The Effects of Silver Nanoparticles on Lower Trophic Levels in Aquatic Ecosystems

THE EFFECTS OF SILVER NANOPARTICLES ON LOWER TROPHIC LEVELS IN AQUATIC ECOSYSTEMS

A Thesis Submitted to the Committee on Graduate Studies in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy in the Faculty of Arts and Science

TRENT UNIVERSITY Peterborough, Ontario, Canada © Copyright by Katarina Ana Cetinic 2019 Environmental and Life Sciences Ph.D. Graduate Program May 2019

ABSTRACT

The effects of silver nanoparticles on lower trophic levels in aquatic ecosystems

Katarina Ana Cetinic

Due to their effective antibacterial and antifungal properties, silver nanoparticles

(AgNPs) have quickly become the most commonly used nanomaterial, with applications in industry, medicine and consumer products. This increased use of AgNPs over the past decade will inevitably result in an elevated release of nanoparticles into the environment, highlighting the importance of assessing the environmental impacts of these nanomaterials on aquatic ecosystems. Although numerous laboratory studies have already reported on the negative effects of AgNPs to freshwater organisms, only a handful of studies have investigated the impacts of environmentally relevant levels of AgNPs on whole communities under natural conditions. This thesis examines the effects of chronic

AgNP exposure on natural freshwater littoral microcrustacean, benthic macroinvertebrate and pelagic zooplankton communities. To assess the responses of these communities to

AgNPs, I focused on a solely field-based approach, combining a six-week mesocosm study with a three-year whole lake experiment at the IISD – Experimental Lakes Area

(Ontario, Canada). Our mesocosm study tested the effects of AgNP concentration (low, medium and high dose), surface coating (citrate- and polyvinylpyrrolidone [PVP]-coated

AgNPs), and type of exposure (chronic and pulsed addition) on benthic macroinvertebrates in fine and stony sediments. Relative abundances of metal-tolerant

Chironomidae in fine sediments were highest in high dose PVP-AgNP treatments; however, no negative effects of AgNP exposure were seen on biodiversity metrics or overall community structure throughout the study. I observed similar results within the

ii whole lake study that incorporated a long-term addition of low levels of AgNPs to an experimental lake. Mixed-effects models and multivariate methods revealed a decline in all species of the littoral microcrustacean family Chydoridae in the final year of the study within our experimental lake, suggesting that this taxon may be sensitive to AgNP exposure; however, these effects were fairly subtle and were not reflected in the overall composition of littoral communities. No other negative effects of AgNPs were observed on the pelagic zooplankton or benthic macroinvertebrate communities. My results demonstrate that environmentally relevant levels of AgNPs have little impact on natural freshwater microcrustacean and benthic macroinvertebrate communities. Instead, biodiversity metrics and community structure are primarily influenced by seasonal dynamics and nutrient concentrations across both lakes. This thesis highlights the importance of incorporating environmental conditions and the natural variability of communities when examining the potential risks posed by the release of AgNPs into the environment, as simplistic laboratory bioassays may not provide an adequate assessment of the long-term impacts of AgNPs on freshwater systems.

Keywords: Silver nanoparticles, zooplankton, benthic macroinvertebrates, littoral microcrustaceans, whole lake experiment, IISD – Experimental Lakes Area

iii Acknowledgements

I would like to express my sincere gratitude to my supervisor Dr. Maggie

Xenopoulos for allowing me the incredible opportunity to conduct large-scale freshwater research, and providing me with endless patience, support and guidance along the way. I would also like to express special thanks to my committee members, Dr. Dave Beresford and Dr. Gary Burness, for their constant words of encouragement and support, and helpful comments and advice in committee meetings and in the writing stages of my thesis. Thank you also to Dr. Jim Schaefer and Dr. James Rusak for being wonderful members of my defence committee, and for providing me with insightful questions and a broader perspective from which to examine my research.

I would like to extend a special thank you to all of the scientists and staff at the

International Institute for Sustainable Development – Experimental Lakes Area for their assistance with the Lake Ecosystem Nanosilver (LENS) project. In particular, I would like to thank Dr. Mike Paterson for his continuous support and mentorship throughout the years. I would also like to acknowledge Dr. Beth Norman for her helpful suggestions and contribution towards the experimental design in the first years of this project. Many thanks to everyone who helped with sample collection and analysis throughout the years:

Jennifer Vincent, Graham Blakelock, Jonathan Martin, Daniel Rearick, Andrea Conine,

Andrew Scott and Paul Finigan. And lastly, to all those at the IISD – ELA that have made every field season so special and memorable. Some of my favourite moments ever were spent up there with you.

A warm thank you to Dr. Tom Whillans and Dr. Stephen Hill, for not only being extraordinary professors to TA for, but for providing a constant source of inspiration and

iv support. I would also like to thank my Master’s supervisor, Dr. Petar Kružić, for continuing to be a wonderful mentor, even after all these years. Thank you for everything you have taught me, and for helping me grow as a researcher.

