding EDX spectra (B)………………………………………………………………………………………… 121

Figure 4.5 TEM images (A and C) of nanosilver after re-circulation through dental tubing for 3 days……………………………………………………………………………………………...… 122

Figure 4.6 Ag 3d XPS spectra of pristine Ag nanoparticles…………………………………………… 124

Figure 4.7 A) Ag-Kα XAS spectra of Ag2S, Ag2O, AgCl and Ag metallic as pure phases and AgNPs (ASAP-AGX-32) after the DUWL disinfection process. B) Ag speciation as identified by LCF of k-edge XANES spectra…………………………………………………………………...… 125

Figure C.1 Schematic of the Dental Unit Water Line (DUWL)-Continuous tap water flow system….. 128

Figure C.2 Schematic of the disinfection process through circulation of the nanosilver solution through the dental tubing. (Re-circulation system)……………………………………...……………… 129

Figure C.3 High magnification of backscattered SEM image of a silver nanoparticle attached to the biofilm surface...... 130

Figure C.4 TEM images of the pristine silver nanoparticle solution (ASAP-AGX-32)………………. 131

Figure C.5 SEM images of the inner surface of dental tubing water lines (DUWL) after 19 days of continuous tap water flow…………………………………………………………………... 132

Figure C.6 SEM images of the inner surface of dental tubing water lines (DUWL) after 4 months of continuous tap water flow...... 133

Figure C.7 TEM images of 2 ppm nanosilver solution (ASAP-AGX-32) after the dental tubing treatment (disinfection process) at various magnification levels. A) 50k B) 200k C) 400k...... 134

Figure 5.1 AgNPs bearing consumer products listed in consumer product inventory of Woodrow Wilson International center for scholars and project on emerging nanotechnology [3] as of March, 2016……………………………………………………………………………………… 178

xiii

List of Abbreviations

AgNPs Silver Nanoparticles

BE Binding Energy

BPEI Branched polyethyleneimine

CFU Colony Forming Units

C/N Carbon to Nitrogen Ratio

COD Chemical Oxygen Demand

DLS Dynamic Light Scattering

DNA Deoxyribonucleic Acid

DO Dissolved Oxygen

DUWL Dental Unit Water Line

EDX Energy-dispersive X-ray Spectroscopy

ENM Engineered Nanomaterial

ENP Engineered Nanoparticle

EPA Environmental Protection Agency

EPS Extracellular Polymeric Substance

FID Flame Ionization Detector

GF-AA Graphite Furnace-Atomic Absorption

HA Humic Acids

HDD Hydrodynamic Diameter

LCA Life Cycle Assessment

MSW Municipal Solid Waste

xiv

NOM Natural Organic Matter

NP Nanoparticle

OTU Operational Taxonomic Unit

PCR Polymerase Chain Reaction

PVP Polyvinylpyrrolidone

ROS Reactive Oxygen Species

SEM Scanning Electron Microscopy

SPR Surface Plasmon Resonance

TCD Thermal Conductivity Detector

TEM Transmission Electron Microscopy

TOC Total Organic Carbon

WWTP Wastewater Treatment Plant

XANES X-ray Absorption Near-Edge Structure

XAS X-ray Absorption Spectroscopy

XPS X-ray Photoelectron Spectroscopy

XRD X-ray Diffraction

Chapter 1

Introduction

1.1 Nanotechnology

Nanotechnology refers to the study, application, and characterization of materials, structures and devices with a size on the nano-scale. By definition, nanotechnology involves the manipulation of matter in a way that produces structures with at least one dimension ranging from 1-100 nm in size; these materials have unique physicochemical properties that differentiate them from their bulk counterparts (1, 2). Nanomaterials can be divided into different categories according to their origin, dimensions and elemental composition. Based on their origin, nanomaterials are divided into naturally-occurring and engineered nanomaterials (ENMs). Naturally-occurring nanomaterials are those commonly produced by nature (e.g., volcanic ash, ocean spray, soot from forest fires, viruses and protein molecules). Engineered nanomaterials (ENM) refer to materials that are engineered by manipulation of matter at the nano-scale (3). ENMs are the main focus of the research reported here. Silver nanoparticles for antibacterial applications or carbon nanotubes for the strengthening of materials are examples of ENMs. Furthermore, a subgroup of nanomaterials exists recognized as incidentally occurring nanomaterials that are the byproduct of man-made processes (e.g., welding fumes, diesel engine exhaust, cigarette smoke and paint pigments). Nanomaterials are currently employed in a wide range of applications and fields, including the medical, chemical, pharmaceutical and environmental industries (4). The objective of nanoscience is to explore and exploit unique chemical, physical, electrical and mechanical properties that emerge when matter is structured at the nanoscale (5).

The use of ENMs is continuously growing in a wide variety of applications ranging from the utilization of carbon nanotubes for the production of vehicle composites that are stronger but considerably lighter than steel (thereby improving fuel economy), to creating medicines capable of targeting specific cells in the body, to using silver nanoparticles in filtration systems for

2 purification of water at the point of collection (6). In addition to the advantages of ENMs due to their nano-specific applications (e. g., AgNPs for antibacterial applications AgNPs or carbon nanotubes for strengthening of materials), universal environmental benefits are associated with the use of nanomaterials as a result of saving in raw materials, less consumption of natural resources and therefore a reduction in environmental pollution and exploitation (7).

According to the Woodrow Wilson database, the number of nanomaterial-containing consumer products in the US increased from 54 products in 2006 to 1015 products in 2010 (8). In a more recent survey, Vence et. al., identified 1814 nanomaterial-containing consumer products in the market as of March 2015. This represents a thirty-fold increase compared to the 54 products originally listed in 2005 (9).

Two primary factors are responsible for the increased interest in ENMs relative to their bulk material counterparts. First, nano-scale solids contain a higher fraction of atoms at their surfaces than bulk materials. This phenomenon causes an increase in the specific surface area, leading to a significant increase in the surface reactivity. Second, at the nano-scale, quantum confinement effects (electrons confined in small area) govern the properties and interactions of matter. The physicochemical and intrinsic properties of nanomaterials can be significantly different from those of the same material (elemental composition) in bulk form (10). For example, silver nanoparticles

(AgNPs) exhibit unique optical properties known as surface plasmon resonance (SPR), oscillations that result in unusually strong scattering and absorption properties, which are absent in bulk silver.

Although a universal classification of nanomaterials has yet to become official, nanomaterials are commonly categorized according to their structural dimensions and elemental composition. On the basis of structural dimensions, nanomaterials are categorized as zero-dimensional, one- dimensional, two-dimensional or three-dimensional nanostructures (11). Zero-dimensional (0-D)

3 nanostructures demonstrate nano-dimension in all three directions (e.g., spherically shaped nanoparticles). One-dimensional (1-D) nanostructures contain a formation in which one dimension exceeds the nanometer range (e.g., nanowires, nanotubes and nano rods of different elemental composition). This definition consistently carries over to two-dimensional (2-D) and three- dimensional (3-D) nanostructures. Examples of (2-D) nanostructures include nanofilms, nanosheets and nanowalls. In (3-D) nanostructures, all the structural dimensions exceed the nanometer range, however, they are made up of individual building blocks that are in the nano- range, and therefore still referred to as nanostructures (e.g., aggregates and agglomerates forming nanoclusters).

Furthermore, on the basis of elemental composition, nanomaterials are classified into different groups, among which the most commonly reported include carbon-based nanomaterials (e.g., single and multi-walled carbon nanotubes), metal-based nanomaterials (metallic, alloy and metal- oxide nanomaterials), dendrimers and composites.

Inside the unlimited diversity of these materials, some nanomaterials may be toxic to biological species, others are relatively benign and yet others are produced for their health benefits. It must be noted that the same properties that make ENMs unique with beneficial technological applications may also endanger human beings directly through the potential induction of cyto- and genotoxic effects, inflammation and cancer (12, 13). Additionally, the environmental health and hazard risks associated with applications of nanotechnology for industrial and commercial uses are not fully known (14, 15). Furthermore, the physicochemical properties associated with ENMs not only differ from the corresponding bulk matter, but also between different forms of ENMs with the same chemical composition (16, 17). ENMs containing the same chemical or elemental composition with different morphological characteristics (e.g., size, shape and surface properties)

4 can exhibit different physicochemical properties (18). This phenomenon further complicates the investigation of these materials as a plethora of ENMs from many materials, in many forms and sizes with a variety of surface coatings are possible. Thus, the potential risks and benefits associated with the utilization of ENMs must be evaluated concentrating on the specific physicochemical characteristics of each nanomaterial. Furthermore, potential alterations to the physicochemical properties of ENMs, caused by time or environmental conditions, may lead to dissimilar environmental and public health implications, and therefore, must be taken into consideration.

1.2 Silver Nanomaterials (AgNPs)

Humankind has known of the antimicrobial properties of silver for over 2000 years. Silver and silver compounds have been utilized for their useful properties in an extensive range of applications (19). Archaeological evidence dating back to 3000 B.C. suggests ancient Persians used silver vessels for water storage to prolong freshness. Similarly, ancient Greeks used silver for healing wounds caused in battle as they were aware of silver’s antimicrobial capabilities. In the

19th century, silver was applied for medicinal purposes in curing common ailments pertaining to the eyes and skin. The US Food and Drug Administration (FDA) approved silver solutions to be used as antibacterial agents in the 1920’s. Silver applications have continued to the 21st century and have been influenced by the development of nanoscience. However, nanosilver is not a new discovery as it has been known and utilized for over 100 years (20, 21); nanosilver was previously referred to as colloidal silver.

Understanding the applications of silver nanoparticles will assist in the prediction of potential exposure routes into the environment. Currently over 50% of silver containing products registered with the USEPA contain silver in nano-form (22). Due to their novel properties, silver

5 nanoparticles have attracted much attention and are applied in diverse areas, including medicine, catalysis, textile engineering, biotechnology and bioengineering, water treatment, electronics and optics (19). Moreover, silver nanoparticles are known to have a strong antibacterial effect for an extensive array of organisms such as viruses, bacteria, fungi and other pathogenic organisms (23).

Interestingly, silver intrinsically demonstrates higher toxicity to microorganisms and lower toxicity to mammalian cells than other metals (24).

Hence, silver nanoparticles are widely used as antibacterial/antifungal agents in a wide range of consumer products such as air sanitizer sprays, socks, pillows, slippers, respirators, wet wipes, detergents, soaps, shampoos, toothpastes, air filters, refrigerator coatings, vacuum cleaners, washing machines, food storage containers, cellular phones and childrens toys (Table 1.1). The rate of applications of silver nanoparticles (AgNPs) has increased, causing them to become the most commercialized nanomaterial today, primarily due to their antibacterial/antifungal applications (25).

The properties of nanomaterials are highly dependent on their method of synthesis. The synthesis method determines the nanoparticle’s shape, size distribution and surface coating. The physiochemical properties of AgNPs are commonly tailored to meet the demands of application by manipulating the synthesis method. A general evaluation of the synthesis processes of silver nanoparticles is important from an environmental standpoint as it allows for the identification and characterization of the most dominantly produced particles. Silver nanoparticles are synthesized using various techniques, resulting in different shapes, sizes and surface properties. AgNP synthesis techniques can be divided into two main categories: “top-down” and “bottom-up” (26).

The principle behind the “top-down” technique is to start with a metal in bulk and mechanically reduce its size to the nano-scale by utilizing various approaches such as evaporation/condensation,

6 laser ablation and lithography (27). In contrast, the “bottom-up” technique begins with atoms or molecules and builds up to nanostructures.

The “bottom-up” synthesis techniques generally involve the dissolution of a silver salt in a solvent, followed by the subsequent addition of a reducing agent, and the supplemental use of stabilizing

(capping) agents. Capping agents are used to prevent ENM aggregation (28). This tendency to aggregate is attributed to the excess surface charge and high thermodynamic instability of the

ENMs (29). The bottom-up synthesis of AgNPs through chemical reduction produces homogenous nanoparticles with regular shapes that can contain impurities carried over from the synthesis process (e, g,. Ag+, residual reducing agents and stabilizing agents) and therefore require an additional step for purification. The top-down techniques have far fewer applications than the bottom-up techniques for the synthesis of AgNPs, mainly because particles synthesized with top- down methods have been reported to contain surface imperfections and poorly controlled particle morphologies (23, 30). Bottom-up synthesis is far more extensively used in the manufacturing process (28). As a result, the AgNPs utilized in the research reported here were synthesized through chemical reduction (bottom-up) techniques.

In principle, countless combinations of AgNPs with various physicochemical properties can be synthesized by using different reactants in the synthesis process. Nevertheless, the stabilization mechanisms used to prevent particles from aggregating fall under one of three main categories: electrostatic, steric or electrosteric stabilization (a combination of electrostatic and steric stabilization) (31). Stabilization is achieved by utilization of capping (stabilizing) agents in the synthesis process. Electrostatic (charge) stabilization is achieved by the adsorption of counterions on the colloidal particle surface or by the protonation of specific functional groups adsorbed on the particle surface. Steric stabilization is achieved by adsorption of nonionic surfactants or

7 polymers onto the nanoparticle surface resulting in repulsion between particles. The polymers utilized may be uncharged or polyelectrolytes. The combined effects of electrostatic and steric stabilization mechanisms lead to electrosteric stabilization. Electrosteric stabilization is generally associated with the adsorption of polymers onto the particle surface similar to steric stabilization; however, the adsorbed polymers contain non-negligible electrostatic charge (32).

While silver has been utilized for antibacterial properties for centuries, its exact toxicological mechanism of action remains unknown. Many researchers hypothesize the toxic effects of silver to be proportional to the rate of release of monovalent silver ions. Silver nanoparticles on the other hand have intrinsic characteristics such as size, surface to volume ratio, solubility, aggregation ability, chemical composition and surface chemistry that differentiate them from the characteristics of bulk silver. Due to their different physicochemical properties in comparison to the material in bulk, it cannot be ruled out that the increased reactivity of AgNPs (due to their high surface area) may lead to increased toxicity.

Many studies have been conducted in an attempt to explain the mechanisms by which nanosilver exerts its antimicrobial activity (33-36), however, the exact mechanisms of toxicity still remain unclear. According to the literature, a number of mechanisms for toxicity have been commonly proposed. The proposed mechanisms include AgNPs toxicity due to the release of silver ions through silver nanoparticle dissolution under aerobic conditions, as well as, nano-specific properties and interactions leading to cell membrane damage, such as nanoparticle transportation by a Trojan-horse type mechanism followed by the generation of Reactive Oxygen Species (ROS)

(33-36). The AgNP-specific mechanism of toxicity, may occur through interaction/attachment with the external membrane or as a result of penetration of into the cell. The attachment of AgNPs to the cell membrane may cause cell lysis through physically damaging the integrity of the cell or

8 by interrupting the permeability and metabolic pathways essential for its survival. Cell permeation by AgNPs can cause cell malfunction through stimulating the production of reactive oxygen species (ROS) and interaction with specific proteins, enzymes and DNA structures inhibiting their activity and replication (37-42). Numerous research studies have reported the toxicity of AgNPs on a widespread array of organisms, mammalian, and human cells (43, 44).

1.3 Disposal of Nanoparticles

Silver nanoparticles, like many other nanoparticles, have the potential to be released into the environment at various points throughout their lifecycle. As a route of environmental exposure, products containing nanoparticles are often disposed of as part of the municipal solid waste stream.

In a disposal scenario, nanoparticles (e.g., AgNPs) may leach from their mother-product into the organic fraction of the solid waste and end up in composting systems or even enter the solid waste stream with biosolids and end up in landfills and wastewater treatment plants. In view of mounting evidence suggesting exponential growth in AgNP consumer applications, there are growing concerns in the research community in regards to the potential impacts of AgNPs on waste management systems that rely on microbial degradation for waste treatment as any interference in these systems can lead to a wide-range calamity.

In the past decade a plethora of research has been conducted on a laboratory scale to understand the fundamental factors governing the environmental impacts of silver nanoparticles. It has been well established that the physicochemical properties of nanomaterials vary drastically from the bulk form of the same material in bulk form; knowledge about the bulk material may not apply to the nanomaterial. Additionally, nanoparticles are manufactured with a wide range of physicochemical properties (size distributions, shapes and surface charge scenarios) that are known to impact their fate, transport and toxicity in the environment (18). Since the mechanisms

9 of toxicity of different silver nanoparticles have been shown to vary under different environmental conditions (1), identification of the different disposal scenarios (routes of environmental exposure) is a key factor in the assessment of the environmental implications of AgNPs.

Due to the lack of regulations, products containing nanoparticles are often disposed of in the municipal waste stream. For instance, AgNPs are reportedly released into both natural systems (e.g., surface water streams, soils and sediments), as well as, engineered systems (e.g., water and wastewater treatment plants, composting systems and landfills). To assess the concerns associated with the environmental impacts of AgNPs, a substantial amount of research on the toxicity of

AgNPs to a wide array of bacterial species has been conducted. However, these assessments were completed under simple laboratory conditions, far from simulating real environmental conditions.

Hence, questions on their impacts on complex environmental systems remains largely unknown.

To date only few studies have started addressing this issue and research is still in its infancy stage.

Therefore, more research is required to fulfill the data gaps related to the toxicity of AgNPs on environmentally relevant media. Also because of the diversity of AgNPs, possessing different physicochemical properties, any direct comparison with the intent of a one size fits all approach would be meaningless.

Hence, research thus far is insufficient to address the antimicrobial impacts and toxicity of AgNPs, for two major reasons: First, the long list of synthesis methods, which utilize different dispersants, reducing agents and capping agents yield AgNPs with a wide variety of physicochemical characteristics (e.g., size, shape, surface charge). In view of the extensive of literature suggesting the toxicity of AgNPs depends on the physicochemical characteristics of the nanoparticles, generalizing a particular form of AgNPs to be representative of all silver nanoparticles may not be appropriate.

Second, in addition to the errors brought by investigating a single form of AgNP as a representative particle, most research assessing the implications of silver nanoparticles on the environment have primarily investigated their impacts on simple systems (environmentally relevant matrices containing few constituents). These simple systems do not represent the most likely scenarios and end of life circumstances in which the silver nanoparticles are to experience and therefore do not portray the entire story.

1.4 Objectives

The primary goal of this research was to explore the end of life scenario and potential impacts of silver nanoparticles on relevant environmental systems. To pursue this goal, the most likely disposal scenarios were evaluated by understanding and considering the mechanisms of toxicity of silver differ under aerobic and anaerobic conditions. Therefore, a thorough evaluation of the process in an example environmental model was followed. Additionally, as laboratory synthesized nanoparticles may not be representative of nano-enabled consumer products, a lifecycle assessment (LCA) of a commonly utilized nanosilver solution was undertaken.

The specific objectives of this research are:

1. To investigate of the impact of silver nanoparticles on the aerobic degradation process by

studying the impact of silver nanoparticles on the composting of municipal solid waste.

2. To evaluate of the toxicity of different silver nanoparticles under anaerobic conditions by

studying the anaerobic toxicity of cationic silver nanoparticles.

3. To assess the fate, transport and physicochemical transformations of a commonly used

consumer product containing silver nanoparticles for the disinfection of dental unit water

lines.

4. To assess the current state of research investigating the fate, transport and environmental

impacts of silver nanoparticles released from consumer products.

1.5 Overview

The objectives of this dissertation are addressed in 7 chapters:

Chapter 1 is an overview of the research topic, scientific problems and objectives.

Chapter 2 reports the findings of a broad study evaluating the impact of polyvinylpyrrolidone

(PVP) coated silver nanoparticles (PVP-AgNPs) on the composting of municipal solid waste. PVP-

AgNPs were used as sterically stabilized NPs have shown minimal aggregation in high ionic strength solutions and therefore resilient to aggregation under the experimental ionic strength. The study also examines relatively low concentrations of AgNP that may be encountered in a real world scenario and how such exposure may impact the function and composition of microbial communities associated with compost samples.

Chapter 3 reports the findings of a study investigating the effect of different AgNPs on microorganisms responsible for the anaerobic degradation of waste. These microorganisms play a major role in the treatment of different waste streams such as sludge digestion in wastewater treatment plants (WWTP) as well landfills containing municipal solid waste. The AgNPs were chosen in a manner to represent different surface charge scenarios as well as stabilization mechanisms. Furthermore, the toxicity of the different AgNPs were compared to that of Ag+ under anaerobic conditions at the same mass concentration.

Chapter 4 reports the findings of a broad investigation of the fate, transport and transformations of a nanosilver containing product commonly used in dental practices for the disinfection of dental lines. The life-cycle assessment (LCA) of the commercially available nanosilver solution was

12 performed by simulation of a dental unit water delivery system. Allowing the impact of the nanosilver solution on the biofilm developments to be investigated under a real-life scenario.

Biofilms are intricate communities of surface-bound microorganisms embedded in a self-produced matrix of extracellular polymeric substance (EPS). The biofilm formation acts as an efficient barrier and protective shield against antimicrobial agents, resulting in a dense undisturbed colonization of the biofilm forming bacteria. Since silver nanoparticles are known to inhibit the formation of biofilms (45), and yet the structure of the bacterial community is designed in a manner to resist bacterial inhibition this effort was taken and of importance and differentiates this chapter from the rest. Additionally, the evaluation of the potential implications of engineered nanomaterials (ENM) to this date have been exclusively focused on pristine, freshly synthesized nanoparticles. However, it is our belief that when considering the life-cycle assessment it is of more relevance to focus on the nanoparticles released from actual commercially available products in different phases of their lifespan such as manufacturing, usage scenario and disposal.

Chapter 5 reports the findings of a comprehensive review addressing the current state of research on nanosilver containing consumer products, as well as, identification of the data gaps. In an effort to better understand the fate, transport and impacts of AgNPs under environmental scenarios, a plethora of studies have been conducted in the past decade. However, these research efforts have been mainly focused on pristine AgNPs either lab-synthesized or purchased as pure commercially available material. Although these studies have contributed to a more fundamental understanding of AgNPs in terms of mechanisms of toxicity and behavior under various environmental conditions, haven’t addressed the real environmental concern associated with AgNPs, which is most likely to occur through their release by consumer products. In an effort to demonstrate the current state of research on actual consumer products, this review will assess previous work reported on

13 characterization techniques, routes of environmental exposure, potential ecological risks of AgNPs, as well as, identification of available toxicity assays currently utilized which may assist in implementing toxicity assessments on AgNPs in consumer products. The main objectives of this review are summarizing the major findings of past research related to AgNPs released from CPs, identifying the issues and knowledge gaps and providing recommendations for future research.

Chapter 6 contains the overlying conclusions and key findings pertaining to each objective.

Chapter 7 is a compilation of data gaps identified in the research and literature, as well as, work laying the groundwork for future work recommended to address some of these gaps in regards to the environmental impact, fate and transformations of AgNPs.

Literature Cited

1. El Badawy, A. M.; Luxton, T. P.; Silva, R. G.; Scheckel, K. G.; Suidan, M. T.; Tolaymat, T. M. Impact of environmental conditions (pH, ionic strength and electrolyte type) on the surface charge and aggregation of silver nanoparticles suspension. Environ. Sci. Technol. 2010, 44, 1260-1266.

2. SCENHIR. Opinion on: the scientific aspects of the existing and proposed definition relating to products of nanoscience and nanotechnologies. Brussels: European commission health and consumer protection directorate general. 2007a.

3. Colvin, V. L. The potential environmental impact of engineered nanomaterials. Nat. Biotechnol. 2003, 21, 1166-1170.

4. Ju-Nam, Y.; Lead, J. R. Manufactured nanoparticles: an overview of their chemistry, interactions and potential environmental implications. Sci. Total Environ. 2008, 400, 396-414.

5. Buzea, C.; Pacheco, I. I.; Robbie, K. Nanomaterials and nanoparticles: Sources and toxicity. Biointerphases. 2007, 2 (4), 17-71.

6. Adlakha-Hutcheon, G.; Khaydarov, R.; Korenstein, R; Varma, R.; Vaseashta, A.; Stamm, H.; Abdel-Mottaleb, M. Nanomaterials, nanotechnology: applications, consumer products and benefits. I. Linkov, J. Steevens (Eds.), Nanomaterials: Risks and Benefits, Springer, Dordrecht. 2009, 195–207.

7. Kuhlbusch, T. A. J.; Fissan, H.; Asbach, C. Nanotechnologies and Environmental risks: measurement technologies and strategies. I. Linkov, J. Steevens (Eds.), Nanomaterials: Risks and Benefits, Springer, Dordrecht. 2009, 233–243.

8. Project on Emerging Nanotechnologies. Silver Database. Washington, DC. Project on Emerging Nanotechnologies. Woodrow Wilson International Center for Scholars. 2013. Available: http://goo.gl/tCKX8 [accessed 25 June 2013].

9. Vence, M. E.; Kuiken, T.; Vejerano, E. P.; McGinnis, S. P.; Hochella Jr, M. F.; Rejeski, D.; Hull, M. S. Nanotechnology in the real world: Redeveloping the nanomaterial consumer products inventory. Beilstein J. Nanotechnol. 2015, 6, 1769-1780.

10. Roduner, E. Size matters: why nanomaterials are different. Chem. Soc. Rev. 2006, 35, 583– 592.

11. Muraviev, D. N. Inter-matrix synthesis of polymer stabilized metal nanoparticles for sensor applications. Contib. Sci. 2005, 3, 19-32.

12. Borm, P, J, A.; Robbins, D.; Haubold, S.; Kuhlbusch, T.; Fissan, H.; Donaldson, K.; Schins, R.; Stone, V.; Kreyling, W.; Lademann, J.; Krutmann, J.; Warheit, D.; Oberdörster, E. The

potential risks of nanomaterials: a review carried out for ECETOC. Part. Fibre Toxicol. 2006, 3, 11.

13. Yang, H.; Liu, C.; Yang, D.; Zhang, H.; Xi, Z. Comparative study of cytotoxicity, oxidative stress and genotoxicity induced by four typical nanomaterials: the role of particle size, shape and composition. J. Appl. Toxicol. 2009, 29, 69–78.

14. Stone, V.; Hankin, S.; Aitken, R.; Aschberger, K.; Baun, A.; Christensen, F.; Fernandes, T.; Hansen, S.F.; Hartmann, N.B.; Hutchinson, G. Engineered Nanoparticles: Review of Health and Environmental Safety (ENRHES); Project Final Report; European Commission: Brussels, Belgium, 2010; pp. 350-365.

15. Savolainen, K.; Pylkkänen, L.; Norppa, H.; Falck, G.; Lindberg, H.; Tuomi, T.; Vippola, M.; Alenius, H.; Hämeri, K.; Koivisto, J.; Brouwer, D.; Mark, D.; Bard, D.: Berges, M.; Jankowska, E.; Posniak, M.; Farmer, P.; Singh, R.; Krombach, F.; Bihari, .; Kasper, G.; Seipenbusch, M. Nanotechnologies, engineered nanomaterials and occupational health and safety – A review. Safety Science. 2010, 48, 957-963.

16. Nowack, B.; Bucheli, T.D. Occurrence, behavior and effects of nanoparticles in the environment. Environ. Pollut. 2007, 150, 5-22.

17. Schwirn, K.; Tietjen, L.; Beer, I. Why are nanomaterials different and how can they be appropriately regulated under REACH? Environmental Sciences Europe. 2014, 26, 4.

18. Simon-Deckers, A.; Loo, S.; Mayne-L’hermite, M.; Herlin-Boime, N.; Menguy, N.; Reynaud, C.; Gouget, B.; Carriere, M. Size-, composition- and shape-dependent toxicological impact of metal oxide nanoparticles and carbon nanotubes toward bacteria. Environ. Sci. Technol. 2009, 43, 8423–8429.

19. Khaydarov, R. R.; Khaydarov, R. A.; Estrin, Y.; Evgrafova, S.; Scheper, T.; Endres, C.; Cho, S. Y. Silver Nanoparticles Environmental and Human Health Impacts. Nanomaterials: Risks and Benefits. 2009, 287-297.

20. Seltenrich, N. Nanosilver: Weighing the Risks and Benefits. Environ. Health Perspect. 2013, 121, A220– A225.

21. Nowack, B.; Krug, H. F.; Height, M. 120 Years of Nanosilver History: Implications for Policy Makers. Environ. Sci. Technol. 2011, 45, 1177–1183.

22. Rosalind, V. EPA has safely regulated nanosilver for decades http://nanotech.lawbc.com/uploads/file/00054380.PDF. 2009

23. Kildeby, N.L.; Roge, R. E.; Larsen, T.; Petersen, R.; Riis, J. F.; Bozhevolnyi, S. I. Silver nanoparticles. Project Group N344. Faculty of Physics and Nanotechnology, Aalborg University; 2005.

24. Zhao, G.; Stevens, S. E. Multiple parameters for the comprehensive evaluation of the susceptibility of Escherichia coli to the silver ion. Biometals. 1998, 11, 27–32.

25. Grosse, S.; Evje, L.; Syversen, T. Silver nanoparticle-induced cytotoxicity in rat brain endothelial cell culture. Toxicology in vitro. 2013, 27, 305-313.

26. del Rocío Balaguera-Gelves, M. Detection of nitroexplosives by surface enhanced Raman spectroscopy on colloidal metal nanoparticles. 2006, Master thesis, University of Puerto Rico, Mayaguez Campus.

27. Amendola, V.; Polizzi, S.; Meneghetti, M. Free silver nanoparticles synthesized by laser ablation in organic solvents and their easy functionalization. Langmuir. 2007, 23, 6766-6770.

28. Tolaymat, T. M.; El Badawy, A. M.; Genaidy, A.; Scheckel, K. G.; Luxton, T. P.; Suidan, M. An evidence-based environmental perspective of manufactured silver nanoparticle in syntheses and applications: A systematic review and appraisal of peer-reviewed scientific papers. Sci. Total Environ. 2010, 408, 999-1006.

29. Olenin, A. Y.; Krutyakov, Y. A.; Kudrinskii, A. A.; Lisichkin, G. V. Formation of surface layers on silver nanoparticles in aqueous and water-organic media. Colloid J. 2008, 70, 71-76.

30. Kildeby, N. L.; Roge, R. E.; Larsen, T.; Petersen, R.; Riis, J. F.; Bozhevolnyi, S. I. Silver nanoparticles 2005, Project Group N344, Faculty of Physics and Nanotechnology, Aalborg University.

31. Jiang, J.; Oberdorster, G.; Biswas, P. Characterization of Size, Surface Charge, and Agglomeration State of Nanoparticle Dispersions for Toxicological Studies. J. Nanopart. Res. 2009, 11, 77-89.

32. Holmberg, K.; Jonsson, B.; Kronberg, B.; Lindman, B. Surfactants and Polymers in Aqueous Solution. John Wiley & Sons Ltda, Chichester, 2002.

33. Marambio-Jones, C.; Hoek, E. M. V. A review of the antibacterial effects of silver nanomaterials and potential implications for human health and the environment. J. Nanopart. Res. 2010, 12, 1531-1551.

34. Choi, O.; Deng, K. K.; Kim, N. J.; Ross, L.; Surampalli, R. Y.; Hu, Z. The inhibitory effects of silver nanoparticles, silver ions and silver chloride colloids on microbial growth. Water. Res. 2008, 42, 3066-3074.

35. Xu, H.; Qu, F.; Xu, H.; Lai, W.; Wang, Y. A.; Aguilar, Z. P.; Wei, H. Role of reactive oxygen species in the antibacterial mechanism of silver nanoparticles on Escherichia coli O157:H7. Biometals. 2012, 25, 45-53.

36. Sotiriou, G. A.; Pratsinis, S. E. Antibacterial activity of nanosilver ions and particles. Environ. Sci. Technol. 2010, 44, 5649-5654.

37. Navarro, E.; Piccapietra, F.; Wagner, B.; Marconi, F.; Kaegi, R.; Odzak, N.; Sigg, L.; Behra, R. Toxicity of silver nanoparticles to Chlamydomonas reinhardtii. Environ. Sci. Technol. 2008, 42, 8959–8964.

38. Choi, O.; Clevengera, T. E.; Deng, B.; Surampalli, R. Y.; Ross, L.; Hu, Z. Role of sulfide and ligand strength in controlling nanosilver toxicity. Water Res. 2009, 43, 1879–1886.

39. Fabrega, J.; Fawcett, S. R.; Renshaw, J. C.; Lead, J. R. Silver nanoparticle impact on bacterial growth: Effect of pH, concentration, and organic matter. Environ. Sci. Technol. 2009, 43, 7285–7290.

40. Skebo, J. E.; Grabinski, C. M.; Schrand, A. M.; Schlager, J. J.; Hussain, S. M. Assessment of metal nanoparticle agglomeration, uptake, and interaction using high-illuminating system. Int. J. Toxicol. 2007, 26, 135–141.

41. Choi, O.; Deng, K. K.; Kim, N. J.; Ross, L.; Surampalli, R. Y.; Hu, Z. The inhibitory effects of silver nanoparticles, silver ions and silver chloride colloids on microbial growth. Water Res. 2008, 42, 3066–3074.

42. Wigginton, N. S.; De Titta, A.; Piccapietra, F.; Dobias, J.; Neasatyy, V. J.; Suter, M. J. F.; Bernier-Latmani, R. Binding of silver nanoparticles to bacterial proteins depends on surface modifications and inhibits enzymatic activity. Environ. Sci. Technol. 2010, 44, 2163–2168.

43. Hernandez-Sierra, J. F.; Ruiz, F.; Cruz Pena, D. C.; MartinezGutierrez, F.; Martinez, A. E.; Guillen, A. P.; Tapia-Perez, H.; Castanon, G. M. The antimicrobial sensitivity of Streptococcus mutansto nanoparticles of silver, zinc oxide, and gold. Nanomed. Nanotechnol. 2008, 4, 237– 240.

44. Ahamed, M.; Karns, M.; Goodson, M.; Rowe, J.; Hussain, S. M.; Schlager, J. J.; Hong, Y. DNA damage response to different surface chemistry of silver nanoparticles in mammalian cells. Toxicol. Appl. Pharmacol. 2008, 233, 404–410.

45. Percival, S. L., Bowler, P. G., Dolman, J. Antimicrobial activity of silver-containing dressings on wound microorganisms using an in vitro biofilm model. Int. Wound. J. 2007, 4, 186-191.

46. Gitipour, A.; ElBadawy, A.; Arambewela, M.; Miller, B.; Scheckel, K.; Elk, M.; Ryu, H.; Gomez, V. A.; Santo Domingo, J.; Thiel, S.; Tolaymat, T. The impact of silver nanoparticles on the composting of municipal solid waste. Environ. Sci. Technol. 2013, 47, 14385-14393.

47. Gitipour, A.; Thiel, S. W.; Scheckel, K. G.; Tolaymat, T. Anaerobic toxicity of cationic silver nanoparticles. Sci. Total. Environ. 2016. 557-558, 363-368.

AgNPs Product Distribution

Automotive Paints/Inks Pet Products Other Household Products Electronics Baby Products Kitchen Supplies Food Supplements Medicinal Products Air and Water Filters Cleaning & Antimicrobial Products Textiles Personal Care Products

0 20 40 60 80 100 120 140 160

NUMBER OF PRODUCTS

Figure 1.1 Silver Nanoparticle Product Distribution

Chapter 2

The Impact of Silver Nanoparticles on the Composting of Municipal Solid Waste

This chapter has been published as

Gitipour, A.; ElBadawy, A.; Arambewela, M.; Miller, B.; Scheckel, K.; Elk, M.; Ryu, H.; Gomez, V. A.;

Santo Domingo, J.; Thiel, S.; Tolaymat, T. The impact of silver nanoparticles on the composting of municipal solid waste. Environ. Sci. Technol. 2013, 47, 14385-14393.

Abstract

This study evaluates the impact of polyvinylpyrrolidone (PVP) coated silver nanoparticles (PVP-

AgNPs) on the composting of municipal solid waste. The results suggest that there was no statistically significant difference in the leachate, gas and solid quality parameters, and overall composting performance between the treatments containing the AgNPs, Ag+ and the negative control. Nonetheless, taxonomical analyses of 25 Illumina 16S rDNA barcoded libraries containing 2,393,504 sequences indicated that the composted samples were highly diverse and primarily dominated by Clostridia (48.5%), Bacilli (27.9%), and beta-Proteobacteria (13.4%).

Bacterial diversity studies showed that the overall compost bacterial community structure changed in response to Ag-based treatments. However, the data suggest that functional performance was not significantly affected due to potential bacterial functional redundancy within the compost samples. The data also indicate that while the surface transformation of AgNPs to AgCl and Ag2S can reduce the toxicity, complexation with organic matter may also plays a major role. The results of this study further suggest that at relatively low concentrations, the organically rich waste management systems would be able to withstand the presence of AgNPs.

2.1 Introduction

Silver has been recognized for its antimicrobial properties for over 2000 years (1). Recently, silver was engineered into nanoparticles, structures having at least one dimension ranging from 1 to 100 nm in size, with unique physicochemical properties that differentiate them from their bulk counterparts. The relatively high surface area to volume ratio of nanoparticles leads to more reactivity and sometimes-higher toxicity than the bulk material (2). Silver nanoparticles (AgNPs) are receiving much attention and are currently employed in a broad range of applications (2). The strong antibacterial properties of AgNPs have encouraged their use as antibacterial/antifungal agents in various types of consumer products (3). The increasing rate of applications has caused

AgNPs to become one of the most widely commercialized nanomaterials today (2).

While a number of studies attempted to explain the mechanisms by which AgNPs exert their antimicrobial activity, the proposed mechanisms are still not fully understood. Proposed mechanisms include a combination of a release of silver ions by AgNP dissolution under aerobic conditions and specific AgNP properties such as nanoparticle transport by a Trojan-horse type mechanism followed by generation of reactive oxygen species (ROS) leading to cell membrane damage (4-7). Several studies have demonstrated AgNPs to be toxic in both aerobic and anaerobic environments (8, 9).

Like many other nanoparticles, AgNPs have the potential for release into the environment throughout their life cycle. At the end of their useful life, products containing nanoparticles are often disposed of with the municipal solid waste stream. In a disposal scenario, nanoparticles (e.g.,

AgNPs) may leach from products into the solid waste. Research by Benn and Westerhoff suggested that AgNPs could be released from AgNP-coated socks. Nanoparticles could also enter the solid waste stream with biosolids from wastewater treatment plants (10). While landfilling is still the

22 dominant method for solid waste management in the United States (U.S.), (11) composting is gaining momentum. Of the approximately 250 million tons of solid waste generated in the U.S in

2010, the recycled and composted fraction of this waste exceeded 85 million tons

(11). Composting is recognized as a low cost, environmentally sound process and has been recommended as an alternative to landfilling of food and green wastes (12).

There are concerns associated with the growing use of AgNPs and their potential impact on waste management systems that largely depend on microbial decomposition for waste stabilization.

Modeling results indicate that up to 15% of the total silver in the forms of Ag+ and/or AgNPs could be released from biocidal plastics and textiles (13, 14). Other studies have shown between 34% and 80% of the Ag released from commercially available functional (nano) textiles are in the form of AgNPs (nanocomposites, elemental AgNPs, and nano-AgCl particles) (15). The properties of the released AgNPs (e.g., stabilization mechanism and the chemistry of the capping agent) and the surrounding environmental conditions (e.g., the pH, ionic strength, electrolyte type, and the presence of natural organic matter (NOM)) govern their environmental fate and toxicity (16,

17). Ho et al. reported that aggregated nanoparticles posed lower toxicity relative to stable nanoparticles (18). In addition to aggregation, AgNPs may undergo surface transformations that will influence their behavior. For example, in a system that has high concentrations of chlorides and sulfides (e.g., wastewater, composter leachate, and landfill leachate), the released AgNPs may transform to silver chlorides (AgCl) and silver sulfides (Ag2S), which are less toxic relative to the metallic AgNPs (19, 20).

The current study aims at investigating the impacts of PVP-AgNPs on the composting process of the biodegradable organic fraction of municipal solid waste. This research represents one of the few studies that evaluate end-of-life management concerns with regard to the increasing use of

23 nanomaterials in everyday life. PVP-AgNPs are used since sterically stabilized NPs showed minimal aggregation in high ionic strength solutions with different background electrolyte valences (up to 1 M Na+ and Ca2+) (21). The conductivity of compost leachate has been reported to range between 750 and 9000 uS cm–1, (22) which translates to approximately 0.01–0.2 M ionic strength. Therefore, PVP-AgNPs are not expected to aggregate under the experimental ionic strength.

The study also examines relatively low concentrations that may be encountered in a real world scenario and how such exposure may impact the function and composition of microbial communities associated with compost samples. More recently, studies measuring the impact of

AgNPs on mixed microbial communities have used molecular profiling techniques such as T-

RFLP (23) and DGGE analysis of 16S-rRNA gene sequences (23, 24). For example, Colman et al.(23) showed that AgNPs can have an effect on the microbial composition of sediments as determined by T-RFLP. It should be noted that these community profiling methods can only provide an idea of how treatments can impact the most abundant populations and therefore generate an incomplete picture for the rest of the bacterial community members. Moreover, such profiling methods do not identify the members of the community. Alternatively, next generation sequencing methods utilized in the current study can identify a large number of the bacterial populations within samples exposed to AgNPs, including relatively low abundant members, and provide relevant information on the population dynamics within compost samples as a result of exposure to AgNP treatments.

2.2 Experimental Section

2.2.1 Nanoparticles, Selection, Synthesis, and Purification

The toxicity of nanomaterials including AgNPs is impacted by the size of the nanoparticles, and thus, it is critical to utilize a nanoparticle that could resist size change resulting from aggregation.

Aggregation may also lead to the nanoparticles settling out of solution and as a result a lower level of interaction between the AgNPs and the microorganisms in the compost media, which may further confound the results of the experiment and increase the uncertainty and accuracy of the conclusions on the toxicological impacts of nanomaterials. Therefore, the highly stable PVP-

AgNPs were utilized in the current study. It has been reported that PVP-AgNPs resist aggregation in high ionic strength solutions with high valence background electrolytes (25) which are typical conditions in a compost system. Additionally, research showed that the most commonly used capping agents for silver nanoparticles are citrate and polyvinylpyrrolidone (PVP) (25). This makes PVP-AgNPs a suitable nanoparticle to be utilized for investigating the impacts of AgNPs in complex environmental settings. The polyvinylpyrrolidone (PVP) coated AgNPs (PVP-AgNPs) were prepared and purified according to a method described previously by El Badawy et al.

(26) (more details are presented in Supporting Information (SI)). The hydrodynamic diameter

(HDD) and zeta potential of the AgNPs were measured using a Zetasizer Nanoseries (Malvern

Instruments). Transmission electron microscopy (TEM) was used to verify nanoparticles’ characteristics (size and shape). TEM samples were prepared by depositing a drop of the sample suspension on a carbon coated copper grid. Samples were air-dried at room temperature overnight in a dust-free box. Images were captured using a JEOL-1200 EX TEM (JEOL Inc.) operated at

120 kV. Total Ag concentrations were measured using a PerkinElmer AAnalyst 800 atomic absorption spectrometer after microwave acid digestion following EPA method 3015A.

2.2.2 The Compost Reactors

Nine composters (Figure 2.1) each with a volume of 130 L were used in the study. To maintain the uniformity of food waste composition, a synthetic food waste mixture (Figure A1, Appendix

A) was prepared in a manner to obtain an initial carbon-to-nitrogen (C/N) ratio of 34 ± 3, which is in the range 18–40 considered ideal for the composting process, and as composting proceeds, C/N gradually decreases until reaching levels below 12, indicating compost maturity (27). Each composter received 27 kg size reduced waste because solid waste systems are rather heterogeneous and to minimize the heterogeneity in the waste composition among the compost reactors, the waste components were size reduced by cutting each individual component into approximately 2 cm × 2 cm × 2 cm pieces. The nine reactors were divided into three batches (three reactors each); the first

+ was used as a negative control (no treatment); the second was treated with Ag (using AgNO3 salt), and the third was treated with PVP-AgNPs. The total silver concentration in both treatments was

2 mg kg–1 of compost. The composters were operated at a temperature of 50 °C (optimum for the composting process) and an initial moisture content of 65% (w/w) to be in the average range reported for compost moisture content (28). Warm humidified air was added to each reactor at a rate of 2.5 L min–1 for the duration of the study. The composters were mixed daily and monitored for 60 days, during which most of the compost activity occurs (29).

2.2.3 Composting Sampling and Analysis

To evaluate the performance of the composters, gas, leachate, and solid samples were monitored.

Gas composition was evaluated at 8-h intervals for the initial two weeks and 12-h intervals for the remainder of the experiment. The gas samples were collected using a 50 mL airtight syringe and immediately analyzed for the concentration of CO2, O2, CH4, H2, and N2O using a gas chromatograph (GC, Shimadzu) equipped with a thermal conductivity detector (TCD) and a flame

26 ionization detector (FID). Prior to gas sampling, air introduction was stopped for 15 min to allow headspace gas to equilibrate.

Leachate samples were collected twice per week and analyzed for pH, conductivity, chemical oxygen demand (COD; HACH, DR890 Colorimeter), total organic carbon (SHIMADZU, TOC-

V, EPA Method 9060A), ammonia-nitrogen (HACH, DR890 Colorimeter), and chloride

(DIONEX ICS-2000 Ion Chromatograph). The leachate samples were also analyzed for total silver content (GF-AA) after digestion using EPA method 3015.

Solid samples were analyzed for carbon to nitrogen ratio (C/N) using a Carlo Erba NC2500 elemental analyzer (more details are presented in the SI). The surface transformations of AgNPs were determined using X-ray Photoelectron Spectroscopy (XPS; Surface Science Laboratories

SSX-100 XPS) and X-ray absorption spectroscopy (XAS; details available in the SI). The total silver concentrations of the solid samples were determined using (GF-AA) after acid digestion method 3051. The solid sample collection was performed after rotating and lifting the compost reactors in order to obtain representative samples. The leachate and gas data were analyzed using an ANOVA test (SigmaPlot 12.0, Systat Software Inc.) where statistical significance was calculated using untreated samples (control) and treated samples (AgNO3 and AgNPs), and those with a p value < 0.05 are considered significant.

2.2.4 DNA Sequencing and Bacterial Composition and Diversity Analyses

Approximately, 0.25 g (wet weight) of each compost sample was used to extract total DNA using the UltraClean Soil DNA kit following the manufacturer’s instructions (MoBio Laboratories Inc.,

Solana Beach, CA). An aliquot of the DNA extracts (2 μL) was used to partially amplify the 16S rRNA gene using barcoded primers 515F and 806R (30), which allows the generation of overlapping forward and reverse reads for each sequence. The amplification step was performed

27 using a BioRad Tetrad and using the following thermal cycling conditions: (1) 95 °C for 3 min;

(2) 35 cycles of 45 s at 95 °C, 60s at 50 °C, and 72C; (3) 72 °C for 10 min. Illumina MiSeq coupled with pair end 250 bp kits were used to generate sequencing data. Prior to analyses, primer sequences were removed resulting in average sequence reads of 251 bp. The sequences were processed and analyzed using the software MOTHUR v1.30.1 (31). Briefly, fastq files for forward and reverse reads were used to form contigs which were first screened for sequence length (no greater than 255 bp). Sequences with ambiguous bases (N’s), containing homopolymers greater than seven bases, and classified as chloroplasts and mitochondria were removed. To reduce computational time, unique sequences were used to identify chimeras and to generate distance matrix and cluster analyses. MOTHUR was used to align and sort sequences with >97% similarity into operational taxonomic units (OTUs). Chimera-free sequences were classified using the tool

Classifier in Ribosomal Database Project II release 10.28 (32).

Barcoded 16S rRNA gene libraries were developed for 25 out of the 27 compost samples available from weeks one, two, and four. After removing undesirable sequences (i.e., chimeras, and contigs containing ambiguous bases, homopolymers, or outside of the sequence length range), a total of

2 393 504 reads were used in downstream analyses. The total number of contigs containing unique sequences was 178 085. The latter data set was composed of 120 699 unique sequences (OTUs) at the 99% sequence identity, which were then used in diversity analyses to further reduce computational needs. The latter sequences were used to determine the bacterial composition of the samples and the impact of Ag-based treatment on community structure and bacterial diversity.

Sequences from rare members were eliminated from analyses that focused on the most abundant

OTUs.

2.2.5 Species Richness, Diversity, and Statistical Analysis of Microbial Communities

Prior to analysis, read libraries (i.e., samples) were normalized by randomization to the smallest data set (i.e., 35 913 reads). Species richness (S), Shannon–Wiener diversity (H), and evenness

(EH) were calculated for each sample according to methods described elsewhere (33).

Subsampling and diversity indices were calculated with MOTHUR (34). In within-sample diversity (i.e., Shannon–Wiener) and taxonomic distribution (e.g., % of phylum Clostridia) for each sampling time were calculated using the coefficient of variation (CV). The CV is described in Shade et al. and is defined by the equation CV = σ/μ, where σ is the standard deviation and μ is the mean (35). Nonmetric multidimensional scaling (nMDS) analysis based on the Bray–Curtis similarity coefficient of the transformed data (log[x + 1]) was used to describe the relationships among microbial communities based on the relative distribution of OTU groups. A two-way crossed analysis of similarities (ANOSIM) was used to identify with significant differences (p <

0.05) the community assemblages among treatments and between sampling times. R values near

0 indicate no difference between groups, whereas those greater than 0 indicate dissimilarities between groups (36). Similarity Percentage (SIMPER) analysis was used to determine which species were most responsible for the differences observed between communities of different treatments and sampling times (34). All statistical analyses were performed with the software

PAST v2.17c (37).

2.3 Results and Discussion

2.3.1 AgNPs Characterization

The PVP-AgNPs investigated, suspension pH of 7.0, had a zeta potential (ζ) of −10.2 mV and a

HDD of 12.3 ± 0.5 nm. The weight of evidence from the literature suggests that engineered nanoparticles are likely to be of concern owing to their unique properties when they have diameters

29 of 30 nm or less (31, 38). The AgNPs average size obtained from TEM was 25.8 (Figure 2.2), which is different than the average size obtained by the DLS (12.3 nm). It has been reported in the literature that the TEM size of nanoparticles may be different from the DLS size for various reasons, including the potential increase in size as a result of the drying process during the preparation of the samples for TEM analysis (25). As previously mentioned, the PVP stabilized

AgNPs were used in this experiment because they are known to resist aggregation under various environmentally relevant conditions (25). Kvitek et al. determined that PVP is the most effective polymer for stabilizing AgNPs, (39) and thus a good candidate for investigating the toxicological impacts of AgNPs under complex matrices such as composting. Furthermore, purification of the suspension ensured that any observed effects were associated with nanoparticles and not impurities carried over from the synthesis process.

2.3.2 Gas Quantity and Quality

Microbial respiration is a commonly used indicator when evaluating the extent of biological activity present in the composting process (40, 41). Therefore, potential impacts of AgNPs on composting could be indirectly evaluated using gaseous emissions. Gas samples were analyzed for

CO2, O2, CH4, H2, and N2O concentration. Regardless of the treatment, CH4 and H2 concentrations were below the method detection limits (0.3% for CH4 and 0.04% for H2). As expected, the average

O2 concentrations (Figure 2.3a) correlated with CO2 concentrations (Figure 2.3b) as a result of aerobic microbial respiration. No statistically significant differences were observed for the average daily levels of CO2 and NO2 (Figure 2.3b and c; p > 0.05; the averages with the standard deviations of the gas measurements are presented in Figure A2, Appendix A). As will be discussed later in the manuscript, most of the silver in the current study is bound to the organic matter present and may have caused a reduction in the potential toxicity of silver.

2.3.3 Leachate Quality

The pH of the compost leachate ranged between 6.7 and 8.8, which is well within the range commonly reported for the process (42) (Figure A3, Appendix A). The conductivity values (10–

60 uS cm–1) for the compost leachate (Figure A4, Appendix A) were a result of the high concentration of the water-soluble salts within the food waste (43). The TOC (4000–25000 mg·L–

1) and COD levels (25 000–100 000 mg·L–1), presented in Figures A5 and A6 (Appendix A), indicate that the compost materials were initially very high in organic matter, and as the composting process proceeded, the organic matter was mineralized by microbial activity, resulting in lower concentrations. In all treatments, the ammonia-nitrogen level peaked between days 8 and

12 and, by day 25, had decreased to relatively stable levels (Figure A7, Appendix A). Overall, no statistically significant differences (p > 0.05) were observed between the various treatments. Low concentrations of total silver (Figure A8, Appendix A) were detected in the leachate of the composters treated with silver nitrate and PVP-AgNPs, which may indicate a precipitation of silver as silver chlorides and adsorption to the waste materials. This is supported by the high concentrations of chlorides detected in the leachate samples as presented in (Figure A9, Appendix

A). In order to verify the transformations of AgNO3 and AgNPs to AgCl and other species, XPS and XAS analyses were performed on leachate samples. Silver was not detected, which can be attributed to the low concentration of silver present in the leachate (Table A1, Appendix A). In order to overcome the detection problem, leachate samples were collected from the positive control

(Ag+) and from the composters treated with PVP-AgNPs after one month. These samples were then spiked with 200 mg L–1 of Ag+ and PVP-AgNPs, respectively. The spiked samples were analyzed using the XAS technique after one week of preparation. In the case of the positive control samples (Ag+), a strong silver signal was obtained, and the results confirmed the transformation of

31 silver ions to AgCl, Ag cystine, and Ag humic phases (Figure 2.4 and Table A2; more detail about the identification of these species is presented in Appendix A). On the other hand, a weak signal was obtained from the leachate samples spiked with the PVP-AgNPs (data not shown). The weak signal may have resulted from the attachment of the NPs on the solid organic fractions present in the leachate and the formation of bigger aggregates that subsequently settle, along with the attached, AgNPs out of solution and as a result were not detected. Nonetheless, under the current experimental conditions, phase transformations can occur and significantly affect the stability, bioavailability, and toxicity of Ag-NPs (19). Also, the Cl– ions present in the media react with the dissolved silver ions and form AgCl as confirmed by XAS analysis. Therefore, the surface transformation of AgNPs as well as Ag+ to AgCl may further explain the results observed in the current study. Choi et al. demonstrated that the toxicity of AgNPs to nitrifying bacteria was reduced in the presence of various anions such as chloride and sulfate (5). It is noted that previous studies have suggested that one of the toxicity mechanisms of AgNPs is through the dissolution to

Ag+ (7). With the presence of a relatively high concentration of Cl– in solution, most of the ionic silver transforms to AgCl (19).

2.3.4 Solids Quality

One of the key indicators of compost decomposition is the carbon to nitrogen (C/N) ratio. A decreasing trend in the ratio of C/N, followed by the eventual stabilization of the organics, can generally be observed as composting proceeds. This is caused by the loss of carbon from the system as a result of the decomposition of organic substrates and the release of CO2 (43, 44). The

C/N ratio of the compost materials used in the current study is presented in (Figure A10, Appendix

A). Initially, the C/N of the utilized composting materials was 34, and as the experiment progressed, the C/N ratio gradually decreased to 10 by the end of the study, indicating compost

32 maturity. The relatively high initial value could be a result of the higher proportion of waste high in carbon (e.g., paper and cardboard) to waste high in nitrogen (e.g., grass; Figure A1, Appendix

A). No significant differences were found for C/N among the treatments (p > 0.05). The C/N ratio obtained herein was in agreement with ratios and trends reported in previous studies (45).

The concentration of total silver in the final composted solid samples ranged between 3 and 6 mg kg–1 (Figure A11, Appendix A), suggesting that a large fraction of the silver was adsorbed onto or complexed with the natural organic matter of the solid media (46). Even with the increase of silver in the solid phase, the silver concentrations in the samples were below the detection limit of the

XPS as presented in Table A1 (Appendix A). XAS analysis was also used to determine possible surface transformation of the silver nanoparticles in the compost solid samples. At time zero (the first possible time a sample was collected once the NPs were mixed into the system), the XAS spectra showed that the PVP-AgNPs attached to the compost material was mostly in the metallic form (Figure 2.4 and Table A2 (Appendix A)). After one month, the silver was not detected using

XAS analysis. It is noted that heterogeneity is rather high in the 27 kg compost reactor utilized in the current study. Even with the mixing, variability in AgNPs content in the collected samples is inevitable. With silver content around the detection limit for the XAS, it is not surprising that silver was not detected.

Therefore, neither the XPS nor the XAS analyses confirm the surface transformations of AgNPs to AgCl in the solid compost samples. Nevertheless, this does not eliminate the possibility of this transformation occurring especially with the high amounts of chlorides present in the system.

Gunawan et al. showed the influence of the oxidation state of Ag on its antimicrobial action (47).

AgNPs are not thermodynamically stable under most environmental conditions and will oxidize or react with organic ligands (48). As an example, silver is known to react strongly with sulfide,

33 chloride, and organic matter in relevant environmental situations (19). Another possible explanation of the lack of overall impact of AgNPs on the composting process is the complexation with the compost products containing high amounts of natural organic matter (NOM), which can limit the AgNPs bioavailability (46).

2.3.5 Bacterial Diversity Analyses

Differences in community structure profiles were observed between control samples and treated samples, although differences were not statistically significant between treatments (Table 2.1 and

Table A3, Figure A13, Figure A14, Appendix A). Moreover, statistically significant differences in community structure were observed over time, suggesting that while Ag-based treatments can impact the microbial community structure, time is also an important factor in the bacterial population dynamics during the composting process (Table A4, Appendix A). These results are confounded by the relatively large variation in diversity between samples from the same type

(Figure 2.5). Differences in the coefficient of variation were more notable during the first week.

However, the variation decreased over the next few weeks, further supporting the importance of time on the changes in bacterial community structure (i.e., regardless of treatment). The variation could also depend on the level of dissolved ions as well as direct contact of NPs with microbial cells.(49) For example, Xiu et al. proposed that under anaerobic conditions environmental parameters that stimulate Ag+ ion release play a larger role that particle size and physical contact

(50).

In contrast with community sequencing profiles, richness, evenness, and Shannon diversity indices were not significantly different between treatments and over time (Figure 2.6, Figures A15 and

A16, Appendix A) suggesting that, while the treatments were perturbing the microbial composition, they were not catastrophic. The data also suggest that highly diverse microbial

34 communities can help to deal with perturbations associated with AgNP exposure, perhaps due to the presence of functional redundant bacterial groups. Relatively low impacts in microbial composition have been observed for other environmental matrices after AgNP exposure. For example, Sun et al. found that AgNPs did not change the community structure of unsettled activated sludge after 24 h treatments (24). Using pyrosequencing, Doolette et al. showed that the dominant populations in wastewater reactors were not significantly impacted by the addition of low doses of AgNPs (51). Microbial processes such as nitrification and methanogenesis were not impacted either.

The presence of 120 699 OTUs suggests that these compost samples are very diverse. Even when rare members were removed (sequences present only once in all the samples, Figure

A12, Appendix A), more than 9368 OTUs were identified. To the authors’ knowledge, this is the highest number of OTUs reported for compost samples. The most represented bacterial taxa at the class level were Clostridia (48.5%), Bacilli (27.9%), and beta-Proteobacteria (13.4%) (Figure 2.5).

Other classes represented were gamma- and alpha-Proteobacteria and Erysipelotrichia, but to a much lesser extent (<1%). The presence of members of the Bacilli and beta-Proteobacteria is in agreement with the predominant aerobic conditions within the reactors.

While aerobic conditions were purposely maintained, several populations that are commonly associated with anaerobic respiration were found. Most of the anaerobic bacteria identified were classified as closely related to Clostridia, a finding that has been previously noted in compost samples (52). While it is possible that a portion of the Clostridia is associated with spore-forming groups, some of these populations may be aerotolerant. However, the presence of anaerobic pockets in the composted material is also possible particularly within the bacterial fraction associated with biofilms.

2.4 Environmental Implications

The AgNPs evaluated in the study did not significantly influence aerobic composting processes at the concentrations that could be expected to be present in the solid waste stream. AgNPs’ toxicity is dose-dependent and has been observed at higher concentrations in both aerobic and anaerobic environments (54, 55). While the surface transformation of AgNPs to AgCl and Ag2S can reduce the toxicity, complexation with organic matter may also play a role (19). The taxonomical analysis results presented in this study are in agreement with Das et al., who found that microbial activity can be impacted within hours of AgNP exposure, although at low AgNP concentrations activity can be relatively stable (53). It is unclear, however, whether the bacterial response is caused by bacterial populations resistant to AgNPs, the decrease in Ag+ release, transformation of AgNPs, or reduction in bioavailability. It should be noted that recovering bacterial phylotypes could be responsible for the differences in community structure between treated and untreated environmental water samples (53). However, there is little information on AgNPs impacts on the microbial community for composted samples. Altogether, data from this and other studies show that toxicity dynamics might be different when dealing with complex microbial communities associated with environmental matrices (53), which suggests that data from pure culture studies may be inaccurate in predicting the impact of AgNPs, and nanoparticles in general, on natural and/or engineered microbial communities.

Extrapolating from the results presented herein, similar toxicological behavior of AgNPs would be expected in the organically rich municipal solid waste landfills if the concentration of AgNPs was relatively low. Additional research, however, is still needed to identify at which concentrations

AgNPs start to have toxicological impact on waste management systems where microbial groups and microbial processes may be more impacted by AgNPs.

Literature Cited

1. Khaydarov, R.R.; Khaydarov, R. A.; Estrin, Y.; Evgrafova, S.; Scheper, T.; Endres, C.; Cho, S. Y. Silver Nanoparticles Environmental and Human Health Impacts. Nanomaterials: Risks and Benefits. Springer. 2009, pages 287-297.

2. Maynard, A.D., Michelson, E., 2006. The Nanotechnology Consumer Product Inventory; http://www.nanotechproject.org/44.

3. Tolaymat, T. M.; El Badawy, A. M.; Genaidy, A.; Scheckel, K. G.; Luxton, T. P.; Suidan, M. An evidence-based environmental perspective of manufactured silver nanoparticle in syntheses and applications: A systematic review and appraisal of peer-reviewed scientific papers. Sci. Total Environ. 2010, 408, 999-1006.

4. Marambio-Jones, C.; Hoek, E. M. V. A review of the antibacterial effects of silver nanomaterials and potential implications for human health and the environment. J. Nanopart. Res. 2010, 12, 1531-1551.

5. Choi, O.; Deng, K. K.; Kim, N. J.; Ross, L.; Surampalli, R. Y.; Hu, Z. The inhibitory effects of silver nanoparticles, silver ions and silver chloride colloids on microbial growth. Water Res. 2008, 42, 3066-3074.

6. Xu, H.; Qu, F.; Xu, H.; Lai, W.; Wang, Y. A.; Aguilar, Z. P.; Wei, H. Role of reactive oxygen species in the antibacterial mechanism of silver nanoparticles on Escherichia coli O157:H7. Biometals 2012, 25, 45-53.

7. Sotiriou, G. A.; Pratsinis, S. E. Antibacterial activity of nanosilver ions and particles. Environ. Sci. Technol. 2010, 44, 5649-5654.

8. Choi, O. K.; Hu, Z, Q. Nitrification inhibition by silver nanoparticles. Water Sci.Technol. 2009, 59, 1699-1702.

9. Navarro, E.; Piccapietra, F.; Wagner, B.; Marconi, F.; Kaegi, R.; Odzak, N.; Sigg, L.; Behra, R. Toxicity of silver nanoparticles to Chlamydomonas reinhardtii. Environ. Sci. Technol. 2008, 42, 8959-8964.

10. Benn, T. M.; Westerhoff, P. Nanoparticle silver released into water from commercially available sock fabrics. Environ. Sci. Technol. 2008, 42, 4133-4139.

11. Barlaz, M. A.; Kaplan, P. O.; Ranjithan, S. J.; Robert, R. Comparing recycling, composting and landfills. BioCycle 2003, 44, 9, 60.

12. Lou, X. F.; Nair, J. The impact of landfilling and composting on greenhouse gas emissions - A review. Bioresource Technol. 2009, 100, 3792-3798.

13. Blaser, S. A.; Scheringer, M.; MacLeod, M.; Hungerbuhler, K. Estimation of cumulative aquatic exposure and risk due to silver: contribution of nano-functionalized plastics and textiles. Sci. Total. Environ. 2008, 390, 396-409.

14. Geranio, L.; Heuberger, M.; Nowack, B. The behavior of silver nanotextiles during washing. Environ. Sci. Technol. 2009, 43, 8113-8118.

15. Lorenz, C.; Windler, L.; Goetz, N. V.; Lehmann, R. P.; Schuppler, M.; Hungerbuhler, K.; Heuberger, M.; Nowack, B. Characterization of silver release from commercially available functional (nano)textiles. Chemosphere. 2012, 89, 817-824.

16. Nel, A.; Xia, T.; Madler, L.; Li, N. Toxic potential of materials at the nanolevel. Science 2006, 311, 622-627.

17. Fabrega, J.; Fawcett, S. R.; Renshaw, J. C.; Lead, J. R. Silver nanoparticle impact on bacterial growth: effect of pH, concentration, and organic matter. Environ. Sci. Technol. 2012, 43, 7285- 7290.

18. Ho, C. M.; Yau, S. K. W.; Lok, C. N.; So, M. H.; Che, C. M. Oxidative dissolution of silver nanoparticles by biologically relevant oxidants: A kinetic and mechanistic study. Chem. Asian J. 2010, 5, 285-293.

19. Levard, C.; Hotze, E. E.; Lowry, G. V.; Brown Jr, G. L. Environmental transformations of silver nanoparticles: impact on stability and toxicity. Environ. Sci. Technol. 2012, 46, 6900- 6914.

20. Kim, B.; Park, C.S.; Murayama, M.; Hochella, M.F. Discovery and characterization of silver sulfide nanoparticles in final sewage sludge products. Environ. Sci. Technol. 2010, 44, 7509- 7514.

21. El Badawy, A. M.; Luxton, T. P.; Silva, R. G.; Scheckel, K. G.; Suidan, M. T.; Tolaymat, T. M. Impact of environmental conditions (pH, ionic strength and electrolyte type) on the surface charge and aggregation of silver nanoparticles suspension. Environ. Sci. Technol. 2010, 44, 1260-1266.

22. Dimambro, M. E.; Lillywhite, R. D.; Rahn, C. R. The physical, chemical and microbial characteristics of biodegradable municipal waste derived composts. Compost Science & Utilization. 2007, 15, 243-252.

23. Colman, B.P.; Wang, S.Y.; Auffan, M.; Wiesner, M. R.; Bernhardt, E. S. Antimicrobial effects of commercial silver nanoparticles are attenuated in natural streamwater and sediment. Ecotoxicology. 2012, 21(7), 1867-77.

24. Sun. X.; Sheng, Z.; Liu, Y. Effects of silver nanoparticles on microbial community structure in activated sludge. Sci Total Environ. 2013, 443, 828-35.

25. El Badawy, A. M.; Luxton, T. P.; Silva, R. G.; Scheckel, K. G.; Suidan, M. T.; Tolaymat, T. M. Impact of environmental conditions (pH, ionic strength and electrolyte type) on the surface charge and aggregation of silver nanoparticles suspension. Environ. Sci. Technol. 2010, 44, 1260– 1266.

26. El Badawy, A. M.; Silva, R. G.; Morris, B.; Scheckel, K. G.; Suidan, M. T.; Tolaymat, T. M. Surface charge-dependent toxicity of silver nanoparticles. Environ. Sci. Technol. 2011, 45, 283-287.

27. Kumar, Sunil. Composting of Municipal Solid Waste Crit. Rev. Biotechnol. 2011, 31, 112– 136.

28. Smårs, S.; Beck-Friis, B.; Jönsson, H.; Kirchmann, H.SE—structures and environment: an advanced experimental composting reactor for systematic simulation studies. J. Agric. Eng. Res. 2001, 78, 415–422.

29. Goyal, S.; Dhull, S. K.; Kapoor, K. K. Chemical and biological changes during composting of different organic wastes and assessment of compost maturity. Bio Tech. 2005, 1584–1591.

30. Shade, A.; Caporaso, J. G.; Handelsman, J.; Knight, R.; Fierer, N. A meta-analysis of changes in bacterial and archaeal communities with time. ISME J. 2013, 7, 1493-4506.

31. Kozich, J.J.; Westcott, S.L.; Baxter, N.T.; Highlander, S.K.; Schloss, P.D. Development of a Dual-Index Sequencing Strategy and Curation Pipeline for Analyzing Amplicon Sequence Data on the MiSeq Illumina Sequencing Platform. Appl Environ Microbiol. 2013, 79, 5112- 5120.

32. Cole J.R.; Wang Q.; Cardenas E.; Fish J.; Chai B.; Farris R.J.; Kulam-Syed-Mohideen A.S.; McGarrell D.M.; Marsh T.; Garrity G.M.; Tiedje J.M. The Ribosomal Database Project: improved alignments and new tools for rRNA analysis. Nucleic Acids Res. 2009, 37, D141.

33. Hill T.C.J.; Walsh K.A.; Harris J.A.; Moffett B.F. Using ecological diversity measures with bacterial communities. FEMS Microbiol. Ecol. 2003, 43, 1.

34. Schloss, P. D.; Westcott, S. L.; Ryabin, T.; Hall, J. R.; Hartmann, M.; Hollister, E. B.; Lesniewski, R. A.; Oakley, B. B.; Parks, D. H.; Robinson, C. J.; Sahl, J. W.; Stres, B.; Thallinger, G. G.; Van Horn, D. J.; Weber, C. F. Introducing mothur: open-source, platform- independent, community-supported software for describing and comparing microbial communities. Appl. Environ. Microbiol. 2009, 75, 7537-7541.

35. Shade, A.; Caporaso, J. G.; Handelsman, J.; Knight, R.; Fierer, N. A meta-analysis of changes in bacterial and archaeal communities with time. ISME J. 2013, 7, 1493-4506.

36. Clarke, K. R. Non-parametric multivariate analyses of changes in community structure. Aust. J. Ecol. 1993, 18, 117-143.

37. Hammer, Ø.; Harper, D. A. T.; Ryan, P. D. PAST: paleontological statistics software package for evolution and data analysis. Palaeontol. Electron. 2001, 4, 1-9.

38. Auffan, M.; Rose, J.; Bottero, J.; Lowry, G.; Jolivet, J.; Wiesner, M. Towards a definition of nanoparticles based on novel size-dependent properties. Nat. Nanotechnol. 2009, 3, 634-641.

39. Kvitek, L.; Panacek, A.; Soukupova, J.; Kolar, M.; Vecerova, R.; Prucek, R.; Holecova, M.; Zboril, R. Effect of surfactants and polymers on stability and antibacterial activity of silver nanoparticles (NPs). J. Phys. Chem. C 2008, 112, 5825-5834.

40. Partanen, P.; Hultman, J.; Paulin, L.; Auvinen, P.; Romantschuk, M. Bacterial diversity at different stages of the composting process. BMC Microbiology 2010, 10, 94.

41. Adams, J. D. W.; Frostick, L. E. Analysis of bacterial activity, biomass and diversity during windrow composting. Waste Manage. 2009, 29, 598-605.

42. Wichuk, M. K.; McCartney, D. Compost stability and maturity evaluation-a literature review. Can. J. Civ. Eng. 2010, 37, 1505-1523.

43. He, X. T.; Logan, T. J.; Traina, S. J. Physical and chemical characteristics of selected U.S. municipal solid waste composts. J. Environ. Qual. 1995, 24, 543-552.

44. Khan, M. A. I.; Ueno, K.; Horimoto, S.; Komai, F.; Tanaka, K.; Ono, Y. Physiochemical, including spectroscopic, and biological analyses during composting of green tea waste and rice bran. Biol. Fert. Soils 2009, 45, 305-313.

45. Mkhabela, M. S.; Warman, P. R. The influence of municipal solid waste compost on yield, soil phosphorus availability and uptake by two vegetable crops grown in a Pugwash sandy loam soil in Nova Scotia. Agriculture, Ecosystems and Environment. 2005, 106, 57-67.

46. Delay, M.; Dolt, T.; Woellhaf, A.; Sembritzki, R.; Frimmel, F. H. Interactions and stability of silver nanoparticles in the aqueous phase: Influence of natural organic matter (NOM) and ionic strength. J. Chromatogr. A 2011, 1218, 4206-4212.

47. Gunawan, C.; Teoh, W. Y.; Marquis, C. P. Lifia, J.; Amal, R. Reversible antimicrobial photoswitching in nanosilver. Small 2009, 5, 341-344.

48. Xiu, Z-M.; Ma, J.; Alvarez, P. J. J. Differential effect of common ligands and molecular oxygen on antimicrobial activity of silver nanoparticles versus silver ions. Environ. Sci. Technol. 2011, 45, 9003-9008.

49. Bondarenko, O.; Ivask, A.; Käkinen, A.; Kurvet, I.; Kahru, A. Particle-cell contact enhances antibacterial activity of silver nanoparticles. PLOS One. 2013, e64060.

50. Xiu, Z.; Zhang, Q.; Puppala, H. L.; Colvin, V. L.; Alvarez, P. J. J. Negligible Particle-Specific Antibacterial Activity of Silver Nanoparticles. Nano Lett. 2012, 12, 4271– 4275.

51. Doolette, C. L.; McLaughlin, M. J.; Kirby, J. K; Batstone, D. J.; Harris, H. H; Ge, H.; Cornelis, G. Transformation of PVP coated silver nanoparticles in a simulated wastewater treatment process and the effect on microbial communities. Chem Cent J. 2013, 7(1), 46.

52. Yamada, T.; Suzuki, A.; Ueda, H.; Ueda, Y.; Miyauchi, K.; Endo, G. Successions of bacterial community in composting cow dung wastes with or without hyperthermophilic pre-treatment Appl. Microbiol. Biotechnol. 2008, 81, 771–781.

53. Das, P.; Williams, C. J.; Fulthorpe, R. R.; Hoque, M. E.; Metcalfe, C. D.; Xenopoulos. M. A. Changes in Bacterial Community Structure after Exposure to Silver Nanoparticles in Natural Waters. Environ. Sci.Technol. 2012, 46 (16), 9120-9128.

54. Choi, O.; Hu, Z. Size dependent and reactive oxygen species related nanosilver toxicity to nitrifying bacteria. Environ. Sci. Technol. 2008, 42, 4583-4588.

55. Yang, Y.; Gajaraj, S.; Wall, J.; Hu, Z. A Comparison of Nanosilver and Silver Ion Effects on Bioreactor Landfill Operations and Methanogenic Population Dynamics. Water Res. 2013, 47, 3422-3430.

Figure 2.1 Schematic of the experimental setup of the compost reactors.

Figure 2.2 TEM image of PVP-coated AgNPs.

Figure 2.3 Average gas levels in composters’ headspace (a) O2, (b) CO2, and (c) N2O.

Figure 2.4 The Ag Kα XAS spectra of AgCl, Ag Cystine, and Ag Humic as pure phases and PVP-

AgNPs reacted with the compost material and AgNO3 with the compost leachate.

Figure 2.5 The relative distribution of bacterial classes. Percentages are shown for control (C), Ag+ treated (A), and AgNP treated (B) samples.

Table 2.1 Results of Two-Way ANOSIM Test Based on Bray−Curtis Similarity Matrix Derived from the Distribution of Microbial Communities.

Global Source of variation Ra p Permutations Two-way crossed ANOSIM Global tests treatments (across all week groups) 0.481 0.001 999 sampling time (across all treatments groups) 0.338 0.003 999 Pairwise tests (treatments)b control vs treatment A 0.833 0.001 999 control vs treatment B 0.750 0.003 999 treatment A vs treatment B 0.033 0.410 999 Pairwise tests (sampling time)b 1 week vs 2 week 0.173 0.075 999 1 week vs 4 week 0.423 0.001 999 2 week vs 4 week 0.549 0.022 999 a R values greater than 0 (up to 1) indicate dissimilarities between groups, whereas those values near 0 indicate a true null hypothesis of no difference between groups. b Significance was set at α = 0.05.

Figure 2.6 Richness for OTUs at 99% sequence identity with rare members included (a, b) and after rare members were removed (c, d).

APPENDIX A

A1. Purification of AgNPs Suspensions

Upon synthesis, most AgNP suspensions contain unquantified Ag+ impurities as a result of the incomplete reduction of the silver salt precursors the AgNP. The suspension will also contain impurities from the synthesis process (reducing, stabilizing agents). These impurities may interfere when evaluating the toxicity of the AgNPs and therefore are critical to be removed. The purification of the AgNPs suspension was performed using an ultrafiltration device purchased from Spectrum Laboratories, Inc. The ultrafiltration membrane used was 10 kDa polyethersulfone

(PES) (MidiKros Hollw Fiber Module (P-X3-010E-300-02N)). Milli-Q water (conductivity of

0.67 uScm-1) was used to wash impurities from the AgNPs suspensions. The conductivity of the

AgNPs suspension was monitored as an indicator of the presence of impurities. Lower conductivity was translated into the presence of fewer impurities in the suspension. The AgNPs suspension’s conductivity was continuously monitored until reaching a conductivity of less than

10 uS cm-1 (Orion 013010MD conductivity cell). This purification technique has been proven to efficiently remove impurities from the synthesized silver suspension as described in more detail by El Badawy et.al. (1).

A2. C: N Ratio Measurement

A finely ground sample was weighed and placed in a tin capsule. Thereafter the capsules were loaded into the autosampler of the elemental analyzer to be combusted for C and N concentrations.

The combustion takes place at 1020 oC in the presence of oxygen, thus converting the sample to its elemental forms of CO2, N2 and H2O. Water was removed by a MgClO4 trap prior to the chromatographic column which separates the gases. The concentration is measured by a Thermal

Conductivity Detector (TCD).

A3. PCR Methodology

PCR amplification of archaeal and bacterial 16S RNA was performed using OneTaq 2X Master

Mix (New England Biolabs, MA) with the following cycling conditions: initial denaturation at

94oC for 1 min, 40 cycles of denaturation at 94oC for 30 sec, annealing at either 58oC or 59oC for

15 sec, and extension at 68oC for 45 sec, followed by final extension at 68oC for 5 min. Bacterial primers used were 984F and 1378R with a 40 base pair GC clamp. The archaeal primers used were

344F and 915R. The PCR products were analyzed by electrophoresis using 2% agarose E-Gels with SYBR Green (Invitrogen, CA). For the WAVE analysis, a flow rate of 0.9 mL min-1 and a denaturing temperature of 62 oC and 65oC for archaea and bacteria, respectively. Individual peaks were determined by setting a minimum peak height of 1mV and between 10 and 20 minutes of buffer flowthrough.

A4. XPS Analysis

Drops of aqueous suspension were placed on pieces of aluminum foil and allowed to dry for several hours under ambient conditions. The prepared samples were then analyzed using Surface Science

Labs SSX-100 XPS. The nominal x-ray beam diameter for the analysis was 600 um.

A5. XAS Analysis

Changes in the chemical speciation of silver in the compost reactors were assessed using X-ray absorption spectroscopy (XAS). The XAS analysis was conducted at the MRCAT beamline 10-

BM of Advanced Photon Source (APS), Argonne National Laboratory (ANL), Argonne, IL. The electron storage ring was operating at 7 GeV in top-up mode. A liquid N2 cooled double crystal

Si (111) monochromator was used to select incident photon energies and a platinum-coated mirror was used for harmonic rejection. The monochromator was calibrated by assigning the first

51 derivative inflection point of the absorption K-edge of Ag metal at 25514 eV and a Ag metal foil spectrum was collected congruently with each sample scan.

Data were collected in both transmission and fluorescence mode. The fluorescence measurements were conducted with an argon filled Lytle detector. The transmission data were utilized for presentation in the manuscript. It is noted that liquid samples were examined as liquids. The samples were transported, stored, and analyzed at 4 oC in a fluid cell designed for XAS studies.

The collected data were compared to a library of over two dozen silver standards. Most are ACS reagent grade chemicals (Ag2CO3, Ag2O, AgCl, AgS, AgO, AgCO3, AgAcetate, AgNO3, Ag2SO4, etc) and some are materials that were carefully prepared in the lab (Ag-Humic, Ag-Citrate, Ag-

Cystine, Ag-Cysteine, Ag sorbed to various iron oxides and clay minerals, Ag-Histidine, etc). This allows multiple lines of evidence to determine Ag-thiol and Ag-carboxyl bonds. For Agthiol, the sample spectra were compared to Ag-cystine, Ag-cysteine, Ag-histidine, and Ag2SO4. Ag-cystine and Ag-cysteine provide convincing evidence of Ag-thiol complexes based on pKa’s in the system even though carboxyl functional groups are present with these amino acids. For Ag-carboxyl complexes, we have evidence that Ag-Acetate, Ag-Humic, Ag-Citrate, and Ag-Histidine form suitable Ag-carboxyl complexes. The humic acid used is a synthetic material from Sigma Aldrich and is characterized to have primarily carboxyl functional groups with some amine groups. Our experimental protocol for preparing these standards is to mix a solution of 2000 ppm AgNO3 at a molar ratio of 1 AgNO3: 10 ligand and to examine the solution in a fluid cell via XAFS; the excessive ligand concentration and solution pH ensures complexation.

The IFEFFIT software package was employed for data processing; specifically, data were processed and analyzed with Athena (2). Principal component analysis (PCA) of the normalized sample spectra were used to estimate the likely number of species contained in the samples, whilst

52 target transformation (TT) was used to identify relevant standards for linear combination fitting

(LCF) of the sample spectra (3). PCA and TT were performed using SixPack (4) while data normalization and LCF were performed using Athena (2). Fitting range was -30 to +100 eV relative to the Ag K-edge. For each sample, the combination of standards with the lowest residual parameter was chosen as the most likely set of components.

9.5

9.0

8.5

8.0

7.5

7.0

6.5 Control Treated with Ag+ 6.0 Treated with AgNPs

5.5 0 10 20 30 40 50 60 70 Time (Day) Figure A.1 pH levels of composter leachate.

60 Control 50 Treated with Ag+ Treated with AgNPs

mS/cm

0 0 10 20 30 40 50 60 Time (Day) Figure A.2 Conductivity of composter leachate.

35000 Control 30000 Treated with Ag+ Treated with AgNPs 25000

20000

15000

10000

Total Organic Carbon (mg/L) Carbon Total Organic

5000

0 0 10 20 30 40 50 60 70 Time (Days) Figure A.3 Total Organic Carbon (TOC) concentration of composter leachate.

140000

Control 120000 Treated with Ag+ Treated with AgNPs 100000

80000

60000

40000

Chemical Oxygen Demand (mg/L) Demand Oxygen Chemical 20000

0 0 10 20 30 40 50 60 70 Time (Days) Figure A.4 Chemical Oxygen Demand (COD) concentration of composter leachate.

12000

Control 10000 Treated with Ag+ Treated with AgNPs 8000

6000

4000

2000

Ammonia Nitrogen Concentration (mg/L) Concentration Nitrogen Ammonia

0 0 10 20 30 40 50 60 Time (Day) Figure A.5 Ammonia-Nitrogen concentration of composter leachate.

Treated with Ag+ Treated with AgNPs 4

Ag Concentration (ppb) Concentration Ag

0 0 10 20 30 40 50 60 Time (Day) Figure A.6 Total silver concentration in composter leachate.

2500

Control + 2000 Treated with Ag Treated with AgNPs

1500

1000

Chloride Concentration (mg/L) Concentration Chloride 500

0 0 10 20 30 40 Time (Days) Figure A.7 Chloride concentration of composter leachate.

40 Control 35 Treated with Ag+ Treated with AgNPs 30

C:N ratio 15

0 0 10 20 30 40 50 60 Time (Days) Figure A.8 C:N ratio of composter solid samples.

20 Treated with Ag+ Treated with AgNPs

Ag Concentration (mg/kg) Concentration Ag 5

0 0 10 20 30 40 50 60 Time (Days) Figure A.9 Total silver concentration of composter solid samples.

Table A.1 XPS analysis results

Sample C O N Ag Si Na Cl K Ca P Treated with Ag+ (AgNO3) C=O C-O, C-N C-H N-C, N-H (30 days) 7.8 15.3 41 24.8 5 < 0.05 1.5 0.9 0.7 3.1

(60 days) 4 13.5 61.3 15.8 2.4 < 0.05 1.9 1

Treated with Ag-NPs (PVP)

(30 days) 6.4 15.8 61.5 11.3 2.1 < 0.05 0.4 0.6 2.1

(60 days) 7.4 16.4 60.2 12.4 1.3 < 0.05 0.9 0.7 0.8

Table A.2 XAS speciation results

Ag- Ag Ag Sample Ag Metal Ag Humic χ2 Thiol Sulfide Chloride Leachate spiked with -- 21 -- 15 64 0.0001 AgNO3 Compost spiked with 62 38 ------0.0161 PVP-AgNPs

0.24 B83 A98 0.18

B98 P98 C98 0.12

H98 0.06E98

2 D98 B822C822

a C83 G98 A822

d E822

o -0.40 -0.32 -0.24 -0.16 -0.08 D822F822 0.08 0.16 0.24 0.32

o G822

-0.06 P822 P83

A83 -0.12 H83 G83

-0.18 D83

-0.24 F83

E83 -0.30 Coordinate 1 Figure A.10 nMDS based on total community (120699 OTUs at 99% cutoff value), 1) A, B, C are the negative control composters, 2) D, E, F are the composters treated with Ag+, and 3) G, H, P are composters treated with AgNPs.

0.24

G83 C98 0.16 P83

B98A98 0.08 H83 E83

H98

2 F83 G98 E98 F822G822D822P98 D98

e t -0.24 -0.16 -0.08 0.08 A822 0.16 0.24 0.32

a P822

n B822

o E822 C822 o -0.08

A83 B83 C83 -0.16

-0.24

-0.32

D83 -0.40 Coordinate 1 Figure A.11 nMDS based on total community (9368 OTUs at 99% cutoff value with no RARE members), 1) A, B, C are the negative control composters, 2) D, E, F are the composters treated with Ag+, and 3) G, H, P are composters treated with AgNPs.

0.8 6.5 a Control b) Treatment A 0.6 ) Treatment B 5.5 0.4 4.5

Control 0.2 3.5 Treatment A

Treatment B Shannon div. (SE) Shannon div. Shannon div. (CV) Shannon div. 0.0 2.5 1 2 4 1 2 4 Weeks Weeks

0.8 5.5 c Control d) Treatment A 0.6 ) Treatment B 4.5

0.4 3.5 Control 0.2 2.5 Treatment A

Treatment B Shannon div. (SE) Shannon div. Shannon div. (CV) Shannon div. 0.0 1.5 1 2 4 1 2 4 Weeks Weeks

Figure A.12 Shannon diversity for OTUs at 99% sequence identity with rare members included (a, b) and after rare members were removed (c, d).

0.8 1.0 Control a) Treatment A b) 0.6 Treatment B 0.8

0.6 0.4 0.4 0.2 Control

0.2 Treatment A Evenness (SE) Evenness Evenness (CV) Evenness Treatment B 0.0 0.0 1 2 4 1 2 4 Weeks Weeks

0.8 1.0 Control c) Treatment A d) 0.6 Treatment B 0.8

0.6 0.4 0.4 0.2 Control

0.2 Treatment A Evenness (SE) Evenness Evenness (CV) Evenness Treatment B 0.0 0.0 1 2 4 1 2 4 Weeks Weeks

Figure A.13 Evenness for OTUs at 99% sequence identity with rare members included (a, b) and after rare members were removed (c, d).

Table A.3 SIMPER analysis representing the top OTUs responsible for 50% of the differences between samples (i.e. treatments-week).

Control Control Control Treat. A Treat. A Treat. A Treat. B Treat. B Treat. B Taxon 1w 2w 3w 1w 2w 3w 1w 2w 3w Domain [%] Phylum [%] Class [%] Order [%] Family [%] Genus [%] OTU015872 6.5 27.7 12.1 17.6 4.8 15.1 1.0 1.7 10.7 Bacteria 100 Proteobacteria 100 Betaproteobacteria 100 Burkholderiales 100 Alcaligenaceae 100 Paenalcaligenes 100 OTU055033 12.9 3.6 3.5 16.3 10.6 8.3 14.5 10.0 11.2 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Clostridiales_(IS)_XI 100 Tepidimicrobium 100 OTU000001 0.0 0.1 0.0 0.0 7.1 0.2 0.0 16.9 1.8 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 Bacillaceae 100 OTU015874 0.1 2.7 8.2 0.1 1.1 7.3 0.1 0.5 6.3 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 OTU015887 1.9 0.4 0.0 5.1 6.4 0.3 4.8 5.1 2.1 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 Bacillaceae 100 OTU000039 3.9 0.1 0.0 10.5 1.1 0.2 0.0 0.4 0.1 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 Bacillaceae 100 OTU000037 0.0 12.2 0.6 0.0 0.0 0.0 0.0 0.0 0.0 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Lachnospiraceae 100 OTU000007 0.0 0.0 9.7 0.0 0.0 0.0 0.0 0.0 0.0 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Lachnospiraceae 100 OTU015868 1.9 2.8 1.2 0.3 2.6 4.8 0.4 0.8 3.5 Bacteria 100 OTU000032 0.3 0.9 0.0 0.0 6.8 1.2 0.0 0.0 0.0 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 OTU000443 0.0 0.0 0.0 1.7 1.5 0.0 0.0 8.1 0.0 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 OTU015867 1.2 0.7 0.3 2.0 1.7 3.0 4.3 1.6 2.9 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 OTU001872 4.4 0.4 0.0 0.1 0.5 0.5 0.3 3.5 1.1 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 OTU000040 0.0 0.0 5.5 0.0 0.0 0.0 0.0 0.0 0.0 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Lachnospiraceae 100 Clostridium_XlVa 100 OTU000772 0.0 0.0 0.0 0.6 0.1 0.2 4.8 0.2 0.1 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 OTU014961 0.0 0.0 0.0 4.0 0.9 0.7 0.0 0.0 0.1 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Clostridiales_(IS)_XI 100 Tepidimicrobium 100 OTU002613 2.5 0.3 0.0 0.0 1.4 1.1 0.0 1.0 0.4 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 Planococcaceae 100 OTU000315 1.8 0.0 0.0 1.6 1.8 0.2 0.2 1.6 0.2 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 Bacillaceae 100 OTU000030 0.0 0.0 4.7 0.0 0.0 0.0 0.0 0.0 0.0 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Ruminococcaceae 100 Clostridium_III 100 OTU001140 1.0 0.3 0.4 2.0 1.1 1.7 1.4 0.9 1.9 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Clostridiales_(IS)_XI 53 Tepidimicrobium 53 OTU081194 0.0 0.0 0.0 2.0 1.0 0.1 1.6 0.4 0.4 Bacteria 100 Firmicutes 100 OTU015871 1.8 1.9 0.9 0.7 1.2 2.3 0.8 2.4 3.1 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 OTU015878 1.2 1.0 0.3 0.0 0.3 1.9 0.0 0.4 1.4 Bacteria 100 Proteobacteria 100 Betaproteobacteria 100 Burkholderiales 100 Alcaligenaceae 100 OTU005593 0.1 0.0 0.0 0.1 0.0 0.0 3.8 0.0 0.0 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Clostridiales_(IS)_XI 100 Sporanaerobacter 100 OTU002602 3.1 0.7 0.0 0.0 0.2 0.1 0.0 0.1 0.4 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 Bacillaceae 100 OTU003217 2.6 0.2 0.0 0.3 0.2 1.5 0.0 0.3 0.1 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 Bacillaceae 100 OTU001668 0.0 0.0 0.0 0.0 1.5 1.5 0.0 0.0 1.8 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 OTU001134 2.0 1.1 0.2 1.0 0.7 1.2 1.6 1.3 1.3 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 OTU015901 0.1 0.7 2.1 0.3 0.2 0.1 0.1 0.8 0.4 Bacteria 100 OTU015905 0.9 0.3 0.0 0.5 0.7 0.2 2.0 0.8 0.4 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 Planococcaceae 100 Ureibacillus 100 OTU054141 0.0 0.0 2.8 0.0 0.0 0.0 0.0 0.0 0.0 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Lachnospiraceae 100 Clostridium_XlVa 100 OTU015900 0.7 0.1 0.0 1.2 0.6 0.2 1.1 1.2 0.6 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Clostridiales_(IS)_XI 100 Tissierella 100 OTU000049 0.0 0.3 2.6 0.0 0.0 0.0 0.0 0.0 0.0 Bacteria 100 Firmicutes 100 OTU000035 0.0 0.0 0.0 0.3 0.0 0.0 2.5 0.0 0.0 Bacteria 100 Firmicutes 100 Clostridia 100 OTU015701 0.2 0.1 0.0 0.1 0.1 0.1 2.5 0.1 0.2 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Clostridiales_(IS)_XI 100 Tepidimicrobium 100 OTU002093 0.0 0.1 0.0 1.4 0.4 0.4 0.0 0.9 0.9 Bacteria 100 Firmicutes 100 OTU016560 0.0 0.0 2.6 0.0 0.0 0.0 0.0 0.0 0.0 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100 Ruminococcaceae 100 Clostridium_III 100 OTU000354 0.4 0.5 0.0 0.0 0.3 1.7 0.0 0.1 0.8 Bacteria 100 Firmicutes 100 Bacilli 100 Bacillales 100 Bacillaceae 100 OTU015870 0.2 0.1 0.0 0.3 0.1 0.3 1.9 0.1 0.3 Bacteria 100 Firmicutes 100 Clostridia 100 Clostridiales 100

Figure A.14 Components in each composter (%w/w).

Literature Cited

1. E1-Badawy, A. M.; Silva, R. G.; Morris, B.; Scheckel, K. G.; Suidan, M. T.; Tolaymat, T. M. Surface Charge-Dependent Toxicity of Silver Nanoparticles. Environ. Sci. Technol. 2011, 45, 283-287.

2. Ravel, B.; Newville, M. ATHENA, ARTEMIS, HEPHAESTUS: data analysis for X-ray absorption spectroscopy using IFEFFIT. J. Synchrotron Rad. 2005, 12(4), 537–541.

3. Malinowski, E.R. Factor Analysis in Chemistry. John Wiley. New York. , 1991

4. Webb, S.M. SIXpack: a graphical user interface for XAS analysis using IFEFFIT. Physica Scripta. 2005, T115, 1011-1014.

Chapter 3

Anaerobic Toxicity of Cationic Silver Nanoparticles

This chapter has been published as

Gitipour, A.; Thiel, S. W.; Scheckel, K. G.; Tolaymat, T. Anaerobic toxicity of cationic silver nanoparticles. Sci. Total. Environ. 2016. 557-558, 363-368.

Abstract

The microbial toxicity of silver nanoparticles (AgNPs) stabilized with different capping agents was compared to that of Ag+ under anaerobic conditions. Three AgNPs were investigated: (1) negatively charged citrate-coated AgNPs (citrate-AgNPs), (2) minimally charged polyvinylpyrrolidone coated AgNPs (PVP-AgNPs) and (3) positively charged branched polyethyleneimine coated AgNPs (BPEI-AgNPs). The AgNPs investigated in this experiment were similar in size (10-15nm), spherical in shape, but varied in surface charge which ranged from highly negative to highly positive. While, at AgNPs concentrations lower than 5 mgL-1, the anaerobic decomposition process was not influenced by the presence of the nanoparticles, there was an observed impact on the diversity of the microbial community. At elevated concentrations

(100 mgL-1 as silver), only the cationic BPEI-AgNPs demonstrated toxicity similar in magnitude to that of Ag+. Both citrate and PVP-AgNPs did not exhibit toxicity at the 100 mgL-1 as measured by biogas evolution. These findings further indicate the varying modes of action for nanoparticle toxicity and represent one of the few studies that evaluate end-of-life management concerns with regards to the increasing use of nanomaterials in our everyday life. These findings also highlight some of the concerns with a one size fits all approach to the evaluation of environmental health and safety concerns associated with the use of nanoparticles.

3.1 Introduction

Nanotechnology has seen a dramatic increase in utilization in environmental, medical, chemical and pharmaceutical industries (1). Engineered nanoparticles (ENPs) possess unique characteristics, compared to their bulk counterparts, such as the large surface area to volume ratio, high chemical reactivity, unique antimicrobial/ fungicidal activity, and biocompatible surface properties (2). Most of these properties are directly attributed to the small particle size of nanomaterials and only manifest within a particular size regime (3). These unique properties of materials at the nanoscale raise the need to examine the behavior of these particles in diverse environmental scenarios. Fundamental research on the fate, transport and toxicity of NPs is essential within an effort to determine their impact on natural and engineered environmental systems such as groundwater, soils and sediments, wastewater treatment and municipal solid waste management systems.

In particular, silver which has been well known for its antibacterial properties for centuries has become among the most commercially used nanoparticles (4). Currently silver nanoparticles are employed in many consumer products such as textiles, biomedical products, plastics, socks, food storage containers and various cleaning products (5). These particles are synthesized using various techniques to produce particles with different chemical and physical characteristics (6). Most of the AgNPs incorporated into consumer products are coated, surface modified and or functionalized to achieve certain properties (7, 8). Different surface functionalization, obtained by applying different capping agents, was shown to influence the toxicity, aggregation and dissolution of

AgNPs (9, 10).

A considerable amount of silver may be released from silver nanoparticle-containing consumer products within just a few washing cycles (11, 12). Once released, depending on the circumstances,

AgNPs can end up in a variety of environmental scenarios. It is highly likely that AgNPs travel through municipal sewer lines and reach wastewater treatment plants (WWTPs) and potentially accumulate in the biosolids (13-15).

One of the primary methods of biosolids treatment utilized is anaerobic digestion which relies on active anaerobic bacterial communities to degrade the organics (16). Therefore, it is vital to assess the potential impacts AgNPs may have on these bacterial species under anaerobic conditions and also the possible transformations of these particles within these systems. In addition to aggregation,

AgNPs may undergo surface changes that influence their behavior. For example, in a system with elevated chlorides and sulfides levels (e.g., wastewater, composter leachate and landfill leachate), the released AgNPs may transform to silver chlorides (AgCl) and silver sulfides (Ag2S) which are less toxic relative to ionic Ag and metallic AgNPs (17-19).

It is worthy to mention that the antibacterial properties of AgNPs on microorganisms under aerobic conditions have been significantly studied (20). The antibacterial mechanism of silver nanoparticles is linked to a combination of release of Ag+ by AgNPs oxidative dissolution or specific nanoparticle properties such as Trojan-horse type mechanism and generation of reactive oxygen species (ROS) leading to cell membrane damage (15, 21). Comparatively, the antibacterial activity of AgNPs under anaerobic conditions has been less studied and understood. According to

Kim et al., 2010, municipal wastewater treatment plants control the flows of silver between anthropogenic and environmental compartments (18). Therefore, the current study aims at investigating the effect of AgNPs on the anaerobic degradation process in simulated environmental systems. The antibacterial impacts of different AgNPs on the digestion process of anaerobic biosolids were studied at various concentrations and compared to that of Ag+. Based on the

69 observed surface charge dependent impacts, the nanoparticles exhibiting toxicity were further investigated by performing real-time speciation analysis and taxonomical analysis.

3.2 Experimental Section

3.2.1 Nanoparticles Synthesis, Purification, and Characterization

Three types of AgNPs with different capping agents were utilized, citrate coated AgNPs (citrate-

AgNPs), Polyvinylpyrrolidone coated AgNPs (PVP-AgNPs), and branched polyethyleneimine- coated AgNPs (BPEI-AgNPs). The nanomaterials were prepared and purified as described by El

Badawy et al. (22). The hydrodynamic diameter (HDD) and zeta potential () of the AgNPs were measured using a Zetasizer Nanoseries (Malvern Instruments). Transmission electron microscopy

(TEM) was used to verify nanoparticles’ size and shape. TEM samples were prepared by depositing a drop of the particular nanoparticle suspension on a carbon coated copper grid. Samples were air- dried at room temperature overnight in a dust-free box. Images were captured using a JEOL-1200

EX TEM (JEOL Inc.) operated at 120 kV. The total Ag concentration of the suspensions was measured using a PerkinElmer AAnalyst 800 atomic absorption spectrometer after performing a microwave acid digestion following EPA method 3015A.

3.2.2 Anaerobic Digesters Setup

Anaerobic biosolids were collected from a wastewater treatment plant. The characteristics of the biosolids are presented in Table S1 in supporting information (SI). In an anaerobic chamber, 30 serum bottles (250 mL total volume) were each filled with 120 mL of biosolids, sealed, covered, and purged with argon for 5 minutes to initiate anaerobic conditions. The biosolids containing bottles were then placed in a temperature-controlled room at 37 ℃ and continuously mixed using a bench top shaker. The samples were given two weeks to equilibrate after which each sample

70 received 1 mL cellulose solution (100 gL-1). The experimental setup included three nanomaterial

+ + treatments and a positive control containing Ag (AgNO3 was used as the source of Ag ). Four dosing levels were examined (0.5, 1, 5 and 100 mgL-1 as Ag). For consistency, after dosing with the various silver concentrations, de-aired DI-water was added to each serum bottle to achieve a final volume of 150 ml. A negative control composed of biosolids only (no addition of Ag) was also examined. All samples were prepared in triplicates.

3.2.3 Sampling and Analysis

3.2.3.1 Biogas Analysis

Gas sampling occurred immediately after the addition of treatments (time 0) and after 2, 5, 10, 17 and 28 days. The duration of the experiment was chosen to mimic the anaerobic biosolids retention time (25-30 days) at the WWTP from where the sludge was collected. The gas volume generated was measured using an 18-gauge needle and an airtight glass syringe according to US EPA OPPTS method 835.3400 (23). The gas composition (O2, CH4, and CO2) was evaluated using a gas chromatograph equipped with a thermal conductivity detector (GC/TCD; Agilent 6980N).

3.2.3.2 Taxonomical Analysis

Concurrent to the gas analysis, a 2 ml sludge aliquot was taken from each serum bottle and centrifuged at 5000 rpm for 15 minutes. After decanting the supernatant, the sample was preserved at -80 oC for taxonomical analysis. Total genomic DNA was extracted from the anaerobic sludge samples using a QIAamp stool DNA mini kit as described by Dowd et al. (24). Because of the intrinsic uncertainty of the analysis, the data are presented to demonstrate a general trend of the microbial communities and not for their absolute values.

3.2.3.3 Silver Speciation Analysis by X-ray Absorption Spectroscopy (XAS)

To evaluate changes in Ag speciation, X-ray absorption spectroscopy (XAS) was conducted at

Sector 10-ID (25) of the Advanced Photon Source at Argonne National Laboratory (ANL),

Argonne, IL. The analysis followed the experimental setup and sample analysis described by

Lombi et al. (26). To avoid changes in metal speciation during the sample preparation process, the anaerobic biosolids were dosed with the particular AgNP treatment in its natural liquid state, under anaerobic conditions, and immediately analyzed under a real-time setting in a reaction cell to observe the reactions occurring. In a 1 mL cell, 800 L of anaerobic sludge was added. Before analysis, 200 L of the BPEI-AgNPs (500 mgL-1) was injected into the cell giving a final concentration of 100 mgL-1 silver and a known concentration of sulfide. The contents of the cell were examined by XAS utilizing a quick scan set up. In addition to the real-time analysis of the

BPEI-AgNPs in natural sludge, additional samples with elevated sulfide levels in the sludge were analyzed operating under the exact experimental settings as described above.

3.2.3.4 Modelling Speciation of Dissolved Ag (MINTEQ)

The chemical speciation of dissolved silver (Ag+) in anaerobic biosolids is an important factor for understanding the toxicity, mobility and bioavailability of the metal within the environment (27).

The chemical speciation of silver was modeled using Visual MINTEQ, an equilibrium model that was developed based on MINTEQ2 (28). The model generates data used to determine the chemical speciation and bioavailability of dissolved Ag+ after introduction to the anaerobic biosolids. The visual MINTEQ2 input parameters, used for the study, were based on the initial characteristics of the utilized anaerobic biosolids (Table S1 in SI). These parameters include pH, redox potential, and dissolved metals. Sulfate, sulfide, phosphate, chloride, and nitrate concentrations were also included in the model due to their importance in dissolved metal complexations (29).

3.2.3.5 Statistical Analysis

The data were analyzed using one-way analysis of variance (ANOVA) test (SigmaPlot 12.0, Systat

Software Inc.) where the statistical significance was calculated using untreated samples (control) and treated samples (AgNO3 and AgNPs) and a p-value > 0.05 were considered as significant.

3.3 Results and Discussion

3.3.1 AgNPs Characterization

Silver nanoparticles with three different capping agents were characterized. The fundamental characteristics that may contribute to the toxicological behavior of these AgNPs (e.g., the (HDD) and the ()) are summarized in Table S2 of the SI (30). The AgNPs synthesized for this experiment were spherical and in the size range between 10 and 15 nm, well below the size at which nanoparticles exhibit unique properties (31). The AgNPs size distributions obtained from the TEM images (Figures S1, S2, and S3, SI) were consistent with those obtained by dynamic light scattering (DLS). The particles used varied in surface charge which ranged from -40 to +40 mV and stabilization mechanisms (see Tables S2, and SI). It is noteworthy to point out that the Ag+ concentrations in the AgNP suspensions were minimal since an ultrafiltration system was used to clean the suspensions (32).

3.3.2 Biogas Volume

The cumulative biogas production after the 28 days anaerobic incubation period is shown in Figure

3.1. At relatively low doses (< 5 mgL-1 as Ag), there was no statistically significant difference

(p>0.05) between the treatment types. These results are in line with previous studies that found methanogenesis to be unaffected by AgNPs at similar concentrations (33, 34). Interestingly, as the

AgNPs concentrations increased to 100 mgL-1 statistically significant differences regarding

73 inhibition were observed between treatment types. At that level, the biogas production of serum bottles containing citrate and PVP coated AgNPs showed no significant difference (p>0.05) compared to the background (negative control). In the case of citrate-AgNPs, the lack of toxicity associated with these nanoparticles might be caused by aggregation and subsequent settling and as a result, the nanoparticles were less likely to interact with the microbial community.

Electrostatically stabilized nanoparticles, like citrate-AgNPs, readily aggregate in solutions with high ionic strength. On the other hand, sterically stabilized nanoparticles like PVP-AgNPs are more stable in solutions with high ionic strength like sludge (35). The lack of surface charge on

PVP-AgNPs may have resulted in lower attraction between the nanoparticles and the negatively charged bacterial membrane thus manifesting a low toxicity.

While citrate and PVP-AgNPs showed no toxicity at the 100 mgL-1 level, both the Ag+ and the

BPEI-AgNPs treatments showed near complete inhibition of the anaerobic biogas production process. The oxidative dissolution of Ag+ may play a role in the aerobic toxicity of AgNPs, but the anaerobic conditions are unfavorable for such dissolution (36). Since all of the synthesized AgNPs were in the same size range, size as a parameter of toxicity, can be ruled out leaving surface characteristics as the distinguishing factor among the particles. Citrate-AgNPs have a negative surface charge, PVP-AgNPs are neutral, and BPEI-AgNPs are positively charged. Taking into account that the vast majority of the bacterial cells in the anaerobic sludge are negatively charged

(37), it is probable that the surface charge of the BPEI-AgNPs would give these particles affinity to the negatively charged bacterial walls while the other particles would not behave in the same manner.

These findings were not surprising as similar toxicity trends were observed with these nanoparticles under aerobic environments (38). Furthermore, Ivask et al. concluded that under

74 aerobic conditions, BPEI-AgNPs impact on E. coli is specific to this type of nanoparticles (39).

The positively charged AgNPs are theorized to attach onto the surface of the bacteria and cause cell lysis through physical interaction.

3.3.3 Taxonomical Analysis

The biogas data show that at the 100 mgL-1 Ag dose level, there was no biogas generation in both the BPEI-AgNPs and the Ag+ treatments which was supported by the taxonomical analysis conducted on these samples. At 5 mgL-1 dose, the biogas production of the BPEI-AgNPs treated samples showed similar trends (p>0.05) to both the negative and positive controls. However, the taxonomical analysis suggests otherwise. For example, Figure 3.2 represents the taxonomical data for one of the most abundant bacterium present in the samples (Flavobacterium). As can be observed, there seems to be time dependent variability in the response of the bacterium to BPEI-

AgNPs as opposed to the negative or the positive control. In this case, it appears that the population of this bacterium is increasing over time in the BPEI treatment and remaining rather constant in the other treatments. Other bacteria evaluated (Table B3, Appendix B) showed similar time- dependent variations.

These results seem to reinforce the concept that the functional redundancy often built into a complex bacterial community, like the one found in biosolids, maintains the best level of performance especially at low AgNPs dosing levels (19). The results further support the observation that BPEI-AgNPs toxicity pathways are different than those associated with Ag+ only

(39). It is also important to note that the data shows that dominant bacterial communities responsible for anaerobic degradation are bacteria with a negative surface charge. This is of importance as the only nanoparticles that demonstrated a toxic and inhibitory effect on the digestion process were BPEI-AgNPs which have a positive surface charge.

3.3.4 XAS (X-ray Absorption Spectroscopy)

Previous studies indicate that transformation of AgNPs to Ag2S is to be expected during wastewater transport/treatment. A rapid conversion of AgNPs to Ag2S during sewage treatment was reported for one AgNP material by Kaegi et al. (40). The mechanism of this process has been studied in detail by Liu et al. who suggested that at high sulfide concentrations, like those present in the anaerobic digesters, oxysulfidation occurs by a direct particle-fluid reaction rather than through an oxidative dissolution of the AgNPs (41).

Since the only toxicity observed was with one of the capping agents (BPEI-AgNPs), these particles were chosen for further investigation. To further analyze the impact of sulfide concentrations on the rate of sulfidation, anaerobic biosolids containing two levels of sulfide were investigated; natural sulfide levels in biosolids ~15 mgL-1 and elevated to 100 mgL-1 sulfides. The concentration of AgNPs was kept constant at 100 mgL-1.

The XAS spectra demonstrated immediate sulfidation of the nanoparticles upon introduction to the media regardless of the sulfide level (Figure 3.3). Nevertheless, a higher percentage of sulfidation was observed for the particles in the elevated sulfide sludge. This is of much importance as the rate of sulfidation may have a direct correlation with the toxicity of AgNPs. These results demonstrate that although AgNPs partially sulfidize upon entering the anaerobic sludge media

(Table 3.1), the transformation to Ag2S was incomplete. The analysis shows that sulfidation kinetics of BPEI-AgNPs are dependent on the concentration of sulfur within the media. As demonstrated, the BPEI-AgNPs in biosolids with natural sulfide levels showed a slower sulfidation rate within the first 24 hours as compared to the same particles in biosolids with elevated sulfide levels. Therefore, based on the amount of sulfide in the system, the sulfidation kinetics vary

76 allowing BPEI-AgNPs to retain their metallic form for various times. Hence, allowing the particles to exhibit their particle specific toxicity.

3.3.5 Speciation of Dissolved Metals in Anaerobic Sludge (MINTEQ)

The chemical speciation of dissolved Ag+ in anaerobic biosolids was modeled using the biosolids parameters as input values summarized in Table B1, Appendix B. The input values produced chemical speciation data that were used to determine the form in which the dissolved silver has within the anaerobic biosolids and also the percentage of the dissolved metals present in their ionic state. Due to the nature of the media under investigation, it is not possible to identify and include all possible interactions in the system. Nevertheless, the chemical species that are known to impact the speciation of AgNPs were measured and considered for this simulation.

Based on these results (Table B4, Appendix B), less than 1% of the dissolved Ag present in the system would be free Ag+. At sulfide concentrations of ~15 ppm, found naturally in the biosolids,

Ag would be predominantly associated with hydrogen sulfide (AgHS (aq)) (94.61%) and also sulfide (AgS-) (3.427%). A small portion of the Ag was also associated with the chloride as (AgCl

(aq)) (1.13%). These results are similar to those reported by Bolyard et al. (42). The data further exclude the toxicity of BPEI-AgNPs due to oxidative dissolution of Ag and further supports the conclusion that other toxicity pathways are the primary mechanism for the toxicity of BPEI-

AgNPs.

3.4 Implications

At low concentrations, AgNPs seem not to drastically impact anaerobic degradation which is the result of the functional redundancy built within the microbial community and not the lack of toxicity of AgNPs. At relatively high concentrations the cationic BPEI-AgNPs demonstrate

77 elevated toxicity as compared with PVP-AgNPs and the anionic citrate-AgNPs. Because of the lack of oxidative dissolution in anaerobic environments, the results further support previous research that suggests a different toxicity mechanism is in place with BPEI-AgNPs as compared to other types of AgNPs. It must be noted that BPEI-AgNPs toxicity did not manifest until relatively high concentrations (100 mgL-1 as Ag). Under normal operating conditions in a WWTP,

AgNPs concentrations would not be expected to reach such levels. Thus, it is reasonable to believe that most WWTPs will be able to manage AgNPs without negatively impacting their operations.

The results of this study also demonstrate that a one size fits all approach to the evaluation of environmental health and safety of nanoparticles may not necessarily be accurate. Even within one class of nanoparticles, there appear to be differences that drastically impact their behavior in the environment.

Literature Cited

1. Nel, A.; Xia, T.; Madler, L.; Li, N. Toxic potential of materials at the nanolevel. Science. 2006, 311, 622-627.

2. Khaydarov, R.R.; Khaydarov, R. A.; Estrin, Y.; Evgrafova, S.; Scheper, T.; Endres, C.; Cho, S. Y. Silver Nanoparticles Environmental and Human Health Impacts. Nanomaterials: Risks and Benefits. Springer. 2009, 287-297.

3. Maynard, A.D., Michelson, E., 2006. The Nanotechnology Consumer Product Inventory; http://www.nanotechproject.org/44.

4. Tolaymat, T. M.; El Badawy, A. M.; Genaidy, A.; Scheckel, K. G.; Luxton, T. P.; Suidan, M. An evidence-based environmental perspective of manufactured silver nanoparticle in syntheses and applications: A systematic review and appraisal of peer-reviewed scientific papers. Sci. Total Environ. 2010, 408, 999-1006.

5. Dobias, J.; Bernier-Latmani, R. Silver release from silver nanoparticles in natural waters. Environ. Sci. Technol. 2013, 47, 4140-4146.

6. Iravani, S.; Korbekandi, H.; Mirmohammadi, S. V.; Zolfaghari, B. Synthesis of silver nanoparticles: chemical, physical and biological methods. Res. Pharm. Sci. 2014, 9, 385-406.

7. Reinhart, D. R.; Berge, N. D.; Santra, S.; Bolyard, S. C. Nanomaterial fate in landfills. Waste Manage. 2010, 30, 2020-2021.

8. Nowack, B. The behavior and effects of nanoparticles in the environment. Environ. Pollut. 2009, 157, 1063-1064.

9. Kvitek, L.; Panacek, A.; Soukupova, J.; Kolar, M.; Vecerova, R.; Prucek, R.; Holecova, M.; Zboril, R. Effect of surfactants and polymers on stability and antibacterial activity of silver nanoparticles (NPs). J. Phys. Chem. C 2008, 112, 5825-5834.

10. Unrine, J. M.; Coleman, B. P.; Bone, A. J.; Gondikas, A. P.; Matson, C. W. Biotic and abiotic interactions in aquatic microcosms determine the fate and toxicity of Ag nanoparticles. Environ. Sci. Technol. 2012, 46, 6915-6924.

11. Benn, T. M.; Westerhoff, P. Nanoparticle silver released into water from commercially available sock fabrics. Environ. Sci. Technol. 2008, 42, 4133-4139.

12. Geranio, L.; Heuberger, M.; Nowack, B. The behavior of silver nanotextiles during washing. Environ. Sci. Technol. 2009, 43, 8113-8118.

13. Kiser, M. A.; Ladner, D. A.; Hritovski, K. D.; Westerhoff, P. K. Nanomaterial transformation and association with fresh and freeze-dried wastewater activated sludge: Implications for testing protocol and environmental fate. Environ. Sci. Technol. 2012, 46, 7046-7053.

14. Shafer, M. M.; Overdier, J. T.; Armstrong, D. E. Removal, partitioning, and fate of silver and other metals in wastewater treatment plants and effluent-receiving streams. Environ. Toxicol. Chem. 1998, 17, 630-641.

15. Marambio-Jones, C.; Hoek, E. M. V. A review of the antibacterial effects of silver nanomaterials and potential implications for human health and the environment. J. Nanopart. Res. 2010, 12, 1531-1551.

16. Donoso-Bravo, A.; Mailier, J.; Martin, C.; Rodriguez, J.; Aceves-Lara, C. A.; Wouwer, A. V. Model selection, identification and validation in anaerobic digestion: a review. Water Res. 2011, 45, 5347-5364.

17. Levard, C.; Hotze, E. M.; Lowry, G. V.; BrownJr, G. E. Environmental transformations of silver nanoparticles: Impact on stability and toxicity. Environ. Sci. Technol. 2012, 46, 6900- 6914.

18. Kim, B.; Park, C.S.; Murayama, M.; Hochella, M.F. Discovery and characterization of silver sulfide nanoparticles in final sewage sludge products. Environ. Sci. Technol. 2010, 44, 7509- 7514.

19. Gitipour, A.; ElBadawy, A.; Arambewela, M.; Miller, B.; Scheckel, K.; Elk, M.; Ryu, H.; Gomez, V. A.; Santo Domingo, J.; Thiel, S.; Tolaymat, T. The impact of silver nanoparticles on the composting of municipal solid waste. Environ. Sci. Technol. 2013, 47, 14385-14393.

20. Choi, O.; Deng, K. K.; Kim, N. J.; Ross, L.; Surampalli, R. Y.; Hu, Z. The inhibitory effects of silver nanoparticles, silver ions and silver chloride colloids on microbial growth. Water Res. 2008, 42, 3066-3074.

21. Sotiriou, G. A.; Pratsinis, S. E. Antibacterial activity of nanosilver ions and particles. Environ. Sci. Technol. 2010, 44, 5649-5654.

22. El Badawy, A. M.; Luxton, T. P.; Silva, R. G.; Scheckel, K. G.; Suidan, M. T.; Tolaymat, T. M. Impact of environmental conditions (pH, ionic strength and electrolyte type) on the surface charge and aggregation of silver nanoparticles suspension. Environ. Sci. Technol. 2010, 44, 1260-1266.

23. USEPA Fate, transport, and transformation guidelines — OPPTS 835.3400 anaerobic biodegradability of organic chemicals, Washington, DC, 1998.

24. Dowd, S. E.; Callaway, T. R.; Wolcott, R. D.; Sun, Y.; McKeehan, T.; Hagevoort, R. G.; Edrington, T. S. Evaluation of the bacterial diversity in the feces of cattle using 16S rDNA bacterial tag-encoded FLX amplicon pyrosequencing (bTEFAP). BMC Microbiology. 2008, 8, 125.

25. The MRCAT Insertion Device Beamline at the Advanced Photon Source", C.U. Segre, N.E. Leyarovska, L.D. Chapman, W.M. Lavender, P.W. Plag, A.S. King, A.J. Kropf, B.A. Bunker,

K.M. Kemner, P. Dutta, R.S. Duran and J. Kaduk, CP521, Synchrotron Radiation Instrumentation: Eleventh U.S. National Conference, ed. P. Pianetta, et al., p419-422, (American Insitute of Physics, New York, 2000).

26. Lombi, E.; Donner, E.; Taheri, S.; Tavakkoli, E.; Jamting, A. K.; McClure, S.; Naidu, R.; Miller, B. W.; Scheckel, K. G.; Vasilev, K. Transformation of four silver/silver chloride nanoparticles during anaerobic treatment of wastewater and post-processing of sewage sludge. Environ. Pollut. 2013, 176, 193-197.

27. Levard, C.; Hotze, M. E.; Colman, B. P.; Dale, A. L.; Truong, L.; Yang, X. Y.; Bone, A. J.; Brown, Jr., G. E.; Tanguay, R. L.; Di Giulio, R. T.; Bernhardt, E. S.; Meyer, J. N.; Wiesner, M. R.; Lowry, G. V. Sulfidation of silver nanoparticles: Natural antidote to their toxicity. Environ. Sci. Technol. 2013, 47, 13440-13448.

28. Visual MINTEQ ver. 3.0, Website, http://www2.lwr.kth.se/English/OurSoftware/vminteq/index.html

29. Tejamaya, M.; Romer, I.; Merrifield, R. C.; Lead, J. R. Stability of citrate, pvp, and peg coated silver nanoparticles in ecotoxicology media. Environ. Sci. Technol. 2012, 46, 7011-7017.

30. Silva, T.; Pokhrel, L. R.; Dubey, B.; Tolaymat, T. M.; Maier, K. J.; Liu, X. Particle size, surface charge and concentration dependent ecotoxicity of three organo-coated silver nanoparticles: Comparison between general linear model-predicted and observed toxicity. Sci. Total. Environ. 2014, 468-469, 968-976.

31. Auffan, M.; Rose, J.; Bottero, J.; Lowry, G.; Jolivet, J.; Wiesner, M. Towards a definition of nanoparticles based on novel size-dependent properties. Nat. Nanotechnol. 2009, 3, 634-641.

32. El Badawy, A. M.; Scheckel, K. G.; Suidan, M.; Tolaymat, T. The impact of stabilization mechanism on the aggregation kinetics of silver nanoparticles. Sci. Total Environ. 2012, 429, 325-331.

33. Yang, Y.; Zhang, C.; Hu, Z. Impact of metallic and metal oxide nanoparticles on wastewater treatment and anaerobic digestion. Environ. Sci.: Processes Impacts. 2013, 15, 39-48.

34. Yang, Y.; Chen, Q.; Wall, J. D.; Hu, Z. Potential nanosilver impact on anaerobic digestion at moderate silver concentrations. Water Res. 2012, 46, 1176-1184.

35. An Huynh, K.; Chen, K. L. Aggregation kinetics of citrate and polyvinylpyrrolidone coated silver nanoparticles in monovalent and divalent electrolyte solutions. Environ. Sci. Technol. 2011, 429, 325-331.

36. Liu, J.; Hurt, R. H. Ion release kinetics and particle persistence in aqueous nano-silver colloids. Environ. Sci. Technol. 2010, 44, 2169-2175.

37. Jia, X. S.; Fang, H. P.; Furumai, H. Surface charge and extracellular polymer of sludge in the anaerobic degradation process. Wat. Sci. Tech. 1996, 34, 309-316.

38. El Badawy, A. M.; Silva, R. G. Morris, B.; Scheckel, K. G.; Suidan, M. T.; Tolaymat, T. M. Surface charge-dependent toxicity of silver nanoparticles. Environ. Sci. Technol. 2011, 45, 283-287.

39. Ivask, A.; ElBadawy, A.; Kaweeteerawat, C.; Boren, D.; Fischer, H.; Ji, Z.; Chang, C. H.; Liu, R.; Tolaymat, T.; Telesca, D.; Zink, J. I.; Cohen, Y.; Holden, P. A.; Godwin, H. A. Toxicity mechanisms in Escherichia coli vary for silver nanoparticles and differ from ionic silver. ACS Nano. 2014, 8, 374-386.

40. Kaegi, R.; Voegelin, A.; Sinnet, B.; Zuleeg, S.; Hagendorfer, H.; Burkhardt, M.; Siegrist, H. Behavior of metallic silver nanoparticles in a pilot wastewater treatment plant. Environ. Sci. Technol. 2011, 45, 3902-3908.

41. Liu, J.; Pennell, K. G.; Hurt, R. H. Kinetics and mechanisms of nanosilver oxysulfidation. Environ. Sci. Technol. 2011, 45, 7345-7353.

42. Bolyard, S. C.; Reinhart, D. R.; Swadeshmukul, S. Behavior of engineered nanoparticles in landfill leachate. Environ. Sci. Technol. 2013, 47, 8114-8122.

43. Mwilu, S. K.; El Badawy, A.; Bradham, K.; Thomas, D.; Scheckel, K. G.; Tolaymat, T. M.; Ma, L.; Rogers, K. Changes in Silver Nanoparticles Exposed to Human Synthetic Stomach Fluid: Effects of Particle Size and Surface Chemistry. Sci. Total Environ. 2013, 447, 90‐98.

Table 3.1 Ag speciation as identified by LCF of k-edge XANES spectra. Silver species proportions are presented as percentages. The R-factor indicates the goodness of fit.

Sample Time (h) Ag2S (%) Ag0 (%) R-factor (x 1000)

0 24 76 0.13

100ppm BPEI-AgNPs in Sludge 4 53 47 0.13 (15 mgL-1 Sulfide) 24 95 5 0.32

0 95 5 0.5

100ppm BPEI-AgNPs in Sludge 4 97 3 0.33 (100 mgL-1 Sulfide) 24 97 3 0.29

Figure 3.1 Cumulative gas volume production of anaerobic biosolids after exposure to different concentrations of AgNPs or Ag+ for a duration of 28 days.

Figure 3.2 Flavobacterium species among different treatments at 5 mgL-1 of silver.

APPENDIX B

B1. Synthesis of AgNPs

Citrate stabilized AgNPs (citrate-AgNPs)

-3 -2 A solution of 1x10 M AgNO3 (99.99%, Aldrich) was mixed with a solution of 1x10 M

Na3C6H5O7.2H2O (99+ %, SAFC supply solution) in a volume ratio of 2:1, respectively.

Subsequently, the mixture was heated to 70°C steadily for a duration of 4 hours inside a water bath.

Polyvinylpyrrolidone stabilized AgNPs (PVP-AgNPs)

-3 A solution of 5x10 M AgNO3 (99.99%, Aldrich) was added drop by drop (~1 drop per second)

-3 to a vigorously stirring ice-cold solution of 2x10 M NaBH4 in 1% PVP at a final volume ratio of 1:3, respectively.

Branched polyethyleneimine stabilized AgNPs (BPEI-AgNPs)

-1 Branched polyethyleneimine (BPEI) (99 %, Mw = 1.20 kg mol ) and AgNO3 were separately dissolved in 1x10-4 M solution of N-(2-hydroxyethyl) piperazine-N’-2-ethanesulfonic acid

(HEPES). The two solutions were then mixed in a volume ratio of 1:1 to give a final molar ratio of 0.5:1:0.1 BPEI: AgNO3: HEPES, respectively. After mixing, the solution was exposed to UV irradiation using a standard low-pressure mercury arc lamp for a duration of 2 hours.

Table B.1 Biosolids Characteristics*

Parameter Unit Value pH -- 7.0 Redox Potential (Eh) mV -200 Temperature ℃ 37 Ca+2 mg/L 500 Fe+2 mg/L 55 K+1 mg/L 75 Na+1 mg/L 100 Cl-1 mg/L 250 Mg+2 mg/L 50 Al+3 mg/L 25

O2(aq) mg/L 0.75 S-2 mg/L 15

- NO2 mg/L 50

-2 SO4 mg/L 25

-3 PO4 mg/L 125

* These characteristics were used as MINTEQ Input Parameters

Table B.2 Hydrodynamic diameter (HDD) and Zeta potential of particles.

AgNPs Properties Stabilization HDD (nm)  (mV) Mechanism

Citrate-AgNPs 10 -40 Electrostatic PVP-AgNPs 13 -4 Steric BPEI-AgNPs 10 +40 Electrosteric

Figure B.1 TEM image of Citrate-AgNPs.

Figure B.2 TEM images of PVP-AgNPs.

Figure B.3 TEM images of BPEI-AgNPs.

Table B.3 Most abundant bacterial species over 28-day period. (Most ubiquitous genera identified from the anaerobic biosolid)

Control Ag + BPEI-AgNPs Bacterial Species 0 2 5 10 17 28 0 2 5 10 17 28 0 2 5 10 17 28 flavobacterium spp. 5413 7631 6101 7074 6932 12299 7592 9786 9156 9270 8163 7444 12816 13630 12957 19583 19006 37164 bacteroides spp. 12701 11965 9597 8407 9471 6621 8177 7121 7132 7015 8199 6247 4338 4457 3384 4389 3770 2273 bacteroides sp. 8611 9974 7453 6811 5940 2685 5562 6348 6392 4763 5169 4969 4013 3611 2583 2882 1210 904 marinoscillum spp. 7775 9839 7584 9367 6223 6544 6303 8106 7267 5605 6108 7216 3425 3335 3015 3547 2754 2672 ignavibacterium album 2032 3109 3841 5420 8846 7310 7999 6296 6010 4792 3122 2803 11355 8936 7870 11339 10853 9084 coprothermobacter spp. 5668 7080 5019 5644 3529 2689 3931 3956 3471 3828 5025 8420 1688 1912 1342 1384 393 583 clostridium sp. 2320 3410 2995 1987 2804 2848 2864 2823 2622 2118 1976 2282 8010 7269 6069 7372 5697 2350 candidatus endomicrobium 1213 1646 1235 1170 1184 1006 4364 5044 6132 2981 2812 4947 5033 4230 3797 4448 2975 1706 treponema spp. 2378 2860 2099 2747 2763 5137 1875 2019 1757 1662 1709 1684 2022 1933 1746 2420 2051 3648 sphingobacterium spp. 2552 2906 2330 2618 3524 5979 2013 2140 1780 1179 1004 391 2384 2428 2085 2543 2582 1006 clostridium spp. 3619 3515 2991 511 1232 3486 541 775 2003 791 1716 820 537 504 761 2056 486 9093 thermosipho spp. 1298 1826 1221 1378 1243 2817 1725 1964 1558 1757 1249 1708 2045 2306 1669 2297 1671 2602 petrimonas spp. 1664 2015 1472 1713 1855 1495 1296 1571 1747 1796 2044 2508 996 1320 980 1579 1046 1202 synergistes spp. 1229 1751 1257 1035 1318 1922 1036 1270 1241 1322 1537 1815 1843 1942 1499 1912 1782 3800 tepidiphilus margaritifer 1465 2245 1432 1367 1270 820 1385 1156 1012 1081 1244 1459 922 1140 740 764 253 93 bellilinea spp. 711 915 727 935 1045 2004 708 962 1118 1461 1422 1720 835 830 857 1322 1345 2625 aminiphilus circumscriptus 332 453 443 408 682 1382 603 679 736 1222 1252 1677 810 978 831 1630 1708 2066 treponema zuelzerae 398 406 277 201 244 118 896 1080 867 549 393 966 1875 1338 892 837 396 52 eubacterium spp. 640 852 591 659 688 833 525 697 624 403 360 403 719 719 704 832 640 972 tepidiphilus spp. 570 869 584 550 490 278 547 480 440 467 474 601 382 500 337 379 118 42 fervidobacterium spp. 489 590 359 410 474 314 447 565 444 353 380 334 326 329 245 332 165 77 proteiniphilum spp. 397 520 352 378 488 387 339 422 513 479 604 793 177 205 150 258 147 102

93 leptolinea spp. 46 65 50 113 330 1720 46 49 72 69 274 306 85 104 181 729 980 2459 acetivibrio spp. 72 243 606 901 979 173 144 133 450 577 289 163 310 246 227 261 387 662 clostridium intestinale 779 572 327 86 92 92 395 390 185 115 102 95 421 380 224 293 185 50 bacillus spp. 184 245 206 160 229 173 106 110 239 1240 980 1460 161 157 98 170 209 708 ruminococcus sp. 48 82 76 96 318 123 248 525 422 117 126 249 994 1020 641 458 251 64 lachnoclostridium clostridium 500 481 274 133 39 14 378 306 168 112 50 23 310 267 115 112 37 13 pleomorphomonas koreensis 126 223 150 189 376 412 219 295 281 521 657 465 211 218 131 293 196 66 petrotoga olearia 145 138 127 85 96 90 117 117 108 134 654 1298 113 106 77 94 28 16 ruminobacillus spp. 217 239 181 140 143 58 168 211 216 167 235 160 98 110 76 97 21 31 spirochaeta spp. 181 220 167 200 176 151 185 192 147 123 163 142 141 145 148 165 95 102 lutispora spp. 49 99 100 384 376 772 102 106 82 82 86 162 131 138 143 127 117 368 ignavibacterium spp. 37 50 80 73 147 482 100 82 122 86 124 379 126 129 64 169 123 1074 ruminiclostridium clostridium 320 311 160 33 10 14 113 163 200 17 67 10 11 16 8 13 14 18 petrotoga spp. 172 165 113 82 148 97 175 164 144 141 153 194 90 87 56 83 50 19 symbiobacterium spp. 167 145 115 84 200 85 202 115 94 49 92 68 180 167 101 122 51 30 propionispora spp. 95 149 115 93 140 125 113 118 136 158 177 226 79 121 92 98 90 106 spirochaeta sp. 128 89 72 56 67 27 359 340 315 219 133 85 75 96 56 40 21 13 alcaligenes faecalis 166 150 98 38 49 27 137 126 75 45 21 39 127 118 40 59 15 5 geodermatophilus nigrescens 116 127 79 68 74 52 29 28 22 44 31 46 46 51 29 65 88 533 dysgonomonas mossii 12 13 9 3 11 6 5 13 104 621 489 555 13 16 9 5 3 22 levilinea saccharolytica 36 51 33 32 42 128 36 51 92 82 89 82 77 85 59 126 89 265 cloacibacillus spp. 111 103 88 114 99 152 20 24 28 33 47 39 69 53 32 36 43 84 thermacetogenium spp. 36 39 37 70 120 104 49 72 46 95 111 148 48 77 50 93 99 165 geothermobacterium spp. 36 46 44 47 61 87 42 55 70 42 45 51 94 94 114 123 107 257 citrobacter spp. 144 138 62 43 35 8 44 25 24 7 3 3 27 26 7 18 3 0 Total sequences 67398 80560 63329 64113 66602 74146 64260 69070 67864 59790 61160 67655 70588 67879 57291 77954 64350 91246

Table B.4 MINTEQ Speciation Output

Component Species Name % of total concentration

Ag+1 0.138

AgCl (aq) 1.133 Ag+1 AgCl2- 0.646

AgHS (aq) 94.606

AgS- 3.427

AgNO2 (aq) 0.046

Figure B.4 Schematic of the ultrafiltration system used for the purification of AgNP suspensions. The processing reservoir contains the AgNPs suspension and the buffer reservoir contains Milli- Q water (1). The polyethersulfone ultrafiltration membrane is encircled.

CHAPTER 4

Nanosilver as a Disinfectant in Dental Unit Waterlines: Assessment of the Physicochemical Transformations of the AgNPs

This chapter has been submitted for publication as

Gitipour, A.; Al-Abed. S. R.; Thiel, S. W.; Scheckel, K. G.; Tolaymat, T. Nanosilver as a Disinfectant in

Dental Unit Waterlines: Assessment of the Physicochemical Transformations of the AgNPs. Chemosphere.

2016, submitted.

Abstract

Containing the essential elements for the promotion of bacterial growth, dental unit water lines

(DUWL) are susceptible to biofilm development, leading to water contamination and potential health hazards impacting dental patients as well as dental care personnel. To prevent health risks associated with water contamination a broad list of disinfecting agents are used to hinder/decrease biofilm buildup. Silver nanoparticles are often utilized for disinfection purposes at dental offices and are commercially available as a DUWL disinfectant. Due to the antibacterial characteristics of

AgNPs, there is growing concern with the release of AgNPs into natural or engineered environmental systems which rely on microbial communities for proper functioning. This study monitors the physiochemical transformations of a commonly used nanosilver disinfectant (ASAP-

AGX-32, an antimicrobial cleaner for dental units, 0.0032% Ag) as a treatment for biofilm and bacterial growth in DUWL. To simulate the disinfection scenario, an in-house DUWL model was assembled and biofilm buildup was allowed. After considerable levels of biofilm growth in the dental lines were confirmed, the disinfection process was simulated according to the manufacturer’s instructions. The interactions between the AgNPs and biofilm were assessed in the process, concentrating on the physicochemical characteristics of the AgNPs and the morphology of the biofilm development. As a confirmation of the biofilm growth, as well as, effectiveness of the AgNPs as a disinfectant, the colony forming units (CFU/ml) of both the effluent water and inner surface of the DUWL were determined at varying times throughout the process. Additionally, the morphology of the biofilm development within the inner surface of the tubing and the size distribution of the silver nanoparticles were examined over the duration of the disinfection process by SEM and TEM, respectively. The pristine nanosilver particles measured between 3-5 nm in diameter and were surrounded by a stabilizing polymer. The polymeric

98 stabilizing agent encapsulating the pristine AgNPs was no longer present after the disinfection process, potentially adsorbing onto the biofilm surface. The diminishing of the polymeric stabilizing agent may have initiated AgNP aggregation. TEM images of the AgNPs after the disinfection process illustrated the existence of aggregated AgNPs measuring between 50-200 nm, as well as, AgNPs in the initial size range. Finally, the surface transformations of the AgNPs were identified before and after the disinfection process by XPS and XAS analysis. The physiochemical characteristics (e.g., size and surface functionalization) of silver nanoparticles are known to govern their fate, transport and environmental implications and therefore, knowledge of the characteristics of the AgNPs after the disinfection process (usage scenario) is of significance. In addition to further understanding the transformations that occur in the process, adsorption of the AgNPs onto the biofilms surface was confirmed using backscattered and secondary electron SEM. This demonstration may help in further explaining the mechanism of AgNPs toxicity to biofilms.

4.1 Introduction

In recent times, nanomaterials have found applications in many aspects of our daily life. They have been progressively incorporated into a wide range of consumer products. Paints, textiles, personal care products, medical apparatus, batteries, and sporting goods are just a few examples of such products currently available in the market. Attributed to their size, nanomaterials have shown to display specific characteristics such as high surface-to-volume ratio and higher reactivity in comparison to the bulk material. Today, nanomaterials are synthesized utilizing various techniques and methods leading to an extensive variety of nanomaterials encompassing different characteristics.

The widespread use of nanomaterials in commercial products may lead to their release and accumulation in the environment (e.g., soil and water) (1). Depending on the inherent nature and characteristics of the nanomaterial, environmental exposure may lead to negative ecological impacts. Silver nanoparticles, which are well-known and primarily utilized for their antibacterial and antifungal properties, are currently the most extensively used nanomaterial in consumer products (2). In the case of silver nanoparticles (AgNPs), based on their antibacterial characteristics, there is concern regarding their possible impacts on beneficial microbial communities in the ecosystem and also waste management processes that heavily depend on microbial degradation (3).

Microbial and biofilm growth in water delivery lines are a persistent problem facing the dental industry (4). Typically, dental unit waterlines contain opaque, small-diameter tubing that carry water to and from the patient's mouth during dental procedures. Due to their small inner diameter and composition, mainly flexible polyurethane or other plastics, dental lines are an excellent environment for the development of microbial biofilms (5, 6). The inner surface of the dental

100 tubing acts as a substrate, allowing bacteria to use the hardeners and additives as a nutrient source promoting biofilm growth (5). The biofilm and bacteria will eventually percolate into the water stream, posing health risks to patients as well as dental health care personnel. Occupational asthma has been reported as a potential risk factor dental staff face from inhalation of aerosols generated from the biofilm contaminated dental unit waterlines (7). Also, the biofilm cause a foul odor and are known to give water a bad taste (8).

In light of the above, a clean water delivery system is essential for dental procedures. In the past, instruments such as ultrasonic scalers were utilized for reducing the growth of microorganisms

(9). Other more common methods used for preventing contamination include installation of independent water systems, water filtration and sterilization systems or a variety of chemical treatment protocols (10). It is important to note that sterile water delivery systems have proven to be expensive to purchase, maintain and less convenient to use compared to the conventional water delivery systems (11). As for chemical treatments, a variety of chemicals have been examined for their decontamination capability against these microorganisms and biofilms (e.g. hydrogen peroxide, ethanol and ozone). Dependent on the intrinsic nature of the chemical agents, the application procedures vary, ranging from intermittent use to continuous introduction to the water supply. However, the potential implications of the chemical treatments on different components of the dental unit, and their biochemical interactions are not always known or predictable (10).

Therefore, it is important to investigate the safety and efficacy of chemical treatments to ensure patient and dental personnel protection.

AgNPs have been established as a beneficial disinfecting agent with applications in countless fields. In dentistry, AgNPs are utilized as a means to prevent or reduce biofilm growth over dental material surfaces (12). As an example, numerous nanosilver solutions are commercially available

101 for the disinfection of dental unit waterlines. The dental lines are routinely flushed with a nanosilver solution for a duration of time, rinsed, and brought back into service. Although this practice has gained popularity, little research has been conducted assessing the physiochemical transformations of the AgNPs during the disinfection process, and more importantly, upon exiting the system. Therefore, beyond assessing the effectiveness of AgNPs as a disinfectant, it is important to monitor their physicochemical characteristics throughout the process.

The objectives of this work are the evaluation of the potential physiochemical alterations of commercial nanosilver cleaner (ASAP-AGX-32, an antimicrobial cleaner for dental units,

0.0032% Ag) as a disinfectant on biofilm growth in DUWL. It is well established that the physicochemical properties (e.g. aggregation, complexations, and transformations) of nanosilver govern their fate, as well as, their potential environmental implications. Hence, it is our goal to demonstrate that the disinfection process commonly utilized by dental offices may impact the vital physiochemical properties of the AgNPs and therefore, the AgNPs released after the usage scenario, regardless of the disposal scenario, contain different physiochemical characteristics.

4.2 Experimental Methodology

An in-house DUWL model was built to simulate common dental unit water lines used in dental practices. A continuous flow system of tap water was utilized as shown in Figure C1. Gray polyurethane tubing with an outer diameter of 1/4"was purchased from 55 Dental Supplies and

Equipment (www.55dental.com) and used for this experiment. The model was designed to allow tap water to run through multiple loops of dental tubing and ultimately drain into a collection tank.

As water was siphoned from the tank, it was automatically replaced using a float valve, allowing unattended continuous water flow. The continuous water flow rate was set at 2 ml min-1 and a total of twelve 4 m in length dental tubing lines were utilized. After 5.5 months of continuous tap water

102 circulating through the DUWL, a substantial accumulation of biofilm was detected according to the microbial analysis. At this time, the dental tubing was transferred to the re-circulating disinfection system (Figure C2) and 2 ppm nanosilver was circulated through the system for a duration of 3 days.

Figure S2 is a schematic of the setup used for the disinfection process. The disinfection procedure was run following the manufacturer’s instructions for a duration of 3 consecutive days. Two liters of commercial nanosilver cleaner ASAP-AGX-32 (0.0032% Ag) were prepared with a final silver concentration of 2 ppm to be used for the course of the disinfection process. Initially, 1.5 L of the diluted nanosilver solution was added to bottle A and 0.5 L to bottle B. The flow rate from bottle

A to B was set at 1.3 mL s-1 for a 10-minute interval resulting in 800 mL of solution transferring from A to B. After 10 minutes, the valves were reversed and 800 mL was transferred back from B to A. This process was continued for the duration of the treatment period. As demonstrated in

Figure S2, the recirculation of the disinfectant solution was also setup to perform automatically by a pre-programed timer/control box with multiple start and stop times per day aimed at achieving the proposed operational instructions.

Subsequent to 3 days of nanosilver treatment, the dental tubing was transferred back to the continuous water flow system following the procedure used in dentist offices. Microbial analysis of the dental tubing and corresponding effluent tap water were performed immediately after the 3 days of nanosilver treatment. Intended for microbial analysis, 1 ft dental tubing was cut into 6 sections and each section was sliced in half. The biofilm was removed from the sectioned pieces of the tubing by means of either physical scraping or sonication. Biofilm collection by scraping was done by scooping the biofilm using disposable inoculating loops. As for sonication, the sliced tubing sections were immersed in centrifuge tubes containing 30 mL sterile water and sonicated

103 for an interval of 10 minutes (Fisher Scientific, Ultrasonic Cleaner, FS30, 100w, 40 kHz).

Sonication was proven to be a more effective way for removal and or re-suspension of biofilms and bacteria in sterile water (13-15). The water samples collected directly from the dental lines

(effluent) and the water samples collected subsequent to the sonication process were filtered through 0.4 µm polycarbonate filters. Subsequently, the water samples were plated onto R2A agar in 10, 100 and 1000 fold dilutions and incubated for one week at 25 ˚C on duplicate plates. The

R2A plates were evaluated for colonies after one week. Colonies were counted and recorded as colony forming units, CFU ml-1.

Transmission electron microscopy (TEM, JEOL JSM 2100), Scanning electron microscopy (SEM,

JEOL JSM 6490LV and JEOL JSM 7600F) and Energy-dispersive X-ray analysis (EDX, Oxford

Isis) were used for characterization of the nanosilver and biofilm, respectively. For TEM analysis, samples were prepared by depositing a drop of nanoparticle suspension onto a carbon coated copper grid and air-drying the samples at room temperature in a dust-free box overnight. For SEM characterization, the dental tubing samples were cut and placed in 2.5% gluteraldehyde in 0.1M cacodylate buffer. Afterwards, samples were placed in 1% OsO4 and washed with double distilled water, dehydrated in a dilution series of ethanol-water solution and placed in a desiccator to be air- dried. Subsequently, samples were mounted and gold coated for examination with SEM. Finally, an elemental composition analysis of the silver nanoparticles by energy-dispersive X-ray spectroscope (EDX, Oxford Isis) was performed to confirm the elemental presence of silver in the electron micrographs.

Prior to the initialization of the experiment, intended for XPS analysis, evaporated (concentrated) pristine Ag nanoparticles were suspended in distilled water and subsequently evaporated in droplets onto the sticky carbon tape and placed on the platen. Surface elemental speciation and

104 composition of the pristine silver nanoparticles was performed using X-ray photoelectron spectroscopy (XPS, Scanning Al-Kα source). The binding energy was set in the region of 0 to

1450eV and the resolution < 0.5 eV using beam width resolution of 10-100μm. The analysis took place in High Power mode with neutralization (anti-charging) by electrons and argon ions. Further details of the XPS analysis can be found in the SI.

To evaluate transformations in silver speciation that may have occurred in the course of the disinfection process, X-ray absorption spectroscopy (XAS) was conducted at Sector 10-ID (16) of the Advanced Photon Source of Argonne National Laboratory (ANL), Argonne, IL following the experimental setup and sample analysis described by Lombi et al. (17). Upon completion of the disinfection process, the used nanosilver solution was collected and stored in a dark serum bottle to circumvent changes in speciation. In order to reach detectable silver concentrations, a 50 ml aliquot of the spent nanosilver solution was filtered through a 0.45µm polycarbonate membrane, and the membrane was dried overnight in an anaerobic chamber. The AgNPs accumulated on the membrane were examined by XAS utilizing a quick scan set up.

4.3 Results and Discussions

4.3.1 Microbial Analysis

Table 4.1 shows the bacterial growth in CFU ml-1 for the dental tubing, as well as, corresponding effluent water samples after 2 weeks, 1 month, 10 weeks, 3.5 months and 5.5 months of continuous tap water flow. According to the results, there were no detectable bacterial developments on the inner surface of the dental tubing nor the effluent during the initial 2 weeks. Microbial analysis of triplicate tap water samples indicated the CFU count negligible, theoretically explaining the lack of significant biofilm formations in the first few weeks. The first indications of bacterial growth came after one month from the effluent water samples. However, at this time point the dental

105 tubing did not measure detectable bacterial growth. Although it has been reported the CFU levels can reach magnitudes of 105 after one or two weeks (18, 19), results from other experiments show a lower CFU level and biofilm growth comparable and similar to ours (20, 21). Variations in the

CFU measure, as well as, biofilm density may be caused by dissimilar planktonic bacterial levels present depending on the water quality. As demonstrated in Table 1, the bacterial growth increases over time and after 5.5 months of tap water flowing through the lines, the CFU is in the order of

102 for the effluent water and 102 – 103 for the tubing.

Table 4.2 shows the data pertaining to the microbial analysis (CFU/ml) of both the dental tubing surface and effluent water before and after the nanosilver disinfection process. As demonstrated, upon treatment no CFU were detected on the dental tubing surface and the CFU count in the effluent water had decreased meeting the level recommended for acceptable water quality (< 200

CFU/ml). Nevertheless, it must be emphasized that the CFU count is presented as general confirmation of the occurrence of the disinfection process, while the primary goal is to evaluate the potential physiochemical alterations occurring to the AgNPs.

4.3.2 Scanning Electron Microscopy (SEM)

Figure 4.1 shows SEM images of the inner surface of pristine dental tubing. Dust-like material and particles can be observed on the large-view image Figure 1A. The soft texture of the polyurethane surface can be clearly seen in Figure 4.1B. The polyurethane inner surface of dental tubing is a favorable substrate for microbial attachment and growth (5). Figure 4.2A and 4.2B display the

SEM images of the inner surface of the tubing after 19 days and 4 months of continuous tap water flow, respectively. In Figure 4.2A, the biofilm is covered by extracellular polymeric substance

(EPS), a slime-like matrix that gives biofilm stability and helps it adhere to the tubing surface (22).

The EPS development at this stage is very thin, however, several rod shaped bacteria can be

106 detected in the biofilm development. The biofilm structure may still be in its early stages of development at this point. Figure 4.2B demonstrates the SEM of the inner surface of the tubing after 4 months of continuous tap water flow. Relative progression of the biofilm growth is evident and an increase in the bacteria population is noticeable within the EPS compared to Figure 4.2A.

However, an overview SEM image of the same area of the tubing indicates the biofilm growth to be sporadic with low density at this time point. This may potentially explain the relatively low

CFU counts encountered in Table 1.

Figures 4.3A and 4.3B display the backscattered and secondary electron SEM of the same area of the biofilm surface after treatment with nanosilver. As a result, the particle spatial distribution is consistent on both SEM images. Distinct bright particles are observed on the surface of the biofilm in Figure 4.3A. In view of the fact that backscattered SEM imaging is capable of displaying distinguishable contrasts amongst different materials, it is our opinion that the bright distinctive particles visible in Figure 4.3A are silver in the form of particles or aggregates. This hypothesis was further established by detection of a large particle on the biofilm surface using high magnification backscattered SEM imaging in Figure C3.

4.3.3 Transmission Electron Microscopy (TEM) and Energy-dispersive X-ray (EDX)

Analysis

The TEM image of the pristine commercial ASAP-AGX-32 nanosilver is shown in Figure 4.4A.

The pristine AgNPs are small in size, ranging between 3–5 nm in diameter, spherically shaped and surrounded by a polymer. The corresponding EDX spectrum however, distributes a weak Ag signal as can be seen in Figure 4.4B. The weak EDX signal is perceived to be related to the small size of the particles and the existence of a surrounding polymer. The utilization of a polymer is primarily known to be for stabilization of AgNPs in aqueous solutions (23). Although the composition of

107 the specific polymer used in this commercial product is not specified by the vendor, TEM images of comparable polymers and particles surrounded by polymers have been previously reported (24-

27). As an example, Brown et al. presents, time-resolved, in situ TEM imagery, demonstrating real-time alterations to the polymer on the TEM grid (24). El-Shall et al. presents TEM images of metallic and bimetallic particles bound and stabilized by polymers (25). All of the reported TEM imaging, demonstrate a polymer morphology similar to that depicted in Figure 4A. Furthermore, the existence of a strong carbon peak confirms the presence of s stabilizing agent in the prepared

EDX samples, Figure 4.4B.

To monitor the physical transformations AgNPs may potentially encounter during the course of the disinfection process, further TEM imaging was performed. Figures 4.5A and 4.5C display

TEM images of the AgNPs subsequent to recirculation through the dental tubing for 3 days

(disinfection process). It is evident that the pre-existing polymer encapsulating the silver nanoparticles are no longer present. The polymer may have adsorbed onto the surface of the dental tubing or biofilm during the treatment period. The corresponding EDX spectra displayed in Figure

4.5B (after disinfection process) is focused on a non-aggregated portion of the AgNPs and demonstrates an Ag signal similar to that of Figure 4.4B (pristine). Although the EDX analysis was performed using similar spot size, it is probable that different locations may contain different numbers of particles leading to an altered Ag signal strength. The absence of the polymer as a stabilizer, appears to have initiated the aggregation of a proportion of the AgNPs, as shown in

Figure 4.5C. Although a fraction of the AgNPs have lost their stability in suspension, forming larger aggregated particles, small particles in the 3-5 nm range are still dominant in number. Figure

4.5D is the EDX spectrum corresponding to the aggregate particles (Figure 4.5C). This spectrum displays a considerably strong Ag signal which is attributed to the accumulation of particles

108 forming the aggregate. The observed aggregated Ag particles measured between 50–200 nm in diameter. This considerable increase in particle size observed in Figure 4.5D, impacts the physiochemical properties of the AgNPs, giving rise to different environmental impacts and transportation scenarios (after disposal) compared to the pristine AgNPs (28). As an example, aggregated silver nanoparticles are far less stable compared to the pristine nanoparticles’ containing the polymer stabilizer and consequently, more likely to precipitate out of solution. In addition, particle aggregation causes a decrease in the surface/volume, lowering their surface of contact (reactivity) and therefore lowering their antimicrobial capabilities (29-32).

4.3.4 Silver speciation by X-ray Photoelectron Spectroscopy (XPS) and X-ray Absorption

Spectroscopy (XAS)

To determine the chemical speciation state of the particle surface atoms for the pristine nanosilver solution, XPS measurements at the Ag 3d core levels were conducted. Previous studies have demonstrated the Ag3d5/2 binding energies for Ag, Ag2O and AgO are approximately 368.2, 367.8 and 367.4 eV, respectively (33, 34). The Ag 3d5/2-3/2 spectrum is shown in Figure 4.6, with binding energies of Ag3d doublet peaks located at 367.7 (Ag3d5/2) and 373.7 (Ag3d3/2) eV . Comparison of these peaks with the 3d peaks available from Ag metal (BE = 368.0 eV) reveals some major differences. The binding energy of Ag3d core levels for the pristine nanosilver solution shifts towards lower binding energy values signifying the existence of silver oxidation states. Therefore, according to the peak shifts of the Ag3d lines observed between both species (i.e. an Ag 3d5/2 peak

3+ at 367.05 ± 0.05 eV and a 3d3/2 peak at 373.25 ± 0.05 eV), the presence of Ag ions is likely.

According to the literature and our experiments, the binding energy of the 3d5/2 level is equal to

+1 368.3 ± 0.1 eV for silver metal, and that of Ag2O (i.e. Ag ) is about 367.9 ± 0.1 eV, therefore, a shift of about 0.3 eV per valence unit to lower energies can be expected. Thus, the measured

109 value of 367.3 eV measured for this peak is well in accordance with the presence of Ag3+ ions in the oxide layer (35, 36). Further analysis of the spectrum indicates the pristine nanosilver solution to be comprised of AgO (~74%) with a small admixture of Ag (I) and Ag (III) (~26%).

Previous studies indicate transformation of AgNPs to AgCl to be anticipated when exposed to chloride containing environments (37). The XAS spectra from the analysis of the nanosilver solution after the disinfection process is demonstrated in Figure 4.7A. A visual inspection of Figure

4.7A clearly demonstrates AgCl as the dominant species developed following the disinfection process. In fact, analysis of the LCF results (Figure 4.7B) indicates that the dominant species formed were AgCl. As tap water contains fairly high levels of chlorine, the formation of AgCl was triggered and anticipated. Therefore, under the current conditions present in the DUWL disinfection system, phase transformations of AgNPs are likely. Phase transformations are of importance as they may impact the stability, bioavailability and toxicity of AgNPs.

4.4 Conclusion

In summary, this research highlights the physiochemical alterations, as well as fate of the commercial AgNPs, during the disinfection process. The analysis of colony forming units

(CFU/ml) for both the effluent water and dental tubing, before and after treatment, reveal the commercial nanosilver solution to be an effective disinfecting agent (<200 CFU/ml).

Characterization of the commercial nanosilver revealed the pristine AgNPs to be spherical in shape, ranging between 3 – 5 nm in size, and bound by a stabilizing polymer. However, after disinfection, the polymeric stabilizing agent surrounding the AgNPs were no longer visible. The disappearance of the polymer was most likely due to adsorption onto the biofilm surface. As a result of the absence of the capping agent, aggregation of a fraction of the silver nanoparticles were initiated. The AgNP aggregates size distribution ranged between 50 – 200 nm in diameter.

110

Additionally, surface transformations of the silver nanoparticles were observed after the course of the disinfection process (treatment process). The pristine nanoparticles demonstrated through XPS analysis to be dominantly silver oxides (AgO) with a miniscule portion of Ag (I) and Ag (III).

While, the spent nanosilver solution displayed substantial surface transformations of the AgNPs to AgCl.

To the best of our knowledge, this manuscript demonstrates the first time Ag nanoparticles have been demonstrated to adsorb onto biofilm surfaces and therefore will assist in further illustration of the toxicity mechanisms of AgNPs to bacteria and biofilms. This work can be an initial step in better understanding how nanomaterials, specifically AgNPs, transform depending on the conditions they are exposed to during their usage scenario. Till this date, most research has been focused on assessing the impacts of pristine nanomaterials on various systems. Where it is our belief that the nanoparticles may go through transformations when utilized and these transformations must be taken into consideration prior to making judgements in regards to their environmental toxicity (natural or engineered systems).

111

Literature Cited

1. Hajipour, M. J.; Fromm, K. M.; Ashkarran, A. A.; de Aberasturi, D. J.; de Larramendi, I. R.; Rojo, T.; Serpooshan, V.; Parak, W. J.; Mahmoudi, M. Antibacterial properties of nanoparticles. Trends Biotechnol. 2012, 30, (10), 499-511.

2. Tolaymat, T. M.; El Badawy, A. M.; Genaidy, A.; Scheckel, K. G.; Luxton, T. P.; Suidan, M. An evidence-based environmental perspective of manufactured silver nanoparticle in syntheses and applications: A systematic review and appraisal of peer-reviewed scientific papers. Sci. Total Environ. 2010, 408, 999-1006.

3. Donoso-Bravo, A.; Mailier, J.; Martin, C.; Rodriguez, J.; Aceves-Lara, C. A.; Wouwer, A. V. Model selection, identification and validation in anaerobic digestion: a review. Water Res. 2011, 45, 5347-5364.

4. Costa, D.; Girardot, M.; Bertaux, J.; Verdon, J.; Imbert, C. Efficacy of dental unit waterlines disinfectants on polymicrobial biofilm. Water Res. 2016, 91, 38-44.

5. Walker, J. T.; Marsh, P. D. Microbial biofilm formation in DUWS and their control using disinfectants. J. Dent. 2007, 35, (9), 721-730.

6. Szymanska, J. Biofilm and dental unit waterlines. Ann. Agr. Env. Med. 2003, 10, (2), 151-157.

7. Pankhurst, C. L.; Coulter, W.; Philpott-Howard, J. N.; Surman-Lee, S.; Warburton, F.; Challacombe, S. Evaluation of the potential risk of occupational asthma in dentists exposed to contaminated dental unit waterlines. Primary Dental Care: Journal of the Faculty of General Dental Practitioners (UK) 2005, 12, (2), 53-9.

8. Shepherd, P. A.; Shojaei, M. A.; Eleazer, P. D.; Van Stewart, A.; Staat, R. H. Clearance of biofilms from dental unit waterlines through the use of hydroperoxide ion-phase transfer catalysts. Quintessence Int. 2001, 32, 755-761.

9. Blake, G. C. The incidence and control of bacterial infection of dental units and ultrasonic scalers. Brit Med J. 1963, 115, 413-416.

10. Pankhurst, C. L.; Johnson, N. W.; Woods, R. G. Microbial contamination of dental unit waterlines: the scientific argument. Int Dent J. 1998, 48, (4), 359-68.

11. Shearer, B. G. Biofilm and the dental office. J Am Dent Assoc. 1996, 127, (2), 181-9.

12. Correa, J. M.; Mori, M.; Sanches, H. L.; da Cruz, A. D.; Poiate Jr., E.; Poiate, I. A. Silver nanoparticles in dental biomaterials. Int J Biomater. 2015, 485275.

13. Manuel, C. M.; Nunes, O. C.; Melo, L. F. Dynamics of drinking water biofilm in flow/non- flow conditions. Water Res. 2007, 41, (3), 551-562.

112

14. Gagnon, G. A.; Slawson, R. M. An efficient biofilm removal method for bacterial cells exposed to drinking water. J. Microbiol. Methods. 1999, 34, (3), 203-214.

15. Jost, G. F.; Wasner, M.; Taub, E.; Walti, L.; Mariani, L.; Trampuz, A. Sonication of catheter tips for improved detection of microorganisms on external ventricular drains and ventriculo- peritoneal shunts. J. Clin. Neurosci. 2014, 21, (4), 578-582.

16. The MRCAT Insertion Device Beamline at the Advanced Photon Source", C.U. Segre, N.E. Leyarovska, L.D. Chapman, W.M. Lavender, P.W. Plag, A.S. King, A.J. Kropf, B.A. Bunker, K.M. Kemner, P. Dutta, R.S. Duran and J. Kaduk, CP521, Synchrotron Radiation Instrumentation: Eleventh U.S. National Conference, ed. P. Pianetta, et al., p419-422, (American Insitute of Physics, New York, 2000).

17. Lombi, E.; Donner, E.; Taheri, S.; Tavakkoli, E.; Jamting, A. K.; McClure, S.; Naidu, R.; Miller, B. W.; Scheckel, K. G.; Vasilev, K. Transformation of four silver/silver chloride nanoparticles during anaerobic treatment of wastewater and post-processing of sewage sludge. Environ. Pollut. 2013, 176, 193-197.

18. Walker, J. T.; Bradshaw, D. J.; Fulford, M. R.; Marsh, P. D. Microbiological evaluation of a range of disinfectant products to control mixed-species biofilm contamination in a laboratory model of a dental unit water system. Appl. Environ. Microbiol. 2003, 69, (6), 3327-3332.

19. Porteous, N.; Sun, Y. Y.; Dang, S. C.; Schoolfield, J. A comparison of 2 laboratory methods to test dental unit waterline water quality. Diagn. Microbiol. Infect. Dis. 2013, 77, (3), 206- 208.

20. Schel, A. J.; Marsh, P. D.; Bradshaw, D. J.; Finney, M.; Fulford, A. R.; Frandsen, E.; Ostergaard, E.; ten Cate, J. M.; Moorer, W. R.; Mavridou, A.; Kamma, J. J.; Mandilara, G.; Stosser, L.; Kneist, S.; Araujo, R.; Contreras, N.; Goroncy-Bermes, P.; O'Mullane, D.; Burke, F.; O'Reilly, P.; Hourigan, G.; O'Sullivan, M.; Holman, R.; Walker, J. T. Comparison of the efficacies of disinfectants to control microbial contamination in dental unit water systems in general dental practices across the European union. Appl. Environ. Microbiol. 2006, 72, (2), 1380-1387.

21. Porteous, N.; Luo, J.; Hererra, M.; Schoolfield, J.; Sun, Y. Y. Growth and identification of bacterial in N-halamine dental unit waterline tubing using an ultrapure water source. Int. J. Microb. 2011, 767314.

22. Hall-Stoodley, L.; Costerton, J. W.; Stoodley, P. Bacterial biofilms: From the natural environment to infectious diseases. Nat. Rev. Microbiol. 2004, 2, (2), 95-108.

23. Iravani, S.; Korbekandi, H.; Mirmohammadi, S. V.; Zolfaghari, B. Synthesis of silver nanoparticles: chemical, physical and biological methods. Res. Pharm. Sci. 2014, 9, (6), 385- 406.

113

24. Brown, R. M.; Barnes, Z.; Sawatari, C.; Kondo, T. Polymer manipulation and nanofabrication in real time using transmission electron microscopy. Biomacromolecules 2007, 8, (1), 70-76.

25. El-Shall, M. S.; Abdelsayed, V.; Khder, A.; Hassan, H. M. A.; El-Kaderi, H. M.; Reich, T. E. Metallic and bimetallic nanocatalysts incorporated into highly porous coordination polymer MIL-101. J. Mater. Chem. 2009, 19, (41), 7625-7631.

26. Liu, J. F.; Liu, R.; Yin, Y. G.; Jiang, G. B. Triton X-114 based cloud point extraction: a thermoreversible approach for separation/concentration and dispersion of nanomaterials in the aqueous phase. Chem. Commun. 2009, (12), 1514-1516.

27. Xing, S. X.; Tan, L. H.; Yang, M. X.; Pan, M.; Lv, Y. B.; Tang, Q. H.; Yang, Y. H.; Chen, H. Y. Highly controlled core/shell structures: tunable conductive polymer shells on gold nanoparticles and nanochains. J. Mater. Chem. 2009, 19, (20), 3286-3291.

28. Agnihotri, S.; Mukherji, S.; Mukherji, S. Size-controlled silver nanoparticles synthesized over the range 5-100 nm using the same protocol and their antibacterial efficacy. RSC. Adv. 2014, 4, 3974-3983.

29. Dorjnamjin, D.; Ariunaa, M.; Shim, Y. K. Synthesis of silver nanoparticles using hydroxyl functionalized ionic liquids and their antimicrobial activity. Int. J. Mol. Sci. 2008, 9, (5), 807– 819.

30. Kvitek, L.; Panacek, A.; Soukupova, J.; Kolar, M.; Vecerova, R.; Prucek, R.; Holecova, M.; Zboril, R. Effect of surfactants and polymers on stability and antibacterial activity of silver nanoparticles (NPs). J. Phys. Chem. 2008, 112, (15), 5825– 5834.

31. Yu, D. G. Formation of colloidal silver nanoparticles stabilized by Na+-poly(gamma-glutamic acid)-silver nitrate complex via chemical reduction process. Colloids Surf. 2007, 59 (2), 171– 178.

32. Li, X.; Lenhart, J. J. Aggregation and dissolution of silver nanoparticles in natural surface water. Environ. Sci. Technol. 2012, 46, (10), 5378-5386.

33. Lai, Y.; Zhuang, H.; Xie, K.; Gong, D.; Tang, Y.; Sun, L.; Lin, C.; Chen, Z. Fabrication of uniform Ag/TiO2 nanotube array structures with enhanced photoelectrochemical performance. New. J. Chem. 2010, 34, 1335-1340.

34. Prieto, P.; Nistor, V.; Nouneh, K.; Oyama, M.; Abd-Lefdil, M.; Diaz, R. XPS study of silver, nickel and bimetallic silver-nickel nanoparticles prepared by seed-mediated growth. Appl. Surf. Sci. 2012, 258, 8807-8813.

35. Lutzenkirchen-Hecht, D.; Strehblow, H. Anodic silver (II) oxides investigated by combined electrochemistry, ex situ XPS and in situ X-ray absorption spectroscopy. Surf. Interface Anal. 2009, 41, 820-829.

114

36. Ferraria, A.; Carapeto, A.; Botelho do Rego, A. X-ray photoelectron spectroscopy: Silver salts revisited. Vacuum. 2012, 86, 1988-1991.

37. Gitipour, A.; ElBadawy, A.; Arambewela, M.; Miller, B.; Scheckel, K.; Elk, M.; Ryu, H.; Gomez, V. A.; Santo Domingo, J.; Thiel, S.; Tolaymat, T. The impact of silver nanoparticles on the composting of municipal solid waste. Environ. Sci. Technol. 2013, 47, 14385-14393.

115

Table 4.1 Microbial analysis of dental tubing and corresponding effluent water (CFU/ml) after 2 weeks, 1 month, 10 weeks, 3.5 months and 5.5 months of continuous tap water flow.

Sample type 2 weeks 1 month 10 weeks 3.5 months 5.5 months (CFU/ml) (CFU/ml) (CFU/ml) (CFU/ml) (CFU/ml) 1 Tubing < 10 < 10 < 10 9.50 × 101 2.64 × 103

2 Tubing < 10 < 10 < 10 2.50 × 102 1.90 × 103

3 Tubing < 10 < 10 < 10 9.65 × 102 7.00 × 102

1 Effluent water < 10 4.00 × 101 < 10 3.40 × 102 1.89 × 103

2 Effluent water < 10 4.70 × 102 2.70 × 102 6.40 × 102 1.70 × 102

3 Effluent water < 10 2.60 × 102 2.70 × 102 6.00 × 102 1.45 × 102

116

Table 4.2 Microbial analysis results of the dental tubing and corresponding effluent water before and after the disinfection process (re-circulation with 2 ppm nanosilver for 3 days).

Sample type Prior to treatment After nanosilver treatment

(CFU/ml) (CFU/ml)

1 Tubing 2.64 × 103 < 10

2 Tubing 1.90 × 103 < 10

3 Tubing 7.00 × 102 < 10

1 Effluent water 1.89 × 103 2.50 × 102

2 Effluent water 1.70 × 102 1.70 × 102

3 Effluent water 1.45 × 102 6.00 × 101

117

Figure 4.1 SEM images of unused dental tubing surface. Images A) and B) comprise of the large view and detailed texture of tubing surface, respectively.

118

Figure 4.2 SEM image of the biofilm on dental tubing surface after 19 days (A) and 4 months (B) of continuous tap water flow.

119

120

Figure 4.4 TEM image of original commercial ASAP-AGX-32 nanosilver (A) and corresponding EDX spectra (B).

121

122

Figure 4.5 TEM images (A and C) of nanosilver after re-circulation through dental tubing for 3 days. Image A shows some small 3 – 5 nm Ag nanoparticles and C shows a large partially- aggregated and Ag particle which is surrounded by some small 3 – 5 nm Ag particles. B and D are corresponding EDX of the particles in image A and C, respectively.

123

Figure 4.6 Ag 3d XPS spectra of pristine Ag nanoparticles.

124

A) 2

Ag2O

AgCl

Spent Nanosilver

Ag Metal Normalized Absorption Normalized

0 25480 25500 25520 25540 25560 25580 25600 25620 25640 Energy (eV)

Sample AgCl (%) Ag0 (%) R-factor (x1000) Spent Nanosilver 56 44 0.041

Figure 4.7 A) Ag-Kα XAS spectra of Ag2S, Ag2O, AgCl and Ag metallic as pure phases and AgNPs (ASAP-AGX-32) after the DUWL disinfection process. B) Ag speciation as identified by LCF of k-edge XANES spectra. Silver species proportions are presented as percentages. Goodness of fit is indicated by the R-factor.

125

APPENDIX C

126

C1. XPS Analysis

X-ray photoelectron spectroscopy (XPS) was performed using a QUANTERA II Spectrometer

(PHI, USA). The samples were prepared and mounted for analysis using standard practices consistent with high vacuum surface analytical procedures. Powders were placed onto the XPS compatible mounts (using nonmagnetic screws and fixtures) prior to analysis. Calibration was performed using a clean Au/Cu sample and a pure Ag sample (99.99%). Measured values for electron binding energies (BE) were 84±0.02 eV and 932±0.04 eV. The samples were irradiated with monochromatic AlKα X-rays (hν = 1486.6eV) using an X-ray spot size of 400×700 μm2.

Surface charging was compensated by means of a filament (I = 1.9A, 3.6 V) inserted in a magnetic lens system, and all spectra were corrected by setting the C1s hydrocarbon component to

284.60eV. For each sample, a survey spectra (0-1200 eV) were recorded at a pass energy of 20 and 160 eV to determine the surface chemical compositions as percentage. The data were processed using a PHI MultiPak data reduction software (Physical Electronics, USA). Sample compositions were obtained from the survey spectra after linear background subtraction using RSF

(Relative Sensitivity Factors) derived from Scofield cross-sections. Curve fitting was carried out using the same initial parameters and inter-peak constraints to reduce scattering. The core level envelopes were fitted with Gaussian-Lorentzian function (G/L=30) and variable full width half maximum.

127

Figure C.1 Schematic of the Dental Unit Water Line (DUWL)-Continuous tap water flow system.

128

Figure C.2 Schematic of the disinfection process through circulation of the nanosilver solution through the dental tubing. (Re-circulation system).

129

Figure C.3 High magnification of backscattered SEM image of a silver nanoparticle attached to the biofilm surface.

130

Figure C.4 TEM images of the pristine silver nanoparticle solution (ASAP-AGX-32).

131

Figure C.5 SEM images of the inner surface of dental tubing water lines (DUWL) after 19 days of continuous tap water flow.

132

Figure C.6 SEM images of the inner surface of dental tubing water lines (DUWL) after 4 months of continuous tap water flow.

133

Figure C.7 TEM images of 2 ppm nanosilver solution (ASAP-AGX-32) after the dental tubing treatment (disinfection process) at various magnification levels A) 50k B) 200k C) 400k.

134

Chapter 5

Fate and Transportation of Silver Nanoparticles Released from Consumer Products: Ecological Risk Assessments

This chapter is to be submitted for publication as

Gitipour, A.; Koralegedara, N.; Al-Abed. S. R.; Thiel, S. W. Fate and transportation of silver nanoparticles released from consumer products: ecological risk assessments. Journal of Hazardous Materials. 2016, in preparation.

135

Abstract

The extensive growth in engineered nanomaterials (ENMs) applications in the last few decades has led to their incorporation in a wide range of commercially-available consumer products (CPs).

AgNPs are the most widely-used ENMs in consumer products, mainly due to their antibacterial properties. This widespread usage has resulted in concerns regarding potential adverse environmental impacts. In an effort to better understand the fate, transport and environmental impacts of AgNPs, many studies have been conducted in the past decade. However, this research has been focused mainly on pristine AgNPs either lab-synthesized or purchased commercially.

Although these studies have contributed greatly to a better fundamental understanding of the mechanisms of AgNP toxicity and the behavior of AgNPs under various environmental conditions, they have not addressed the environmental concerns associated with the release of AgNPs from consumer products. AgNPs are seldom used as pure materials; they are mainly integrated into or embedded onto the surface of commercial product.

This review assesses previously reported work on characterization techniques, routes of environmental exposure, potential ecological risks of AgNPs, as well as on the identification of available toxicity assays that may be useful in toxicity assessments on AgNPs in consumer products. The main objectives of this review are to summarize the major findings of past research related to AgNPs released from CPs, to identify key issues and knowledge gaps and to provide potential directions for future research.

136

5.1 Introduction

Silver is well-known for its antimicrobial activity (1). Silver nanoparticles (AgNPs) are fine particles of metallic silver with a size less than 100 nm in at least one dimension. AgNPs may exhibit increased antimicrobial activity compared to bulk-Ag due to their large surface area (2).

Therefore, the applications and exploitations of AgNPs have witnessed a dramatic increase over the past decades. According to the nanotechnology consumer product inventory (CPI), AgNPs are currently the most widely used nanoparticle (5); a total of 442 AgNP-enabled consumer products were listed in the CPI as of 2016 (Figure 5.1) (3). These products include textiles, personal care products, medical equipment, food packages, plastic and silver ware, cook wares, antimicrobial agents, toys, baby products, paints, inks, air and water filters, vacuum cleaners, humidifiers, washing machines, electronics, refrigerators, dietary supplements and medicines (3, 4).

Although significant attention has been given to the production of AgNP-containing CPs, their potential health and ecological risks are not yet fully known. To date, extensive research has been conducted on the synthesis and characterization of AgNPs (6-8), their physicochemical behaviors under different conditions (9-13) and their toxicity to specific microorganisms (14-16). However, all of these studies have investigated pristine (laboratory-synthesized or commercially available)

AgNPs. This is of concern as pristine AgNPs may lack characteristics associated with AgNP-CPs incorporation process, as well as possible alterations to the physicochemical characteristics that may occur during the CPs lifecycle. Consequently, pristine AgNPs may not accurately represent

AgNPs likely to be released into the environment.

The behavioral differences between pristine AgNPs and AgNPs released from CPs are not associated with the basic functionality of the AgNPs but rather with the release form, concentration and associated matrix. As an example, AgNPs incorporated in paints are released with the organic

137 binders used in the paint, altering the characteristics and behavior of these AgNPs. Surface-coated

AgNPs are reportedly more easily released from textiles than are fiber-coated AgNPs (17).

Moreover, the concentrations of AgNPs that are integrated into and therefore potentially released from different CPs depend on the manufacturing process and may differ among applications. The wide variety of applications has led to a wide variety of usage scenarios that also impact the characteristics of AgNPs. For instance, AgNPs used in textiles may be transformed into different

Ag-species due to the interactions with the laundry detergents (18). These observations highlight the importance of investigating the use of AgNP-containing CPs in environmental exposure and health risk assessments. Only a few of the commercially-available AgNP-containing CPs have been evaluated. Therefore, the main objectives of this review are to summarize the major findings of past research related to AgNPs released from CPs, to identify key issues and knowledge gaps and to provide potential directions for future research.

5.2 Characterization of AgNPs in Consumer Products

According to recommendations made by the U.S. EPA (19), and data from the literature, the following characteristics of AgNPs are important for assessing the potential health and ecological impacts:

 Size, including clustering tendencies

 Concentration

 Morphology, including shape and crystal structure

 Surface charge

 Chemical composition

 Surface chemistry (coating) and reactivity

 Solubility

138

 Conductivity, magnetic and optical properties

5.2.1 Sample Collection and Characterization Techniques

The characterization and sample collection of pristine AgNPs is rather different from that of

AgNP-CPs. The intricate nature of the AgNPs associated in CPs as well as their form of release cause difficulties in reproducible sampling and reliable analysis. The complex nature and composition of environmental matrices (air, natural water sources, wastewater, sludge and soil) further complicate sample collection and characterization. In the summary below, sample collection and characterization techniques presented in the literature will be discussed based on their physical state of the AgNP-CPs.

5.2.1.1 Solid Materials

AgNP containing solids are classified into two types:

 Solid AgNP-containing CPs

 Soils, sediments, solid wastes (biosolids) in which AgNPs released from CPs may have

accumulated

The physical characterization of AgNPs in solid CPs is considerably easier than that of environmental samples, as the latter may require sample separation techniques. Nevertheless,

AgNP-CPs also need pre-treatment prior to chemical analyses.

Samples assayed for Ag content are usually digested prior to analysis with inductively coupled plasma mass spectrometry (ICP-MS). The most common acid digestion procedures use HNO3 (20), a HNO3-HCl mixture or a HNO3-H2O2 mixture (21-23). In certain instances, when the material is not digested completely by acid digestion, X-ray Fluorescence (XRF) analysis can be used to determine the Ag content. However, sample pre-treatment is required for this analysis as well. As

139 reported by Lorenz et al., textile samples were incinerated at 600oC in a muffle furnace prior to creating the pellets for XRF analysis (22).

One of the most utilized methods for AgNPs characterization is using Electron Microscopy techniques such as Transmission Electron Microscopy (TEM) and Scanning Electron Microscopy

(SEM)). In these methods, a representative portion of the sample is paced on a TEM/SEM grid for imaging. This technique is effective for observing the AgNPs on the surface of the material but is not applicable for identification and characterization of AgNPs embedded in CPs. For particles embedded in solid matrices, thermal ashing has been used to break down the solid matrix, allowing more precise identification of AgNPs (24-26). However, morphological and physicochemical changes may occur during the ashing procedure can render subsequent measurement of size, shape, and aggregations unrepresentative of the AgNPs present in the original CPs (26).

X-ray Absorption Spectroscopy (XAS) is the most sensitive and sophisticated method for the identification of AgNPs speciation in CPs or environmentally relevant matrices compared to other speciation analyses techniques such as X-ray Photoelectron Spectroscopy (XPS) and X-ray

Diffraction (XRD) (27). Several studies have reported the use of XAS for AgNP characterization in solid matrices. However, some difficulties have been reported in differentiation of the Ag species of interest due to similarities in X-ray near edge spectroscopy (XANES) (18, 27).

Atomic force microscopy analysis (AFM) is also used for identification and characterization of

AgNPs on the surface of materials and has been utilized to assess the size distribution of AgNPs in textiles (28).

140

5.2.1.2 Aqueous Materials

A number of instruments and techniques are utilized for separation and characterization purposes of AgNPs in aqueous (liquid) phase materials. These techniques have been employed in the past to characterize AgNPs in CPs such as antimicrobial solutions, disinfectant sprays, cosmetic products (antiperspirants, deodorants, and shampoo), paints and inks.

Particle size distributions (PSD) of AgNPs in liquid matrices are mainly measured by dynamic light scattering (DLS). These measurements have been reported in a number of studies using a

Zetasizer (29-31).

Field flow fractionation (FFF), sedimentation field flow fractionation (SdFFF) and centrifugal field flow fractionation (centrifugal-FFF) are commonly used to separate AgNPs from liquid matrices based on their size. Further characterization of AgNPs could be obtained by coupling the separation techniques with other analytical instruments (e. g., ICP-MS, UV-visible light spectrometer (UV-vis), TEM, SEM and AFM). As an example, Cascio et al., characterized AgNPs in commercial liquids by coupling FFF and UV-vis analysis (FFF-UV-vis) (29, 32). They noted that sample heterogeneity in terms of size and shape may alter the surface plasmon resonance

(SPR) band position resulting in inaccurate results. Since the presence of various constituents in the AgNP suspension affects the UV-vis spectra, this method is only reliable for identification of

AgNPs in liquids with known background components.

Basic filtration techniques (using filters of different sizes) have been used to separate AgNPs from environmental water samples (20, 23, 24, 34). However, most of these techniques take advantage of forced separation methods capable of altering the characteristics of the AgNPs, and therefore, further characterization of the seperated AgNPs may not be an appropriate representative of the

141 original suspension. Furthermore, these filtration techniques are generally used for separation of the two states of Ag (AgNPs and Ag+) and achieving suspension concentrations.

Cloud point extraction (CPE) using Triton X-114 (TX-114) is another method used to separate

AgNPs from antimicrobial solutions and environmental water (34). This method can separate

AgNPs from the aqueous matrix without altering the size and shape of the AgNPs, as they are concentrated in the Triton X-114 phase. The AgNPs in the Triton phase are then characterized

(after microwave digestion) by either ICP-MS, TEM/energy-dispersive X-ray spectroscopy (EDX) or UV-absorption. All of the Ag+, including Ag+ in free dissolved form, Ag+ adsorbed on the

AgNPs and Ag+ associated with the matrix, remains in the aqueous phase.

The hydrodynamic particle size distribution of any nanoparticle with a refractive index different from that of the liquid medium in which it is suspended can be measured using photon correlation spectroscopy analysis. The hydrodynamic particle size distribution of AgNPs from commercially available spray solutions have been measured by combining a zetaPLS 90 with an induced particle sizing software (36).

Stuart et al., were able to detect AgNPs in colloidal silver sprays by using an electrochemical technique known as particle-impact voltammetry. This technique measures and applies the oxidation potential of AgNPs for direct detection and size measurements. Furthermore, they have reported applying this method not only for CPs but also for environmental samples (e. g., sea water) as well (37).

5.2.1.3 Aerosols

The detection and characterization of AgNPs in aerosols require special techniques and instruments. The most practical approach for studying AgNPs in aerosols is by conducting

142 experiments inside of a glove box. Sample collection of aerosols for characterization purposes is primarily done by filters using single stage impactors. An electrical low pressure impactor (ELPI) can be used to fractionate particles up to 30 nm size. Electrostatic precipitation can also be used based on the electrical mobility of particles. However, to obtain enough sample for subsequent analyses, especially for short term applications such as that of a spraying mechanism, an online technique such as scanning mobility particle sizer (SMPS) may be more appropriate. Several studies have reported measuring particle size distributions (PSD) and particle number counts of released AgNPs from aerosols using a differential mobility analyzer (DMA) followed by a condensation particle counter (CPC). This approach allows the separation of particles based on their electrical mobility followed by a particle number count of each fraction associated with

SMPS and used during spraying (22, 31, 37, 38, 39).

All of the research studies on AgNPs containing aerosols mentioned up to this point took advantage of a common method for particle characterization. Several instruments were connected to a glove box to achieve multiple analyses simultaneously. First, a thermo-desorber was used to dry the aerosol prior to parallel analysis with SMPS and electrostatic sampler. TEM grids were placed on electrostatic samplers to collect AgNPs for morphological assessment using TEM/SEM and compositional analysis using EDX.

A detailed study of the characterization of an aerosol containing AgNPs was conducted using different types of AgNP-containing sprays (anti-odor spray for hunters, disinfectant spray and throat spray) (33). In this experiment, an optical particle counter (OPC) and a diffusion charger

(DC) were used in addition to SMPS and CPC to measure concentrations and surface area of

AgNPs, respectively.

143

In most of these studies, the PSD and particle concentrations are measured concurrently with sample generation, while the chemical and morphological characteristics are assessed later. Losert et al., have considered this issue and suggested a SMPS-ICPMS coupling technique to analyze both PSD and physicochemical characteristics of the released NPs simultaneously. This technique has successfully demonstrated the capability to analyze AgNPs released from commercial spray solutions (31).

Furthermore, an innovative nano-aerosol chamber has been recently reported for in-vitro toxicity studies (40). This chamber is capable of controlling the environmental conditions (e. g., moisture and temperature) and allows for long-term exposure analyses.

5.3 Past Research on AgNP-containing Consumer Products

AgNPs are the most widely used metal nanoparticles in consumer products (3). The potential ecological and health toxicity associated with AgNPs have been shown in a number of past studies, primarily using laboratory-synthesized AgNPs. However, similar ecological or health exposure analyses for AgNP-containing CPs are limited to a small number of studies primarily focused on identification of the release concentrations, speciation and morphological changes. of AgNPs. The major findings, experimental difficulties and data gaps from previous studies on AgNP-CPs are discussed below under two main categories: Solid phase products, Liquid/aerosol phase products.

5.3.1 Solid Phase

A number of solid-phase CPs are available in the market that contain AgNPs, however, the majority of the past studies have focused on textiles such as antibacterial socks, T-shirts, trousers, bed sheets, medical clothes, medical masks and wound dressings. Textiles of different brands, fiber compositions (100% cotton, 100% polypropylene and combinations of polyamide, elastane and

144 wool), countries of origin (Germany, Japan, Switzerland, USA), Ag species (Ag+, Ag(0), AgCl), concentrations (1.5 mg/kg – 2925 mg/kg), and usage scenarios have been used in previous studies

(20-28). Textiles and products used for medical purposes typically contained more Ag than ordinary products. For example, a medical cloth and mask contained a total of 270,000 μg Ag/g product (20). However, apart from the medical textiles such as gloves, masks and wound dressings, research on other AgNP-enabled devices such as catheters, cardiac valves, and endotracheal tubes is limited. The main objective of those studies was to identify and improve the performances of antibacterial function of AgNPs (73, 74)

Among the other AgNP-containing household equipment (humidifiers, vacuum cleaners, hair dryers, washing machines, water and air filters, and refrigerators), washing machines and humidifiers are the only products for which research has been reported. These washing machines provide antibacterial capabilities during operation by releasing AgNPs into the washing solution.

Farkas et al., reported the presence of silver nanoparticles in washing machine effluent.

Furthermore, the effluent exhibited strong bactericidal effects on the natural bacterial community tested. (43).

Similar to washing machines, AgNP-emitting humidifiers are available for air sanitation.

However, humidifiers may cause both aerosol and liquid contamination in addition to potential soil contamination after disposal in landfills (20, 41).

Furthermore, food storage containers with embedded AgNPs are widely available and are of concern as they are directly exposed to the food consumed by humans. A number of studies have been conducted on commercially available AgNPs containing food storage containers (75-78).

According to these studies, the release and incorporation of the AgNPs to the food was dependent on, temperature, storage time and food chemistry. Furthermore, a higher release rate of AgNPs

145 under microwave heating conditions was demonstrated compared to conventional oven heating

(76).

5.3.2 Liquid/Aerosol phase

The most broadly studied and reported AgNP-containing CPs are antibacterial solutions. The main mode of application for these antibacterial solutions is spraying, producing AgNP aerosols. Due to difficulties in sample collection, experiments on AgNPs released in aerosols are challenging and therefore limited. Nevertheless, several studies have characterized the AgNP aerosols released from antimicrobial sprays (22, 31-33, 36, 38, 39). The release of both individual and aggregated

AgNPs from the commercial sprays was confirmed in two studies (36, 39).

Both solution characteristics and the spraying mechanism of the bottle have been reported to govern the characteristics of the released Ag-NPs (33, 36). The dispersion composition, storage mechanism, dispensing mechanism and nozzle specifications are important and must be considered when implementing rules and regulations pertaining to AgNP-containing CPs in the aerosol phase.

There have been frequent reports of the absence of NPs in consumer products claimed to contain

NPs. For instance, an experiment conducted by Chao et al. on the speciation of AgNPs in six commercially available antibacterial solutions found only three of the six products actually contained AgNPs (34). Similarly, the reverse scenario of the presence of NPs in products not claiming to contain NPs has been observed. Consequently, AgNP-CPs should be characterized prior to use in experimental studies to verify the presence of NPs regardless of the manufacturer’s information. This also highlights the complications associated with investigations on nanoparticles in CPs compared to laboratory synthesized NPs brought by the lack of information available on the characteristics, method of incorporation and other important parameters of AgNPs-CPs.

146

Incorporation of AgNPs into air and water filtration systems has also been reported. A higher possibility of AgNPs leaching from these products exists, as they are directly exposed to air and water. However, to date no studies have been reported on characterization or fate of AgNPs released from these CPs. Similarly, humidifiers, hair dryers and vacuum cleaners containing

AgNPs can also produce AgNPs aerosols. Research conducted by Benn et al. used two humidifiers made of plastic containing AgNPs to evaluate the fate of released AgNPs (20). Although, the product claimed to contain AgNPs, the authors were unable to confirm the presence of AgNPs in the humidifier using SEM/EDX analysis due to the instability of the plastic sample under the electron beam. However, they found AgNPs in the water vapors released from the humidifiers, but at concentrations too low (1.1 ± 0.4 μg Ag L–1 at 100 mL/hr flow rate and 0.19 μg Ag L–1 at

420 mL/hr flow rate) to allow sufficient characterization.

In a similar experiment, Quadros et al., simulated the use scenario of a humidifier and an antibacterial disinfectant spray by spraying inside a furnished room with carpeted floors, painted walls, a door and a window (41). According to their finding, an Ag concentration of (2.3±0.7 ppb) was present in the condensed vapor released from the humidifier.

Other studies assessing the fate of AgNPs released from toothpaste and shampoo have been reported (20). The toothpaste and the shampoo used in that study were claimed to contain colloidal

Ag as well or a mixture of colloidal and ionic Ag, respectively. The major objective of the study was to characterize the Ag particles in the product before and after being released to the environment through washing with tap water. However, this experiment does not simulate the real usage of these products. For instance, the influences of water quality and interactions with saliva

(in the case of toothpaste) on the AgNPs were not considered. Nevertheless, the majority of the

Ag particles released from both the toothpaste and shampoo samples were larger than 100 nm. In

147 addition, the release of AgNPs was confirmed by SEM/EDX analysis. No other experiments have been reported using personal care products.

AgNPs are also incorporated into paints. It has been reported that paints release the highest percentage of AgNPs into water (79). However, few studies of the fate of AgNPs released from paints have been reported. A study simulating the outdoor applications of facade panels painted with AgNPs containing paints is the first investigation on the AgNPs released from paints in aquatic environments (35).

All of the previous research reported on AgNP-containing CPs are summarized in Table 5.1.

5.4 Potential Ecological Risks Associated with AgNPs

The increased incorporation of AgNPs into CPs elevates the chance they will be released into the environment. Further knowledge of the fate of AgNPs in different environmental compartments will assist in understanding the potential ecological and health risks associated with them. It is thought that the eventual destinations of AgNPs released from CPs are water, air and soil.

Throughout the lifecycle of the AgNPs, from manufacturing to use to disposal (cradle to grave),

AgNPs are naturally exposed to a range of conditions. Therefore, AgNPs may undergo physicochemical transformations throughout their life cycle, completely altering their characteristics (e. g., toxicity) significantly. Therefore understanding the fate and transformations of nanomaterials (AgNPs), especially those incorporated into consumer products, in various environmental conditions and scenarios is of great importance.

5.4.1 Transformations of AgNPs in Aquatic Environments

All bodies of surface water may potentially be contaminated by AgNPs released from CPs. For instance, textiles contain high concentrations of AgNPs which have demonstrated to be released

148 into water during washing. Humidifiers, antibacterial soaps, detergents, paints, inks, washing machines and water filters can directly release AgNPs into water sources. Furthermore, the quality of water plays an important role in the potential transformations (size, shape and speciation) of

AgNPs; a number of studies have reported such transformations (50-55). Nevertheless, the reported studies primarily investigated pristine AgNPs; only a few looked at AgNPs released from

CPs. According to research conducted on environmental exposure of AgNPs from consumer products, the following possible scenarios have been investigated:

5.4.1.1 Usage and Disposal Scenarios

 Release into water from washing of textiles

AgNPs incorporated into textiles can be released in to water during washing (21, 22, 24, 27, 34).

The impact of water quality (tap water versus ultrapure water) on the release of AgNPs in socks was assessed by simulating washing conditions. The Ag released were characterized and reportedly contained both Ag+ and AgNPs as observed in the source material (24). However, higher Ag release rates were observed with ultrapure water compared to tap water due to the corrosive nature of ultrapure water. Interestingly, a similar study on AgNP-containing socks resulted in the formation of AgCl in the presence of a hypochlorite (bleach)/detergent solution

(18). Although the prevailing mechanism of AgCl formation was not confirmed, two possibilities were proposed:

 The Ag+ released (dissolution) from socks reacts with Cl forming AgCl

 The AgNPs are encapsulated by AgCl as a result of surface oxidation of AgNPs by

hypochlorite

149

The detergent used in this study was not representative of typical laundry detergents due to containing high hypochlorite concentrations relative to actual household washing conditions.

Nevertheless, the results demonstrate, with an oxidizer and chloride present, a significant fraction of the AgNPs in the fabric will be converted to AgCl. In 2009, a detailed washing experiment was conducted by Geranio et al., using different types of AgNP-containing textiles (21). That experiment assessed the influence of pH, surfactants and oxidizing agents on AgNPs; it was observed that Ag+ release was higher at pH 7 than at pH 10. The presence of bleach accelerated the release of Ag+ following the chemical reaction given below.

+ . - Ag (s) + H2O2 Ag (aq) + OH + OH (aq)

The produced hydroxyl radicals promote further oxidation of metallic Ag and Ag+ release.

. Ag (s) + OH Ag+ (aq) + OH- (aq)

This experiment also identified the impact of mechanical abrasion on Ag+ release during washing.

A significant release of coarse (> 450 nm) AgNPs was observed with most of the textiles after washing with a washing machine. The coarse fraction may include AgNP aggregates, precipitates and AgNP-embedded textile fibers, all of which may exhibit significantly different behaviors under wastewater treatment conditions. These results demonstrate the importance of thorough characterization of the coarse fraction of nanoparticles in future studies. In a similar experiment conducted by Lorenz et al., the release of different forms of AgNPs was observed from textiles with different fabric compositions and types of AgNPs (Ag, AgCl, Agnanowires) (22). This study concluded that not only are aggregated AgNPs released but also nano-AgCl or nano Ag2S may be

150 released from textiles, further emphasizing the importance of studying the fate of different forms of AgNPs under wastewater treatment conditions. The washing procedures not only caused AgNPs to be released from textiles, but also modified the Ag speciation remaining on the textile. This phenomenon must be taken into consideration, as the AgNPs integrated into the consumer product may be completely altered after prolonged usage. These changes may in turn affect the antibacterial function of the products leading to unknown environmental scenarios.

The transformation of the Ag species in textiles was studied by Lombi et al., before and after washing using XAS analysis (27). Commercially-available textiles containing different species of

Ag and AgNPs (metallic AgNPs, AgCl-NPs, Ag2S-NPs, Ag-oxides, Ag-phosphates, Ag-sulfates,

Ag-zeolites) were investigated in this experiment. The authors reported the formation of more stable Ag species such as AgCl and Ag2S in the textiles after washing.

The concentrations of AgNPs released are highly dependent on their mode of association/integration into the fabric material. Quadros et al. reported that once AgNPs embedded onto the surface of textiles are depleted during initial washing cycles, further release is not observed (41). The antibacterial properties of the textiles after long term washing should be reduced; however, this reduction has not been experimentally confirmed.

The effects of the manufacturing process, including the method of AgNPs incorporation into the

CPs (e. g., AgNPs integrated into or coating the surface), on AgNPs release has been investigated

(17). A greater release of AgNPs were reported from textiles that have AgNPs coating the surface

(51% of the total theoretical concentration of Ag incorporated into the textiles) compared to that of textiles with AgNPs integrated in them (21% of the total theoretical concentration of Ag incorporated into the textiles).

151

In similar research, Meyer et al. conducted a life-cycle assessment of AgNPs in commercially available socks (42). The impact of a number of factors on AgNP transformations were assessed, primarily focused on the production/manufacturing phase. The potential transformations that may occur during the usage and disposal scenarios were not accounted for in the life-cycle assessment.

 CP utilization – release into water from products other than textiles

Although many AgNP-CPs are not textiles, few studies have been reported assessing the fate of these AgNPs. One of the few reported studies was conducted using 13 different AgNP-containing baby products, including plush toys, baby blankets, sippy cups, baby wipes, breast milk storage bags, humidifiers and disinfecting sprays (41). The products were analyzed in a leaching study using different fluids (milk, orange juice, saliva, urine and sweat) to simulate realistic usage scenarios. Regardless of the product, the highest release of AgNPs occurred when the CPs were exposed to urine or sweat due to the presence of Cl- salts. Although the authors quantified the concentrations of AgNPs released in the simulated scenarios, little attention was given to the characterization of these AgNPs, especially to the potential speciation alterations that may occur under each scenario.

Another example in which AgNPs may be released into the water source are washing machines with AgNP-promoted antibacterial functioning. A study by Farkas et al., reports simulation of common clothes washing process utilizing a nano-Ag washing machine (43). The investigation was conducted with and without the use of detergents as well as using both activated and deactivated modes of the “nano-silver wash” function. 11μg/L of Ag was present in the wash water from the “nano-silver wash” mode. The AgNPs discharged mainly consisted of particles < 20 nm.

Once again, a detailed characterization of the AgNPs was not attempted. However, the effect of the discharged water (containing AgNPs) on bacterial growth was further investigated. The

152 exposure impacts of the washing machine discharge on natural bacterial-plankton communities after 7 days were measured by flow cytometry. A significant reduction in bacterial growth (85% at 12.5 μg/L and 62% at 2.5 μg/L AgNP exposure) was reported after 7 days, demonstrating the ecotoxicity of the released AgNPs.

In addition to solid CPs, liquid CPs containing AgNPs may also enter various water sources. One of the most common cases of liquid AgNP-CPs is paint. AgNP-containing paints may potentially release AgNPs to water sources through routes such as exposure to rain water and washing tools

(e. g., paint brush, containers, and gloves). To further assess this scenario, the water run-off from façade panels coated with AgNPs-paints has been investigated. The run-off contained both individual AgNPs as well as AgNPs bound by organic binders. The interaction of Ag(0) with dissolved H2S in rain water, leading to transformations of Ag(0) to Ag2S, was confirmed by

SEM/EDX analysis (35).

 Wastewater treatment scenarios

A number of physical and chemical processes are used in WWTP facilities, including aeration, activated sludge treatment, separation, filtration, O3 treatment, UV exposure and Cl treatment.

Understanding the potential transformations of AgNPs in wastewater after exposure to different processes and different conditions is essential. Several studies have been conducted to investigate the potential transformations of AgNPs under simulated wastewater treatment conditions using pristine AgNPs (44-47). Most of these studies confirmed the formation of Ag2S as a more stable/nontoxic form of Ag in the sludge and effluent water. However, according to Thalmann et al., Ag2S can undergo further transformations under certain conditions resulting in a more toxic form. (48). Specifically, this study reveals that ozone treatment of WWTP effluent results in the oxidation of Ag2S, and therefore, the creation of a more toxic form of silver in the effluent,

153

However, research on the fate of AgNPs released from CPs under wastewater treatment conditions is missing and research thus far has been focused on pristine laboratory synthesized AgNPs.

Activated sludge digestion is generally used by WWTPs to degrade organic matter. The fate of

AgNPs released from socks was assessed after exposure to activated sludge in a quasi-equilibrium batch adsorption study by Benn et al. (24). An adsorption capacity of 3.4 - 17 μg-Ag/g-biomass was calculated and consistent with a Freundlich isotherm model. The data were applied to a general fate model for sorption to estimate the Ag concentrations in the effluent and sludge from a WWTP.

A similar adsorption study simulating WWTP conditions was conducted with dietary supplementary beverages containing Ag in metallic form (30). About 80-90% of the Ag was removed with a biomass dose of 1000 mg/L; however, no analyses or model predictions were presented on the physicochemical characteristics of the AgNP in this study.

Studies following the fate of AgNPs released from CPs from initial release to sewage sludge have yet to be reported. Kim et al., characterized real sewage sludge attempting to identify AgNPs; in situ formation of Ag2S in water treatment processes was confirmed, resulting from the reaction between reduced sulfur and AgNPs under anaerobic conditions (49).

 Natural surface water

AgNPs accumulated in natural bodies of surface water may be exposed to various environmental conditions and components, such as direct sunlight, fluctuating temperatures, dissolved oxygen, natural organic matter (NOM), pH and salinity. These components play a major role in the transformations of AgNPs in these waters. The impact of some of these parameters on AgNPs transformations have been evaluated in the past for laboratory-synthesized AgNPs. These studies included investigations of the transformations of AgNPs as a function of pH, ionic strength and

154 background electrolytes (50), transformations of Ag+ to AgNPs in the presence of natural organic matter (NOM) (51); aggregation and dissolution as a function of sunlight and different capping agents (52), morphological transformations due to the influence of inorganic anions (53, 54) and a function of both NOM and light irradiation (55).

However, detailed investigations on the AgNPs released from CPs are relatively scarce. In a rare example of such investigations, Cleveland et al., simulated exposure of AgNP-containing CPs (a teddy bear, a wound dressing and socks) to estuarine conditions using mesocosm systems contained with seawater, sediment, marine organisms (crabs, mussels), biofilm slides, and marsh grass (56). The objective of the study was to measure the bioaccumulation of AgNPs from CPs in aquatic organisms under realistic conditions. The transformations of AgNPs after exposure to the relevant environmental condition was not investigated in this study. In general, detailed experiments evaluating transformations of AgNPs released from CPs into natural surface waters are still lacking.

 Landfill disposal

AgNP-enabled consumer products have varying lifespans depending on the specific nature and application of the product prior to disposal; one of the disposal scenarios commonly discussed is landfilling. At the time of disposal, AgNP-CPs, especially solid AgNP-CPs, may contain significant amounts of AgNPs. The increased potential accumulation of AgNPs in landfills risks disturbing the system due to the antimicrobial properties of AgNPs. However, few studies have been conducted to evaluate the fate and impact of AgNPs from CPs disposed in landfills. One of the first studies in this area was conducted by Benn et al. (20). This experiment was designed to understand the fate of common AgNP-CPs (a toy teddy bear, a plastic humidifier and medical masks) under landfill leaching conditions using a standard EPA leaching protocol commonly used

155 to simulate landfill leaching conditions, the toxicity characteristic leaching protocol (TCLP). The authors reported release of Ag from humidifiers under landfill leaching conditions was ten times faster than the intended release rate into air during use. A more recent study assessed Ag leaching from AgNPs containing commercial textiles with different fabric compositions (synthetic and natural) as a function of the method of AgNP incorporation (surface coated versus embedded in fabric) (57). The total Ag leached from the consumer products was 7-53% of the total content of each product. Nevertheless, the concentrations of Ag released were below the RCRA regulatory level of Ag (5 mg/L) established by the US EPA (58). Furthermore, landfill disposal conditions demonstrated a higher release of AgNPs from the synthetic fabric with surface coated AgNPs.

However, this experiment was conducted with new consumer products and as a result, the data may be exaggerated from actual disposal scenarios expected to occur after CPs have been used.

5.4.1.2 Air

In some applications scenarios, such as antibacterial and disinfectant sprays, air conditioning filters, humidifiers and vacuum cleaners, AgNPs in CPs can be directly released to the air. Air release of AgNPs is important as AgNPs may directly enter the human body through inhalation.

Atmospheric gases, water vapors and particulates present in the air can interact with the AgNPs, resulting in physicochemical transformations. The wide range of potential transformations of

AgNPs may result in the formation of species exhibiting higher levels of toxicity compared to the pristine AgNPs. However, research assessing the transformations of AgNPs in aerosols, under atmospheric conditions is scarce. This literature review found only one study that evaluated the environmental fate of AgNPs in CPs (AgNP-paint) after release to air (35). The authors of that study postulated the formation of an Ag2O ring around the AgNPs caused by oxidation due to air

156 exposure. There is a wide gap of knowledge on the potential transformations of AgNPs in aerosols

(CPs) under diverse and relevant conditions still exists.

5.4.1.3 Soils

Although direct soil contamination by AgNPs released from CPs may be uncommon, AgNPs can accumulate in soils and sediments indirectly. AgNPs can accumulate in sediments through adsorption onto inorganic and organic matter in water sediments, as well as through disposal of

AgNPs containing biosolids from wastewater treatment plants into soil.

The accumulation of AgNPs in soils may disturb key microbial communities, organisms and plant life. A few studies have reported the impact of laboratory synthesized AgNPs on soil ecosystems, which were discussed separately in Section 4.2. However, a comprehensive investigation of the impacts of AgNPs from CPs on vital soil parameters has not yet been reported.

5.5 Ecotoxicity Analyses

Bioavailability, regardless of the route of exposure, underlying mechanisms of toxicity, remains a vital factor for the determination of the environmental impacts of AgNPs released from consumer products. The bioavailability of AgNPs depends on the physicochemical characteristics of AgNPs, the surrounding environmental conditions; and the mode of interactions between AgNPs and target organisms (80) Thus, the environmental risk assessment of AgNP-CPs requires thorough characterization of NPs before, during and after exposure scenarios.

5.5.1 Tools for Ecotoxicity Impact Evaluation

Ecotoxicity analyses for AgNPs exposure are conducted using exposure and transformation analyses such; toxicity assays performed on ecologically relevant and significant organisms; and toxicogenomic, metallomic and proteomic approaches (Omics endpoints).

157

A number of studies have been conducted on AgNPs using ecologically relevant organisms (60-

63). Some of these studies investigated impacts associated to AgNPs exposure in aquatic environments and on aquatic organisms such as fish (61, 63-65), gastropods (66, 67), mussels (68), endobenthic organisms (12), and bacteria (62, 69). Other studies focused on AgNP-induced phytotoxicity on terrestrial plants (60, 70, 71) and impacts on soil-microorganisms (72). Generally, these studies reported that AgNPs induced toxicity in living cells through increased phenoloxidase and lysozyme activity, increased ROS generation and oxidative stress, damaged DNA, altered genes, and decreased embryonic development. The AgNPs induced toxicity to terrestrial plants by reducting root and plant growth, reduced chlorophyll content, and increased superoxide dismutase activity and reduced fruit productivity.

Based on the existing data of AgNP-induced ecotoxicity, it is confirmed that AgNPs can cause harmful effects on ecosystem. However, these studies used pristine AgNPs. Due to the characteristics and behavioral differences between the laboratory-synthesized AgNPs and the ones released from CPs, it is highly important to perform ecotoxicity studies using the AgNPs- containing CPs prior to drawing any conclusions.

5.6 Research Recommendations and Gaps

Currently, many AgNPs containing consumer products are available in the market. The majority of these products have not been evaluated under relevant environmental conditions. The overall assessment of the fate, transport and environmental impacts of AgNPs in consumer products is deficient in addressing the following areas.

158

5.6.1 Consideration of All Possible Interactions in a given Scenario

The possible interactions between AgNPs-CPs and relevant environmental factors must be taken into consideration according to the specific applications of the consumer product (e.g., the potential transformations of AgNPs during the use of textiles due to body temperature, excess heat from ironing or drying clothes, exposure to sunlight (UV light), and the presence of other substances).

5.6.2 Effects of Capping Agents

Due to the instability of AgNPs caused by their high surface area to volume ratio, a wide variety of capping agents are used stabilize them. Citrate, Tween 20, PVP and polyethlene glycol (PEG) are some of the most widely used capping agents for the synthesis of AgNPs. Although coated

AgNPs are commonly incorporated into CPs, most of the studies cited in this report did not mention or report any details on the type, characteristics or potential effects of the coating agents incorporated into the products.

5.6.3 Effects of Aging and Disposal Scenario

The effect of aging on AgNPs released from consumer products has not been studied thoroughly.

Once AgNPs are released as composites of a sample matrix, it is important to understand the fate and potential physicochemical transformations occurring over time. For instance, the AgNPs integrated in façade paints were demonstrated to be released combined with the organic binder in the paint (35). The organic binder may degrade over time and result in the release of increased concentrations of AgNPs. Aging may also impact the physicochemical characteristics of the

AgNPs and must be considered in future studies.

159

The fate of AgNPs in consumer products has not been commonly evaluated and remains largely unknown. For instance, AgNPs are widely used in the medical field (e. g., in instruments and appliances), but the fate of these AgNP-CPs after use has rarely been investigated. Additionally, due to their short life span, as well as, high consumption, understanding the fate of food packaging and AgNP-enabled plastics are critical.

5.6.4 Potential Transformations as a Function of Natural Environmental Factors

Physicochemical transformations of AgNPs are possible when they are exposed to natural environmental factors such as exposure to sunlight, oxygen, sulfides and pH changes. With the exception of a few experiments on textile washing scenarios, the possible transformations of

AgNPs released from different CPs under various environmental conditions have not been reported. Furthermore, the potential environmental impacts of the transformed AgNPs have been rarely studied.

5.6.5 Wastewater Treatment Conditions

A comprehensive analysis of the fate of AgNPs released from CPs under wastewater treatment conditions has not been conducted. Research on this subject consists of two studies which reported the biosorption of AgNPs by activated sludge (20, 30). However, the impact of other treatments commonly utilized in wastewater treatment facilities, such as chlorination, O3 or UV disinfection, has not been reported.

5.6.6 Aerosol and Atmospheric Conditions

Studies on AgNPs in aerosols have been focused primarily on the concentrations of the released nanoparticles, particle size and form. However, transformations of AgNPs in different environments (air composition, humidity and the presence of H2S) are likely and must therefore

160 be evaluated. While a limited number of liquid-phase AgNPs CPs are available, the fate of the

AgNPs after exposure to sunlight, atmospheric CO2, H2S dissolved in rain and other particulates in air has not been extensively studied.

5.6.7 Ecotoxicity Analyses

Many research studies have included ecotoxicity analyses on pristine laboratory synthesized

AgNPs using different aquatic, bacterial, and plants species. However with the exception of one study, similar ecotoxicity analyses have not been performed with AgNP-containing CPs.

Cleveland et al., conducted a mesocosm experiment by exposing AgNP-CPs (teddy bear, wound dressing and socks) to estuarine water conditions (56). Even in this study, a detailed toxicity assessment was not conducted using standard toxicity assays. Detailed ecotoxicity analyses on

AgNP-CPs for different ecosystems (water, soil, and sediments) are essential. Previous studies that used laboratory-synthesized AgNPs, evaluated the potential toxicity and interactions of AgNPs on a wide range of different organisms and environmentally relevant bacterial species. Similarly for

AgNP-containing CPs, a detailed ecotoxicity assessment of on a variety of organisms should be completed.

161

Literature Cited

1. Chernousova, S.; Epple, M. Silver as antibacterial agent: ion, nanoparticle and metal. Angewandte Chemie. 2013, 52, (6), 1636-1653.

2. Rai, M.; Yadav, A.; Gade, A. Silver nanoparticles as a new generation of antimicrobials. Biotechnology Advances. 2009, 27, (1), 76-83.

3. http://www.nanotechproject.org/cpi/browse/nanomaterials/silver-nanoparticle/ Accessed – 3/24/2016.

4. DiRienzo, M. New applications of silver. The LBMA Precious Metals Conference 2006, Montreux.

5. Vance, M. E.; Kuiken, T.; Vejerano, E. P.; McGinnis, S. P.; Hochella, M. F.; Rejeski, D.; Hull, M. S. Nanotechnology in the real world: Redeveloping the nanomaterial consumer product inventory. Journal of Nanotechnology. 2015, 6, 1769-1780.

6. Tolaymat, T.; El Badawy, A.; Genaidy, A.; Scheckel, K.; Luxton, T.; Suidan, M. An evidence- based environmental perspective of manufactured silver nanoparticle in syntheses and applications: A systematic review and critical appraisal of peer-reviewed scientific papers. Science of the Total Environment. 2010, 408, (5), 999-1006.

7. Balan, L.; Malval, J.; Schneider, R.; Burget, D. Silver nanoparticles: new synthesis, characterization and photophysical properties. Material Chemistry and Physics. 2007, 104, 417–421.

8. Sharma, V. K.; Yngard, R. A; Lin, Y. Silver nanoparticles: green synthesis and antimicrobial activities. Adv. Colloid Interface Sci. 2009, 145, 83–96.

9. Baalousha, M.; Nur, Y.; Romer, I.; Tejamaya, M.; Lead, J. R. Effect of monovalent and divalent cations, anions, fulvic acid on aggregation of citrate coated silver nanoparticles. Science of the total environment. 2013, 455, 119-131.

10. El-Badawy, A. M.; Luxton, T. P.; Silva, R. G.; Scheckel, K. G.; Suidan, M. T.; Tolaymat, T. M. Impact of environmental conditions (pH, ionic strength and electrolyte type) on the surface charge and aggregation of silver nanoparticles suspensions. Environmental Science and Technology. 2010, 44, (4), 1260-1266.

11. Brunetti, G.; Donner, E.; Laera, G.; Sekine, R.; Scheckel, K.G.; Khaksar, M.; Vasilev, K.; De Mastro, G.; Lombi, E. Fate of zinc and silver engineered nanoparticles in sewerage networks. Water Research. 2015, 77, 72-84.

12. Buffet, P. E.; Zalouk-Vergnoux, A.; Chatel, A.; Berthet, B.; Metais, I.; Perrein-Ettajani, H.; Poirier, L.; Luna-Acosta, A.; Thomas-Guyon, H.; Risso-de Faverney, C.; Guibbolini, M.;

162

Gilliland, D.; Valsami-Jones, E.; Mouneyrac, C. A marine mesocosm study on the environmental fate of silver nanoparticles and toxicity effects on two endobenthic species: the ragworm Hediste diversicolor and the bivalve mollusc Scrobicularia plana. Science of the Total Environment. 2014, 470, 1151-1159.

13. Zhang, C.; Hu, Z.; Deng, B. Silver nanoparticles in aquatic environments: Physiochemical behavior and antimicrobial mechanisms. Water Research. 2016, 88, 403-427.

14. Asharani, P. V.; Low, K. M. G.; Hande, M. P.; Valiyaveettil, S. Cytotoxicity and genotoxicity of silver nanoparticles in human cells. ACS Nano. 2009, 3, (2), 279-290.

15. Hussain, S. M.; Hess, K. L.; Gearhart, J. M.; Geiss, K. T.; Schlager, J. J. In vitro toxicity of nanoparticles in BPL 3A rat liver cells. Toxicology in Vitro. 2005, 19, (7), 975-983.

16. Navarro, E.; Piccapietra, F.; Wagner, B.; Marconi, F.; Kaegi, R.; Odzak, N.; Sigg, L.; Behra, R. Toxicity of silver nanoparticles to Chlamydomonas reinhardtii. Environmental Science and Technology. 2008, 42, (23), 8959-8964.

17. Stefaniak, A. B.; Duling, M. G.; Lawrence, R. B.; Thomas, T. A.; LeBouf, R. F.; Wade, E. E.; Virji, M. A. Dermal exposure potential from textiles that contain silver nanoparticles. International Journal of Occupational and Environmental Health. 2014, 20, 220-234.

18. Impellitteri, C. A.; Tolaymat, T. M.; Scheckel, K. G. The speciation of silver nanoparticles in antimicrobial fabric before and after exposure to a hypochlorite/detergent solution. Journal of Environmental Quality. 2009, 38, 1528-1530.

19. U.S. EPA. Nanomaterial Case Study: Nanoscale Silver in Disinfectant Spray (Final Report). U.S. Environmental Protection Agency, Washington, DC, EPA/600/R-10/081F, 2012.

20. Benn, T.; Cavanagh, B.; Hristovski, K.; Posner, J. D.; Westerhoff, P. The release of nanosilver from consumer products used in the home. Journal of Environmental Quality. 2010, 39, 1875- 1882.

21. Geranio, L.; Heuberger, M.; Nowack, B. The behavior of silver Nanotextiles during washing. Environmental Science and Technology. 2009, 43, 8113-8118.

22. Lorenz, C.; Windler, L.; von Goetz, N.; Lehmann, R. P.; Schuppler, M.; Hungerbuhler, K.; Heuberger, M.; Nowack, B. Characterization of silver release from commercially available functional (nano)textiles. Chemosphere. 2012, 89, 817-824.

23. Mitrano, D. M.; Rimmele, E.; Wichser, A.; Erni, R.; Height, M.; Nowack, B. Presence of nanoparticles in wash water from conventional silver and nano-silver textiles. ACS Nano. 2014, 8, (7), 7208-7219.

163

24. Benn, T. M.; Westerhoff, P. Nanoparticle silver released into water from commercially available sock fabrics. Environmental Science and Technology. 2008, 42, 4133-4139.

25. Kulthong, K.; Srisung, S.; Boonpavanitchakul, K.; Kangwansupamonkon, W.; Maniratanachote, R. Determination of silver nanoparticle release from antibacterial fabrics into artificial sweat. Particle and Fibre Toxicology. 2010, 7:8.

26. Tulve, N. S.; Stefaniak, A. B.;Vance, M. E.; Rogers, K.; Mwilu, S.; LeBouf, R. F.; Schwegler- Berry, D.; Willis, R.; Thomas, T. A.; Marr, L. C. Characterization of silver nanoparticles in selected consumer products and its relevance for predicting children’s potential exposure. International Journal of Hygiene and Environmental Health. 2015, 218, 345-357.

27. Lombi, E.; Donner, E.; Scheckel, K. G.; Sekine, R.; Lorenz, C.; Von Goetz, N.; Nowack, B. Silver speciation and release in commercial antimicrobial textiles as influenced by washing. Chemosphere. 2014, 111, 352-358.

28. Bianco, C.; Kezic, S.; Crosera, M.; Svetlicic, V.; Segota, S.; Maina, G.; Romano, C.; Larese, F.; Adami, G. In vitro percutaneous penetration and characterization of silver from silver- containing textiles. International Journal of Nanomedicine. 2015, 10, 1899-1908.

29. Cascio, C.; Geiss, O.; Franchini, F.; Ojea-Jimenez, I.; Rossi, F.; Gilliland, D.; Calzolai, L. Detection, quantification and derivation of number size distribution of silver nanoparticles in antimicrobial consumer products. Journal of Analytical Atomic Spectrometry. 2015, 30, 1255- 1265.

30. Reed, R. B.; Faust, J. J.; Yang, Y.; Doudrick, K.; Capco, D. G.; Hristovski, K.; Westerhoff, P. Characterization of nanomaterials in metal colloid-containing dietary supplement drinks and assessment of their potential interactions after ingestion. Sustainable Chemistry & Engineering. 2014, 2, 1616-1624.

31. Losert, S.; Hess, A.; Ilari, G.; von Goetz, N.; Hungerbuehler, K. Online characterization of nano-aerosols released by commercial spray products using SMPS–ICPMS coupling. Journal of Nanoparticle Research. 2015, 17, 293.

32. Cascio, C.; Gilliland, D.; Rossi, F.; Calzolai, L.; Contado, C. Critical experimental evaluation of key methods to detect, size and quantify nanoparticulate silver. Analytical Chemistry. 2014, 86, 12143-12151.

33. Quadros, M. E.; Marr, L. C. Silver nanoparticles and total aerosols emitted by nanotechnology- related consumer spray products. Environmental Science and Technology. 2011, 45, 10713- 10719.

164

34. Chao, J.; Liu, J.; Yu, S.; Feng, Y.; Tan, Z.; Liu, R.; Yin, Y. Speciation analysis of Silver nanoparticles and Silver ions in antibacterial products and environmental waters via cloud point extraction-based separation. Analytical Chemistry. 2011, 83, 6875-6882.

35. Kaegi, R.; Sinnet, B.; Zuleeg, S.; Hagendorfer, H.; Mueller, E.; Vonbank, R.; Boller, M.; Burkhardt, M. Release of silver nanoparticles from outdoor facades. Environmental Pollution. 2010, 158, 2900-2905.

36. Nazarenko, Y.; Han, T. W.; Lioy, P. J. Mainelis, G. Potential for exposure to engineered nanoparticles from nanotechnology-based consumer spray products. Journal of Exposure Science and Environmental Epidemiology. 2011, 21, (5), 515-528.

37. Stuart, E. J. E.; Tschulik, K.; Omaovic, D.; Cullen, J. T.; Jurkschat, K.; Crossley, A.; Compton, R. G. Electrochemical detection of commercial silver nanoparticles: identification, sizing and detection in environmental media. Nanotechnology. 2013, 24: 444002.

38. Kim, E.; Lee, J. H.; Kim, J. K.; Lee, G. H.; An, K.; Park, J. D.; Yu, I. J. Case study on risk evaluation of Silver nanoparticle exposure from antibacterial sprays containing Silver nanoparticles. Journal of nanomaterials. 2015, Article ID-346586.

39. Hagendorfer, H.; Lorenz, C.; Kaegi, R.; Sinnet, B.; Gehrig, R.; Goetz, N. V.; Scheringer, M.; Ludwig, C.; Ulrich, A. Size-fractionated characterization and quantification of nanoparticle release rates from a consumer spray product containing engineered nanoparticles. Journal of Nanoparticle Research. 2010, 12, 2481-2494.

40. Jeannet, N.; Fierz, M.; Kalberer, M.; Burtscher, H.; Geiser, M. Nanoaerosol chamber for in- vitro toxicity (NACIVT) studies. Nanotoxicology. 2015, 9, (1), 34-42.

41. Quadros, M. E.; Pierson, R.; Tulve, N. S.; Willis, R.; Rogers, K.; Thomas, T. A.; Marr, L. C. Release of Silver from nanotechnology-based consumer products for children. Environmental Science and Technology. 2013, 47, 8894-8901.

42. Meyer, D.; Curran, M. A.; Gonzalez, M. A. An examination of silver nanoparticles in sock using screening-level life cycle assessment. Journal of Nanoparticle Research. 2011, 13, 147- 156.

43. Farkas, J.; Peter, H.; Christian, P.; Urrea, J. A. G.; Hassellov, M.; Tuoriniemi, J.; Gustafsson, S.; Olsson, E.; Hylland, K.; Thomas, K. V. Characterization of the effluent from a nanosilver producing washing machine. Environmental International. 2011, 37, 1057-1062. 44. Liu, J.; Pennell, K. G.; Hurt, R. H. Kinetics and mechanisms of nanosilver oxysulfidation. Environmental Science and Technology. 2011, 45, 7345-7353.

165

45. Kaegi, R.; Voegelin, A.; Sinnet, B.; Zuleeg, S.; Hagendorfer, H.; Burkhardt, M.; Siegrist, H. Behavior of metallic Silver nanoparticles in a pilot wastewater treatment plant. Environmental Science and Technology. 2011, 45, 3902-3908.

46. Wang, Y.; Westerhoff, P.; Hristovski, K. D. Fate and biological effects of silver, titanium dioxide, and C60 (fullerene) nanomaterials during simulated wastewater treatment processes. Journal of Hazardous Materials. 2012, 201-202, 16-22.

47. Ma, R.; Levard, C.; Judy, J. D.; Unrine, J. M.; Durenkamp, M.; Martin, B.; Jefferson, B.; Lowry, G. V. Fate of Zinc Oxide and Silver nanoparticles in a pilot wastewater treatment plant and in processed biosolids. Environmental Science and Technology. 2014, 48, 104-112.

48. Thalmann, B.; Voegelin, A.; Von Gunten, U.; Behra, R.; Mogenroth, E.; Kaegi, R. Effect of ozone treatment on nano-sized silver sulfide in wastewater effluent. Environmental Science and Technology. 2015, 49, 10911-10919.

49. Kim, B.; Park, C.; Murayama, M.; Hochella, M. F. Discovery and characterization of silver sulfide nanoparticles in final sewage sludge products. Environmental Science and Technology. 2010, 44, 7509-7514.

50. El Badawy, A. M.; Luxton, T. P.; Silva, R. G.; Scheckel, K. G.; Suidan, M. T.; Tolaymat, T. M. Impact of environmental conditions (pH, ionic strength, and electrolyte type) on the surface charge and aggregation of Silver nanoparticles suspensions. Environmental Science and Technology. 2010, 44, 1260-1266.

51. Yin, Y.; Shen, M.; Zhou, X.; Yu, S.; Chao, J.; Liu, J.; Jiang, G. Photoreduction and stabilization capability of molecular weight fractionated natural organic matter in transformation of Silver ion to metallic nanoparticle. Environmental Science and Technology. 2014, 48, 9366-9373.

52. Li, X.; Lenhart, J. J. Aggregation and dissolution of Silver nanoparticles in natural surface water. Environmental Science and Technology. 2012, 46, 5378-5386.

53. An, J.; Tang, B.; Zheng, X.; Zhou, J.; Dong, F.; Xu, S.; Wang, Y.; Zhao, B.; Xu, W. Sculpturing effect of chloride ions in shape transformation from triangular to discal silver nanoplates. The Journal of Physical Chemistry C. 2008, 112, (39), 15176-15182.

54. Zhang, L.; Li, X.; He, R.; Wu, L.; Zhang, L.; Zeng, J. Chloride-induced shape transformation of silver nanoparticles in a water environment. Environmental Pollution. 2015, 204, 145-151.

55. Zou, X.; Shi, J.; Zhang, H. Morphological evolution and reconstruction of silver nanoparticles in aquatic environments: The roles of natural organic matter and light irradiation. Journal of Hazardous Materials. 2015, 292, 61-69.

166

56. Cleveland, D.; Long, S. E.; Pennington, P. L.; Cooper, E.; Fulton, M. H.; Scott, G. I.; Brewer, T.; Davi, J.; Petersen, E. J.; Wood, L. Pilot estuarine mesocosm study on the environmental fate of silver nanomaterials leached from consumer products. Science of the Total Environment. 2012, 421-422, 267-272.

57. Limpiteeprakan, P.; Babed, S. Leaching potential of silver from nanosilver-treated textile products. Environmental Monitoring and Assessment. 2016, 188:156.

58. Code of Federal Regulations – Title 40 Protection of Environment – Hazardous waste characteristics. http://www.gpo.gov/fdsys/pkg/CFR-2011-title40-vol26/pdf/CFR-2011- title40-vol26.pdf

59. Dale, A. L.; Lowry, G. V.; Casman, E. A. Modeling nanosilver transformations in freshwater sediments. Environmental Science and Technology. 2013, 47, 12920-12928.

60. Song, U.; Jun, H.; Waldman, B.; Roh, J.; Kim, Y.; Yi, J.; Lee, E. J. Functional analyses of nanoparticle toxicity: A comparative study of the effects of TiO2 and Ag on tomatoes (Lycopersicon esculentum). Ecotoxicology and Environmental Safety. 2013, 93, (1), 60-67.

61. Bae, E.; Lee, B.; Kim, Y.; Choi, K.; Yi, J. Effect of agglomeration of silver nanoparticle on nanotoxicity depression. The Korean Journal of Chemical Engineering. 2013, 30, (2), 364- 368.

62. Das, P.; Xenopoulos, M. A.; Williams, C. J.; Hoque, A. E.; Metcalfe, C. D. Effects of silver nanoparticles on bacterial activity in natural waters. Environmental Toxicology and Chemistry. 2011, 31, (1), 122-130.

63. Griffitt, R. J.; Brown-Peterson, N. J.; Savin, D. A.; Manning, C. S.; Boube, I.; Ryan, R. A.; Brouwer, M. Effects of chronic nano particulate silver exposure to adult and juvenile sheepshead minnow (Cyprinodon variegatus). Environmental Toxicology and Chemistry. 2012, 31, (1), 160-167.

64. Hoheisel, S. M.; Diamond, S.; Mount, D. Comparison of nanosilver and ionic silver toxicity in Daphnia magna and Pimephales promelas. Environmental Toxicology and Chemistry. 2012, 31, (11), 2557-2563.

65. Asharani, P. V.; Lian, W. Y.; Gong, Z.; Valiyaveettil, S. Toxicity of silver nanoparticles in zebrafish models. Nanotechnology. 2008, 19 (25): 255102.

66. Ramskov, T.; Forbes, V. E.; Gilliland, D.; Selck, H. Accumulation and effects of sediment- associated silver nanoparticles to sediment-dwelling invertebrates. Aquatic Toxicology. 2015, 166, 96-105.

167

67. Croteau, M.; Misra, S. K.; Luoma, S. N.; Valsami-Jones, E. Silver bioaccumulation dynamics in a freshwater invertebrate after aqueous and dietary exposures to nanosized and ionic Ag. Environmental Science and Technology. 2011, 45, (15), 6600-6607.

68. Zuykov, M.; Pelletier, E.; Demers, S. Colloidal complexed silver and silver nanoparticles in extrapallial fluid of Mytilus edulis. Marine Environmental Research. 2011, 71, (1), 17-21.

69. Priester, J. H.; Singhal, A.; Wu, B.; Stucky, G. D.; Holden, P.A. Integrated approach to evaluating the toxicity of novel cysteine-capped silver nanoparticles to Escherichia coli and Pseudomonas aeruginosa. Analyst. 2014, 139, 954-963.

70. Zuvera-Mena, N.; Armendariz, R.; Peralta-Videa, J. R.; Gardea-Torresdey, J. L. Effects of silver nanoparticles on radish sprouts: root growth reduction and modifications in the nutritional value. Frontiers in Plant Science. 2016, 77- article 90.

71. Gubbins, E. J.; Batty, L. C.; Lead J. R. Phytotoxicity of silver nanoparticles to lemna minor L. Environmental Pollution. 2011, 159, (6), 1551-1559.

72. Sillen, W. M. A.; Thijs, S.; Abbamondi, G. R.; Janssen, J.; Weyens, N.; White, J. C.; Vangronsveld, J. Effects of silver nanoparticles on soil microorganisms and maize biomass are linked in the rhizosphere. Soil Biology and Biochemistry. 2015, 91, 14-22.

73. Djokic, S. Treatment of various surfaces with silver and its compounds for topical wound dressings, catheter and other biomedical applications. ECS Transactions. 2008, 11, (21), 1-12.

74. Chaloupka, K.; Malam, Y.; Seifalian, A. M. Nanosilver as a new generation of nanoproducts in biomedical applications. Trends in Biotechnology. 2010, 28, (11), 580-588.

75. Huang, Y.; Chen, S.; Bing, X.; Gao, C.; Wang, T.; Yuan, B. Nanosilver migrated into food‐ simulating solutions from commercially available food fresh containers. Packaging technology and science. 2011, 24, 291-297.

76. Echegoyen, Y.; Nerin, C. Nanoparticle release from nano-silver antimicrobial food containers. Food and Chemical Toxicology. 2013, 62, 16-22.

77. Mackevica, A.; Olsson, M. E.; Hansen, S. F. Silver nanoparticle release from commercially available plastic food containers into food simulants. Journal of Nanoparticle Research. 2016, 18:5. 78. Ramos, K.; Gomez-Gomez, M. M.; Camara, C.; Ramos, L. Silver speciation and characterization of nanoparticles released from plastic food containers by single particle ICPMS. Talanta. 2016, 151, 83-90.

168

79. Mueller, N.; Nowack, B. Exposure modeling of engineered nanoparticles in the environment. Environmental Science and Technology. 2008, 42, (12), 4447-4453.

80. Kahru, A.; Dubourguier, H. From ecotoxicology to nanoecotoxicology. Toxicology. 2010, 269, 105-119.

169

Table 5.1 Summary of past research using AgNP-containing consumer products.

Consumer Product- characteristics Environmental Exposure Human exposure

AgNPs behavior Reference Form of Characterizing Toxicity for Exposure Product Toxicity assays AgNP techniques ecosystem pathways Scenario Transformation AgNP Ag+/ colloidal Ag Washing with ultrapure Model simulation

water and tap water was performed to 100-500 nm Colloidal Ag Ag+ Benn and SEM, TEM, without detergents predict the Ag Socks sized Ag (after prolong exposure to Not studied Not studied Westerhoff, IES, ICP-AES leaching from particles water) 2008 biosolid under Waste water treatment Sludge was not characterized TCLP conditions (adsorption to activated for Ag speciation sludge) Elemental Washing with detergents Impellitteri et Socks XAS AgNP AgCl Not studied Not studied Not studied AgNPs (hypochlorite/detergent) al., 2009 AgNP bound to fiber Ag(0) Ag+ surface,

incorporated Washing (ISO method for More Ag+ release at pH 7 in to fibers, ICP-OES, XFS, washing test) – effect of Geranio et al., Socks than at pH 10 Not studied Not studied Not studied AgCl ISE pH, surfactants, oxidizing 2009 Ag + H O Ag + + incorporated agents 2 2 OH. + OH- in binder on Ag + OH. Ag+ +OH the fiber surface, Release as particles or associated with particles > 100 nm (liquid phase products) Washing with tap water

Released as agglomerates of Socks, T- AgNP (< 20 nm) shirt, (textiles/fabric) medical ICP-OES, Benn et al., Rate of release 0.08 – 0.11μg Not studied Not studied Not studied masks, SEM/EDX, 2010 Ag hr-1 toothpaste, Released in to air during shampoos, use (humidifiers) Characterization was not successful due to the low concentration 1.7 μg/g (medical mask), 54 Landfill disposal ( TCLP μg/g (medical cloth), 0.2 μg/g leaching test to simulate) (humidifier),

170

Consumer Product- characteristics Environmental Exposure Human exposure AgNPs behavior Form of Characterizing Toxicity for Exposure Reference Product Toxicity assays AgNP techniques Scenario Transformation ecosystem pathways

Release into artificial Release of Ag in different GFAAS, TEM, Kulthong et Shirts sweat ( four types with concentrations depend on the Not studied Not studied Not studied SEM/EDX al., 2010 pH range from 4.3 – 8) sweat composition.

ICP-MS, Affect on the SEM,STEM, Washing <20 nm size growth of natural Farkas et al., Nano particle Washing cycle Release of AgNP (11 μg/L) Not studied Not studied Machine AgNPs bacterioplankton 2011 tracking was tested analysis Based on the Ag+, Ag- information wires, Ag obtained from integrate in Not studied manufacturer Release of AgCl, Ag2S, Lorenz et al., Socks, T- polyamide (antibacterial and SEM/EDX, Washing with tap water 2012 shirts, fiber, AgNP activity of the Not studied Not studied Total and detergents Aggregations of metallic trouser incorporated textile after washing concentration AgNPs in cotton has been tested) has measured fibers, by ICP-OES Ag(0), AgCl and XRF Sea water = continuous Ag leaching

Intertidal sediment

Organisms (snail, Wound Estuarine ecosystem with clam, shrimps) = dressing, sea water, intertidal Ag level increased, Cleveland et dress sock, ICP-MS, sediments, mud snail, Not studied peaked and Not studied Not studied al., 2012 stuffed toy clams, gras shrimps, decreased except in bear cordgrass and biofilm snails (snails’ Ag level is X9 higher than others and didn’t decreases)

Plant (cordgrass)

Biofilm More Ag+ release into acidic Ag Release in to artificial sweat than alkaline sweat integrated ICP-OES, External dermal Von Goetz et Socks, shirt sweat, effect of pH (5.5 More particulate Ag in Not studied Not studied polyamide STEM exposure al., 2013 and 8) alkaline sweat than acidic fiber, AgNP sweat

171

Consumer Product- characteristics Environmental Exposure Human exposure Form of Characterizing AgNPs behavior Toxicity for Exposure Reference Product Toxicity assays AgNP techniques Scenario Transformation ecosystem pathways Baby Most product released Ag in Ingestion, blanket, Release into liquid media ionic form via dissolution dermal ( plush toy, ( tap water, synthetic NIOSH Method SEM/EDX, Not studied Not studied breast milk sweat, saliva, urine, milk Synthetic sweat and urine 9102) storage bag, formula, orange juice) extract the highest amount of sippy cup Ag (rubber ring), Quadros et al.,

each product is given given is product each Aerosol release ( in a disinfecting simulated the product use 2013 spray, SMPS, optical in a room with carpeted surface particle counter floor, painted walls, a wipes, very low emission rate Not studied Inhalation Not studied door and a window, two kitchen desks, two cushioned scrubber, chairs, book case and humidifier wooden wardrobe) accessory cube in the AgNPs of Sizes article original inthe Ag integrate in polyamide fiber, AgNP Washing procedure (ISO incorporated Increased the AgCl- NP and 105-IS washing in cotton Ag S-NP fraction. procedure and machine 2 fibers,(Ag S- However, highly depend on Lombi et al., Textiles 2 XANES washing) Not studied Inhalation Not studied NP, Ag the association mode of Ag 2014

zeolite,Ag S, NPs in the original fabric 2 AgNPs transformation AgNO ,AgN 3 within fabric P, AgCl- NPs, Ag oxides) Extraction to artificial ICP-MS, AAS, AgNPs release Ag+ body fluids (sweat and Stefaniak et Textiles SEM/EDX, In sweat, Not studied Not studied Not studied saliva)- effect of pH, al., 2014 DLS Ag+ + Cl- = AgCl temperature Metalic Ag Metallic Ag particles of 20-30 in high ICP-MS, nm. Mitrano et al., Socks concentratio Washing with detergents Not studied Not studied Not studied STEM/EDX, AgNPs in Si matrix or 2014 n (14500 associate with S mg/kg)

In vitro analysis AgNP ICP-AES, ICP- 1. skin analysis ( containing MS, penetration of Ag Bianco et al., textiles 50 – 200 nm Not studied Not studied Not studied dermal SEM/EDX, through skin) 2015 (Acticoat AFM Aggregations of TM) AgCl on the skin layers

172

Consumer Product- characteristics Environmental Exposure Human exposure

Form of Characterizing AgNPs behavior Toxicity for Exposure Reference Product Toxicity assays AgNP techniques Scenario Transformation ecosystem pathways In vivo analysis SEM/EDX, Bianco et al., Shirts Not studied Not studied Not studied dermal Skin analysis ICP-MS 2015 Urine analysis Wound dressing, underwear, baby ICP-AES, blanket, SEM/EDX, gloves, TEM/EDX, Potential pants, UV-vis exposures wipes (ingestion) have Plush toy, estimated using Tulve et al., Not studied Not studied Not studied Not studied sippy cup mathematical 2015 formulae disinfecting spray,

antibacteria ISE, DLS, l spray, TEM/EDX, antiviral UV-vis spray, dietary suppliant in the given is product each in the AgNPs of Size article original Body suit, From 7-53% of total Ag car sheet, released as AgNPs. nursing Nano-Ag cover, fiber, Nano- Simulated landfill Synthetic fabric release more FE-SEM/EDX, Limpiteepraka shoes, pet Ag spray on leaching conditions AgNPs than cotton Not studied Not studied Not studied TEM, n et al., 2016 sheet,socks, the surface, (TCLP leaching protocol) face mask, Nano-Ag Higher release from fabrics shower coated with smaller Nano-Ag cloth in the absence of a binder. SEM, TEM, AgNPs ICP-MS, DLS, Wound extraction on to Nano Ag and UV-Vis , Sussman et al., dressings, Not studied Not studied Not studied water , saline Not studied Ag+ nanoparticle 2015 Catheters and human tracking plasma analyzer (NTA) ActicoatTM , PolyMem SilverR Nano Ag, Boonkaew et (wound Ag- None Not studied Not studied Not studied Dermal Cytotoxicity assays al., 2014 dressings ), sulafdiazine Flamazein eTM (cream)

173

Consumer Product- characteristics Environmental Exposure Human exposure

Form of Characterizing AgNPs behavior Toxicity for Exposure Reference Product Toxicity assays AgNP techniques Scenario Transformation ecosystem pathways Migration of AgNP from the packages to food 1. ultra pure water Plastic bag 2. 4% acetic acid < 300 nm size Ag was ( Sunriver 3. 95% ethanol migrated, Huang et al., Nano Ag SEM/EDX Not studied Not studied Not studied Industrial 4. hexane Time and temperature 2011 Co., China) dependent migration Room temperature and oven temp (40oC, 50oC) for 3, 6, 9, 12 and 15 days Kinetic go greenTM , AgNP of different size and Oso fresh Migration of AgNP from aggregates released from the (Plastic the packages to food food packages food Nano-Ag in 1. ethanol (50%) and storage LDPE (bags) ICP-MS, acetic acid (3%) exposure Echegoyen et Release of nano-plastics Not studied Not studied Not studied containers), or PP SEM/EDX, for 10 days (at 40oC) and al., 2013 along with AgNPs in acidic FreshLonge (containers) 2 hr (at 70oC) conditions rTM (plastic Migration of AgNPs is higher food 2. Microwave heating in microwave than oven storage bags) Kinetic go Green TM , Highest release at acidic Always conditions and higher Fresh Nano Ag, temperatures. Migration of AgNP from Containers the packages to food, TM (food Polyethylene Migration to water and 10% 1. Milli-Q water storage (storage ICP-MS, ethanol was below detection Mackevica et 2. 3% acetic acid Not studied Not studied Not studied box), Fresh box), HDPE TEM/EDX, limit al., 2016 3. 10% ethanol Longer TM, (bags) At 400C for 10 days Miracle Released particles are

Food spherical shape, in Storage TM agglomerates or embedded in (storage polymer matrix bags) NANO Migration of AgNP from BeBe+ Nano Ag the packages to food (Baby 1. Milli-Q water Migration of AgNp depend feeding Polycarbonat 2. 3% acetic acid on the nature of polymer in bottle, e (feeding ICP-MS, SEM - 3. 95% ethanol Ramos et al., the product Not studied Not studied Not studied Baby bottle) EDX 4. 10% ethanol 2016 dream Co. Low release of AgNPs. Ltd, Korea) Polypropylen At 20oC, 40oC for 1, 4, 7 and food e (food box) and 10 days box At 70oC for 2 hr

174

Consumer Product- characteristics Environmental Exposure Human exposure

Form of Characterizing AgNPs behavior Toxicity for Exposure Reference Product Toxicity assays AgNP techniques Scenario Transformation ecosystem pathways Nano-Ag spray in two types Propellant gas sprays release ICP-MS, Spraying the antibacterial of sprays nano-sized particles even NPs Hagendorfer et TEM/EDX, solution inside a glove Not studied Not studied Not studied (propellant are not present in the spray al., 2010 SMPS box gas spray solution and pump spray) Inhalation exposure 8 nm – μm (adapted model Antiperspir size range proposed by ant, plant (solutions) TEM/EDX, Spraying the antibacterial European Lorenz et al., Not studied Not studied Not studied strengtheni ICP-MS, SMPS solution in a glove box chemical 2011 ng agent nm-size agency for aerosols assessing inhalation exposure) Anti-odor spray, Ag particles ICP-MS, TEM, surface > 450 nm, Spraying inside a Quadros and SEM/EDX, Not studied Not studied Not studied Not studied disinfectant ionic Ag and polyethylene chamber Marr, 2011 DLS, UV-vis, , a throat nano Ag spray Nano –Ag TEM, photon containing correlation commercial spectroscopy, Inhalation spray Size of the SMPS, exposure solutions ( AgNPs in differential (spraying near Spraying the antibacterial Ag-spray, each product mobility to a Nazarenko et solution inside a bio Not studied Not studied Not studied hair care is given in particle commercially al., 2011 safety cabinet spray, the original analyzer, available multipurpos article condensation mannequin’s e article counter, head) disinfectant aerodynamic s) particle sizer Develop a Antibacteri separation al hydrogel method for lotion, UV-vis, SEM, Spraying the antibacterial Chao et al., AgNP and Not studied Not studied Not studied Not studied antibacteria TEM solution 2011 Ag+ in l nasal antibacterial spray solutions

175

Consumer Product- characteristics Environmental Exposure Human exposure

Form of Characterizing AgNPs behavior Toxicity for Exposure Reference Product Toxicity assays AgNP techniques Scenario Transformation ecosystem pathways Margin of exposure(MOE) approach to calculate, 1.Dermal SMPS, CPC, Spraying the antibacterial exposure(>MO Antibacteri dust monitor, Kim et al., 10-200 nm solution inside a glove Not studied Not studied E, no risk Not studied al sprays AAS, 2015 box concern level) TEM/EDX 2. Inhalation exposure(

176

Consumer Product- characteristics Environmental Exposure Human exposure

Form of Characterizing AgNPs behavior Toxicity for Exposure Reference Product Toxicity assays AgNP techniques Scenario Transformation ecosystem pathways Total release after 1 year = 0.5 mg/m2 (30% total Ag in the panels) Facades Façade panels painted AgNP released as (with TEM/EDX, with AGNP-paint incorporated with organic Kaegi et al., Not studied Not studied Not studied AgNPs ICP-MS, exposed to outdoor for 17 binder in paint (100 nm – μm 2010 paint) months range)

AgNP Ag2S & Ag2O

Cyclic voltammetry, Colloidal particle-impact silver spray Stuart et al., 0.6-5 nm voltammetry, Not studied Not studied Not studied Not studied Not studied (Higher 2013 nanoparticle Nature) tracking analysis Colloidal silver 5-10 nm (10 solution for ppm Clinical study oral solution) and DLS, UV-vis, Munger et al., consumptio Not studied Not studied Not studied Ingestion 25-40 nm ICP-MS, abdominal and 2014 n (32 ppm cardiac MRI scans (American solution) Silver ,LLC)

177

140 120 100 80 60 40 20 0

Figure 5.1 AgNPs bearing consumer products listed in consumer product inventory of Woodrow Wilson International center for scholars and project on emerging nanotechnology as of March, 2016 (3).

178

Chapter 6

Conclusions and Implications

179

Overall, the work reported in this dissertation intended to examine systematically the potential impacts of AgNPs on complex natural and engineered systems that rely on biodegradation for waste treatment/nutrient removal.

With the growth of nanomaterial applications, nanosilver, due to its antibacterial capabilities, has received special attention and has been extensively utilized. AgNPs have been integrated into an overwhelming number of commercially available consumer products, some of which previously would use metallic/ionic silver or other antibacterial agents to achieve this task. The application and production of nanomaterials and replacement of conventional materials, as the case for silver, brings with it economical and environmental incentives. Generally, nanomaterials exhibit desired properties while utilizing far less raw material. It has been projected that most recognized applications for metallic silver (e. g., antimicrobial) may utilize nanosilver in place of silver to take advantage of nanosilver’ s unique properties.

The underlying antimicrobial properties that have caused AgNPs to be the most commercially incorporated nanomaterial, also render AgNPs toxic to a large variety of bacteria, fungi and viruses. A consequence of the many beneficial applications of AgNPs is that their increasing use and production will also increase their release into environmentally significant waste degradation/nutrient removal processes. These processes rely on biodegradation for processing waste; exposure to AgNPs may negatively impact key microorganisms and bacterial communities essential to those processes.

As the mechanisms of toxicity of nanomaterials can differ from those of bulk matter, concerns exist regarding the potential health, safety and environmental implications of these materials. In the past decade, a significant body of research has been conducted demonstrating the toxicological effects of silver nanoparticles to a wide spectra of bacterial species. However, there is no universal

180 agreement about the exact antimicrobial mechanism of AgNPs, and contradictory hypotheses have been routinely reported. Nonetheless, three modes of interactions between AgNPs and bacteria have been commonly identified as the governing mechanisms of microbial toxicity: direct physical contact between AgNPs and cells, production of reactive oxygen species (ROS) and release of ionic silver (Ag+) from the AgNPs as the source of toxicity.

Research conducted thus far assessing the toxicological impacts of AgNPs on ecologically- significant microorganisms have primarily been conducted under conditions not environmentally representative; the majority of the experiments reported have been completed under controlled laboratory conditions by isolating specific bacterial species and evaluating direct toxicological impacts caused by exposure to AgNPs. Although this approach may assist in understanding the interactions of AgNPs with the specific bacteria under investigation, it does not provide an accurate representation of environmentally relevant scenarios in which the interaction between AgNPs and the target bacterial species are likely to occur. Under natural environmental conditions it is highly unlikely to encounter such simple systems and many more key factors/players/constituents exist and therefore, must be taken into consideration. Additionally, the complex nature of environmental degradation systems and processes heavily depend on proper microbial function. However, as no two complex systems are exactly alike, some variability is inevitable among such systems.

Furthermore, the physicochemical characteristics of silver nanoparticles must be taken into account when addressing the potential environmental impacts associated with AgNPs. Among the ecotoxicological studies reported in the literature, an extensive range of AgNPs with different physicochemical characteristics have been evaluated. The AgNPs used in these studies have been produced employing different methods of synthesis, stabilization mechanisms and purification techniques. However, the environmental implications of AgNPs are reportedly closely associated

181 and determined by the specific physicochemical characteristics (e. g., size, shape, surface charge) of the nanoparticles. Consequently, the variety of reducing agents, capping agents and dispersants used for the synthesis and stabilization of AgNPs, may render direct comparisons, generalizations and categorizations of AgNPs solely based on core elemental composition misleading and possibly inaccurate. As an example, the capping agents utilized for stabilization of AgNPs determine the surface charge, thickness of surface layer, stabilization mechanism and dissolution rate (e. g., the dissolution of Ag+ as the source of toxicity. As a result, the synthesized AgNPs may withstand aggregation and exhibit their nano-specific characteristics for different durations of time. The time in which the AgNPs remain stable in suspension (e. g., nano-form) is directly related to the duration of time in which they can exhibit their antibacterial properties, whether it is through direct physical nanoparticle interactions or oxidative dissolution of metal ions. When the capping agents wear off or lose their stabilizing effects, the dynamics and interactions of the system change entirely. Silver in metallic form Ag(0), without the use of a capping agent (e.g., PVP, Citrate, BPEI and etc) is unstable and will rapidly aggregate or react with common environmentally abundant elements or compounds (e.g., oxygen, sulfur and chloride). In either case, the toxicological capabilities will be greatly reduced.

Therefore, the characteristics of the capping agents play a vital role in determining the duration of time AgNPs exhibit their nano-specific properties, further demonstrating why generalization of nanoparticles according to their core elemental composition may not be accurate or truly representative. The persistence and behavior of AgNPs in an environmental scenario are also fundamentally dependent on the composition of the environmental matrix under consideration.

The behavior of nanoparticles in the environment are expected to depend not only on the physical and chemical characteristics of the nanomaterial, but also largely on the characteristics of the

182 receiving environment. Wastewater treatment plants, landfills, ponds, rivers and lakes are a few examples of natural and engineered environmental compartments through which AgNPs may accumulate.

The following section of this chapter contains the conclusions and key findings associated with each of the four objectives stated in the introduction.

Objective 1: Assessment of the impact of silver nanoparticles on the aerobic degradation process.

The composting process (aerobic degradation) of organic waste was not functionally impacted by the presence of silver nanoparticles (PVP-AgNPs) at concentrations of 2 mgL-1. The functioning and performance of the composting process was evaluated in the presence of 2 mgL-1 PVP-AgNPs.

PVP-AgNPs were chosen for evaluation as the representative nanoparticle because of their resilience under harsh environmental conditions (e. g., changes in pH, conductivity and existence of background electrolytes), to assess a worst case scenario. The results of this work indicate that the primary functional parameters essential for the composting process were not significantly impacted by the presence of PVP-AgNPs at the concentrations tested. In terms of microbial respiration, no significant differences were observed among the Ag treated (Ag+ and AgNPs) groups and negative control. The leachate and solid functional parameters also did not display statistically significant differences between the negative control and treatment groups. However, the taxonomical analysis of the prominent species present revealed time dependent variations among the control and AgNP treated composters.

183

Key Findings:

 In an aerobic environment the AgNPs evaluated in the study did not significantly influence the

composting processes at the concentrations that could be expected to be present in the solid

waste stream.

 Microbial activity can be impacted within hours of AgNP exposure, although at low AgNP

concentrations microbial activity can be relatively stable.

 Toxicity dynamics might be different for complex microbial communities associated with

environmental matrices than for pure cultures. Data from pure culture studies may be

inaccurate in predicting the impact of AgNPs, and nanoparticles in general, on natural and/or

engineered microbial communities.

 AgNPs’ toxicity is dose-dependent and has been observed at higher concentrations in both

aerobic and anaerobic environments. While the surface transformation of AgNPs to AgCl and

Ag2S can reduce the toxicity, complexation with organic matter may also play a role.

 It is unclear whether the bacterial response is caused by bacterial populations resistant to

AgNPs, the decrease in Ag+ release, transformation of AgNPs, or reduction in bioavailability.

Recovering bacterial phylotypes could be responsible for the differences in community

structure between treated and untreated environmental water samples.

 Extrapolating from the results presented herein, similar toxicological behavior of AgNPs

would be expected in the organically rich municipal solid waste landfills if the concentration

of AgNPs was relatively low.

184

Objective 2: Evaluation of the toxicity of different silver nanoparticles under anaerobic conditions.

One of the mechanisms of toxicity associated with AgNPs is reportedly due their natural ability to serve as a source of dissolved Ag+, however, for oxidative dissolution of Ag+ to occur at an adequate rate (e. g., concentrations sufficient for bacterial inhibition), oxygen must be present.

Furthermore, the release of AgNP-containing consumer products into the water stream through washing is reportedly one of the most concerning routes of environmental exposure. Since anaerobic degradation processes are utilized at wastewater treatment plants (WWTPs) to reduce and digest biosolids, the potential toxicity of AgNPs on the anaerobic microorganisms may bring undesired consequences. However, few research studies have been reported on the toxicity of

AgNPs to anaerobic organisms. This investigation assessed the toxicity of three commonly synthesized AgNPs on the anaerobic degradation/decomposition process commonly utilized for sludge and biosolid reduction at WWTPs. It has been reported that the surface charge of AgNPs play a direct role in their mechanisms of toxicity. Therefore, the AgNPs investigated in this study were similar in size distributions (10-15nm), shape (spherical), while displaying different surface charge scenarios ranging from negative (citrate-AgNPs) to near neutral (PVP-AgNPs) to positive

(BPEI-AgNPs). The absence of oxygen prevents oxidative dissolution of the AgNPs to Ag+ as the source and mechanism responsible for the observed toxicity.

Key Findings:

 At low concentrations, regardless of their capping agents and consequential surface charge,

AgNPs did not drastically impact the anaerobic degradation process. The lack of toxicity at

low concentrations is attributed to the functional redundancy built within the microbial

community and not the lack of toxicity of AgNPs.

185

 At high concentrations the cationic BPEI-AgNPs demonstrated elevated toxicity compared to

PVP-AgNPs and the anionic citrate-AgNPs.

 At high concentrations the cationic BPEI-AgNPs demonstrated similar toxicity to that caused

by the positive control group (Ag+) according to the biogas evolution.

 The vast majority of the bacterial cells in the anaerobic sludge are negatively charged (gram

negative), raising the probability that the positive surface charge of the BPEI-AgNPs would

give these particles affinity to the negatively charged bacterial walls while the other AgNPs

would not behave in the same manner.

 These results further support previous research that suggests a different toxicity mechanism is

in place with BPEI-AgNPs as compared to other types of AgNPs.

 Surface-dependent, nano-specific toxicity of BPEI-AgNPs were demonstrated.

 Although toxicity was demonstrated at high concentrations for the positively-charged BPEI-

AgNPs, the lack of toxicity observed at environmentally relevant concentrations suggests that

most WWTPs will be able to process AgNPs without negatively impacting their operations.

 A one size fits all approach to the evaluation of environmental health and safety of

nanoparticles may not be accurate or sufficient.

 Even within one class of nanoparticles (elemental composition), different stabilizing agents

and mechanisms lead to different surface functionality of the AgNP that drastically impact

their behavior in the environment.

 The complex nature of environmental degradation processes whether aerobic or anaerobic

result in the uniqueness of each individual system (anaerobic digester). Therefore, anaerobic

digesters at different WWTPs may vary in composition, level of biological activity, microbial

density, microbial diversity and microbial resistance to heavy metals. Therefore, the

186

toxicological results of this experiment may not directly apply to other systems. For instance,

this research monitored the speciation changes occurring to the AgNPs under simulated

parallel conditions while only altering the sulfide concentrations (e. g., natural levels ~15 ppm

and elevated levels of ~100 ppm). Regardless of the sulfide level the transformation to Ag2S

occurred immediately. However, differences were observed in this rate of transformations over

a 24-hour period seemed to vary depending on the sulfide concentrations present.

 A general characteristic of systems where biological degradation is the soul of the process is

the abundance of nutritional sources. Therefore, the destruction or inhibition of a bacterial

group or even a bacterial community will not necessarily disrupt the entire process. As other

functionally-similar bacterial communities will compensate and possibly even flourish farther

as the impacted bacteria may serve as an additional nutritional source, further enriching the

media. Therefore, at concentrations anticipated to occur in such systems the impacts will likely

be rapidly accommodated by the system.

Objective 3: Lifecycle assessment of a commercially available nanosilver solution commonly used for the disinfection of dental unit waterlines. Focusing on the potential physicochemical transformations, as well as, fate of the AgNPs.

The research presented thus far was intended to advance the fundamental understanding of the potential environmental implications of AgNPs on complex aerobic and anaerobic systems that rely on biodegradation for waste processing. The environmental concerns associated with AgNPs were assessed focusing on vital system parameters which are routinely used to detect disturbances within environmental systems while also concentrating on the fate, mode of interaction, transport and transformations of the AgNPs. Hence, having the capacity to control the different

187 characteristics of the AgNPs through synthesis was essential for this purpose. However, the AgNPs reportedly most likely to be released into various environmental systems are mainly from consumer products.

Therefore, a commonly used AgNP-containing consumer product commonly used at dental offices for disinfection purposes was investigated. This research was conducted with a focus on the fate, mode of interaction and physicochemical transformations of the AgNPs under simulated conditions. The work accomplished prior to this point, as well as available literature, clearly suggests, understanding the transformations and surface modifications of AgNPs under environmental conditions to be critical for predicting mobility and overall expected toxicity.

Key Findings:

 The commercial nanosilver solution is an effective disinfecting agent.

 The pristine AgNPs were spherical with diameters ranging from 3 to 5 nm, bound by a

stabilizing polymer. The particle surface was predominantly silver oxide (Ag) with miniscule

portions of Ag (I) and Ag (III).

 After the disinfection process, the polymeric stabilizing agent surrounding the AgNPs

disappeared, most likely due to adsorption on the biofilm surface. The absence of the capping

agent initiated aggregation of a fraction of the AgNPs, producing aggregates with sizes ranging

from 50 to 200 nm. The surface speciation of the nanoparticles was transformed to AgCl.

 AgNPs may go through transformations during the usage scenario; these transformations must

be considered before making judgements about to their environmental toxicity to natural or

engineered systems.

188

Objective 4: Fate and transportation of silver nanoparticles released from consumer products: A review on ecological risk assessments.

In an effort to better understand the fate, transport and impacts of AgNPs under various environmental scenarios, many studies have been conducted in the past decade. However, research efforts have been mainly focused on pristine AgNPs, either lab-synthesized or purchased as a commercially available material. Although these studies have contributed greatly to a better fundamental understanding of the AgNPs in terms of mechanisms of toxicity and behavior under various environmental conditions, they have not addressed the real environmental concerns associated with AgNPs, which are most likely to occur through their release by consumer products.

To achieve a more accurate environmental impact assessment of silver nanoparticles, it is important to consider the majority of AgNPs in consumer products are either integrated into or embedded on to the surface of the specific material, and therefore are seldom available as pure materials. This chapter demonstrates the importance of studies on actual consumer products in an effort to address the current state of research. A comprehensive evaluation of the previous research on characterization techniques, routes of environmental exposure, usage scenarios and potential ecological risks of AgNPs from consumer products was conducted. Furthermore, toxicity assays currently available for environmental risks assessments were identified. The objectives of this review were to summarize the major findings of past research related to AgNPs released from CPs, while identifying the issues and knowledge gaps and providing recommendations for future research.

Key Findings:

 During their lifecycle (e. g., cradle to grave), AgNPs in consumer products are naturally

exposed to different conditions. The impact of various environmental parameters on AgNPs

189

have proven to initiate a range of physicochemical transformations. These alterations may

potentially alter the characteristics (e. g., toxicity) of AgNPs, and therefore, must be considered

for accurate environmental impact assessments.

 The effects of the manufacturing process, considering the method of AgNPs incorporation into

consumer products on AgNPs release and ecological impact scenarios must be further

investigated and considered.

 Due to the instability of AgNPs (e. g., aggregation) caused by their high surface area to volume

ratio, a wide variety of chemicals are used as stabilizers known as capping agents. Citrate,

Tween 20, PVP, BPEI, and polyethlene glycol (PEG) are some of the most widely used capping

agents for the synthesis of AgNPs. The physicochemical characteristics of AgNPs are closely

related to the capping agents used in their synthesis and therefore, may vary greatly. Although

coated AgNPs are commonly incorporated into CPs, most studies did not mention or report

any details on the type, characteristics or potential effects of the coating agents incorporated

into the products.

190

Chapter 7

Recommended Future Work

191

This chapter is a discussion of recommended future work to achieve a more comprehensive understanding of the fate, transport and environmental impacts of AgNPs. The last section of this chapter describes research currently in progress.

 Characterization of Silver Nanoparticles in Consumer Products

The regulatory aspects of nanotechnology are far behind the wide scale manufacturing and application of nanomaterials. Information regarding the characteristics (e. g., size, shape, surface charge, surface chemistry and synthesis method), production quantities, product incorporation techniques and demographic distribution of silver nanoparticles used in consumer products is scarce. However, such data are necessary for conducting risk assessments, lifecycle analyses, and implementation of environmental exposure guidelines. Currently, there are no regulations are in place requiring manufacturers to identify the types of ENMs used in consumer products. In conjunction with further developments in analytical methods and instrumentations for detection and characterization of ENMs, regulations mandating manufacturers to providing information pertaining to ENMs-CPs may be necessary.

 Characterization of AgNPs in Complex Environmentally Relevant Matrices

While analytical methods and instrumentation are available for the characterization of ENMs in simple matrices, far fewer methods have been established for the characterization of ENMs in complex environmental matrices such as biosolids, wastewater, landfill leachate and soils. The development of methods for ENM separation, detection and characterization (concentrations, morphological characteristics, speciation and surface transformations) in complex environmentally relevant matrices is crucial.

192

 Pristine vs Aged AgNPs

Research on the potential environmental risks associated with the production and applications of

AgNPs has focused on pristine, fresh, lab-synthesized AgNPs. However, most of the AgNP- containing CPs are likely to go through a series of events before they may be released into the environment. Depending on the form of product integration, usage scenario and surrounding environmental conditions of the AgNPs, fate-determining transformations such as aggregation, dissolution and complexations are likely. Such transformations may significantly affect the physicochemical properties of AgNPs, leading to uncertainties regarding their condition when they are released to the environment (3). As an example, according to research presented in this dissertation silver nanoparticles are prone to surface transformations in the presence of environmentally-abundant elements and anions such as oxygen, chloride and sulfide. The research presented here, combined with research conducted by others, emphasize the importance of a complete and relevant life-cycle assessment of nanomaterial containing CPs. The point to be evaluated is whether ENMs released from CPs (after usage, storage and aging) show differences in reactivity or toxicity relative to the pristine, newly-synthesized materials after usage or aging

(1). To attain a more precise and comprehensive understanding of the potential ecological impacts of AgNPs the potential transformations must be considered.

 Routes of Environmental Exposure

Further research is required to better understand the points in which ENMs, specifically nanosilver, are most likely released from the mother product into environmental compartments. In general,

AgNPs are either integrated into or embedded onto the surface of a product or matrix prior to being commercially available for consumers. Every year increased numbers of AgNPs containing CPs are manufactured to meet an increasingly wide range of applications. In an effort to expand the

193 regulatory governance, identification and consideration of the pathways commonly traveled prior to environmental exposure is essential. Further investigations of the specific usage scenarios associated to different nano-enabled consumer products during their lifecycle are necessary.

 Long-term Impact Assessments

Most of the environmental assessment research performed on ENMs have neglected the potential long-term impacts of ENMs in the environment. The priority of this thesis was to assess the potential direct impacts of AgNPs on one cycle of the biodegradation process occurring in various systems. Further exposure studies assessing the long term impacts of AgNPs are necessary to better understand the long-term effects of environmental exposure to AgNPs.

 Investigation of The Impact of Sample Preparation on ENM’s Data Quality

The physical and chemical properties distinguishing nanomaterials from their bulk counterpart are directly attributed to the particle morphology (size and shape) and only occur within a finite size range. To investigate the environmental fate, transport and implications of nanomaterials, fundamental data about the characteristics of the nanomaterial must be obtained. Several characterization techniques are widely employed to fulfill this purpose including X-Ray

Diffraction (XRD), and Electron Microscopy (EM) techniques such as Transmission Electron

Microscopy (TEM) and Scanning Electron Microscopy (SEM). These techniques are generally used to determine the morphological characteristics (size and shape) of the nanomaterials. In the study of nanomaterials, although all characterization techniques carry considerable importance,

EM techniques are of special interest as they provide information on the size of the nanomaterials.

Proper sample preparation is a vital step to any microscopic analysis, especially so for TEM imaging. Sample preparation for TEM analysis involves loading of the samples onto a TEM grid.

TEM grids are generally constructed from different materials such as copper, nickel, gold, etc.

194

The metallic base of the grid is commonly coated. As a result, interactions between the grid and the nanomaterial are possible, potentially impacting the morphology of the nanomaterial under investigation. Other manipulations are theoretically possible during sample preparation in the presence of atmospheric oxygen or hydrogen sulfide. If this hypothesis holds, the data quality obtained relating to the morphological characteristics of the ENMs may be significantly degraded.

Further research studying the impact of grid composition (grid type/coating) and or existence of atmospheric gases such as oxygen and H2S on the morphology (size and shape) of the ENMs is essential. The results of this research will shed light on certain provisions that ought to be taken into consideration when preparing nanosilver samples (AgNPs) for TEM analysis.

195

Literature Cited

1. Nowack, B.; Ranville, J. F.; Diamond, S.; Gallego-Urrea, J. A.; Metcalfe, C.; Rose, J.; Horne, N.; Koelmans, A. A.; Klaine, S. J. Potential scenarios for nanomaterial release and subsequent alteration in the environment. Environ. Toxicol. Chem. 2012, 31, 50-59.

2. Nowack, B.; Boldrin, A.; Caballero, A.; Hansen, S. F.; Gottschalk, F.; Heggelund, L.; Hennig, M.; Mackevica, A.; Maes, H. M.; Navratilova, J.; Neubauer, N.; Peters. R. J. B.; Rose, J.; Schaeffer, A.; Scifo, L.; van Leeuwen, S.; von der Kammer, F.; Wohlleben, W.; Wyrwoll, A. J.; Hristozov, D. Meeting the needs for aged and released nanomaterials required for further testing-the SUN approach. Environ. Sci. Technol. 2016.

3. Mitrano, D. M.; Motellier, S.; Clavaguera, S.; Nowack, B. Review of nanomaterial aging and transformations through the life cycle of nano-enhanced products. Environment International. 2015, 77, 132-147.

196