I extend my sincere gratitude to Linda Cardwell and Mary-Lynn Scriver. The encouragement, patience and support you have shown me over the years have helped me get through many difficult and challenging moments, and I am beyond grateful.

Thank you to Holly Kuin and Heathyr Francis at the Centre for Academic Testing for being such amazing supervisors. Your enthusiasm and positive attitude have made working at the Centre a wonderful experience.

I am so thankful to all my friends in Canada and Croatia for always being there for me, even when they were thousands of kilometers away. Thank you to Samantha

Stephens, Allison Kwok, and Debbie Jenkins for their endless encouragement, especially in the moments leading up to my defence. Your passion, drive and strength are truly inspirational. I would also like to thank Adrian Borlestean for being an exceptional friend. Our road trips and engaging discussions on every topic imaginable are something

I will never forget. You have taught me so much about life, have been there for me and supported me in countless ways over the years, and I feel so fortunate to have you in my life.

I am especially grateful to my sister Marija and her husband Jeff for all their support and love. You have inspired me immensely, both academically and personally.

Thank you for always being there, and for all the adventures and moments spent together.

Having you in Canada while I pursued my degree has made all the difference.

v Thank you to my beautiful nieces and nephews for their boundless curiosity and wonder. Your interest and presence have taught me so much about what truly matters in life.

Lastly, I would like to thank my family in Croatia, especially my mother Biserka and my father Boris, for their endless love and encouragement over the years. You have been there for me, supported me and believed in me in every way imaginable, and I am endlessly grateful. I couldn’t have done it without you.

vi TABLE OF CONTENTS

PAGE TITLE PAGE ...... i ABSTRACT ...... ii KEYWORDS ...... iii ACKNOWLEDGEMENTS ...... iv TABLE OF CONTENTS ...... vii LIST OF TABLES ...... xi LIST OF FIGURES ...... xiii LIST OF ABBREVIATIONS ...... xviii

Chapter 1: General Introduction ...... 1

1.1. Background ...... 1 1.2. Objectives ...... 4 1.3. Novelty and Importance ...... 6 1.4. References ...... 8

Chapter 2: The effects of silver nanoparticles on natural benthic macroinvertebrate communities in littoral mesocosms

2.1 Preface ...... 15 2.2 Abstract ...... 16 2.3 Introduction ...... 17 2.4 Materials and Methods ...... 21 2.4.1 Study Area ...... 21 2.4.2 Mesocosms ...... 21 2.4.3 Mesocosm treatments and sampling ...... 22 2.4.4 Characterization and fate of silver nanoparticles ...... 23 2.4.5 Analysis of water quality parameters ...... 24 2.4.6 Benthic macroinvertebrate community collection and analysis .. 25 2.4.6.1 Stony sediment samples ...... 25 2.4.6.2 Fine sediment samples ...... 25 2.4.7 Benthic macroinvertebrate enumeration and identification ...... 25 2.4.8 Statistical analysis ...... 26 2.5 Results ...... 28 2.5.1 AgNPs and water chemistry ...... 28

vii 2.5.2 Macroinvertebrate community composition in fine sediment samples ...... 28 2.5.3 Macroinvertebrate community composition in stony sediment samples ...... 29 2.5.4 Multivariate analyses ...... 30 2.6 Discussion ...... 31 2.7 References ...... 37 2.8 Tables and Figures ...... 45

Chapter 3: Benthic macroinvertebrate community responses to an addition of silver nanoparticles: A whole lake ecosystem approach

3.1 Preface ...... 54 3.2 Abstract ...... 55 3.3 Introduction ...... 56 3.4 Methods ...... 59 3.4.1 Study area ...... 59 3.4.2 Silver nanoparticles and whole lake dosing ...... 59 3.4.3 Collection and processing of lake water samples ...... 60 3.4.4 Ag analysis ...... 62 3.4.5 Benthic macroinvertebrate sampling and taxonomic analyses ... 62 3.4.6 Statistical analysis ...... 64 3.5 Results ...... 66 3.5.1 Benthic macroinvertebrates ...... 66 3.5.2 Multivariate analyses ...... 68 3.6 Discussion ...... 69 3.7 References ...... 75 3.8 Tables and Figures ...... 85

Chapter 4: Responses of natural littoral microcrustacean communities to a whole lake addition of silver nanoparticles

4.1 Preface ...... 95 4.2 Abstract ...... 96 4.3 Introduction ...... 97 4.4 Methods ...... 100 4.4.1 Study area ...... 100 4.4.2 AgNP characterization and whole lake addition ...... 100

viii 4.4.3 Collection and processing of lake water and AgNP samples ...... 101 4.4.4 Littoral microcrustacean sampling and taxonomic analyses ...... 101 4.4.5 Periphyton analysis ...... 102 4.4.6 Statistical analysis ...... 103 4.5 Results ...... 105 4.5.1 Water quality parameters ...... 105 4.5.2 Periphyton biomass and Ag accumulation ...... 106 4.5.3 Littoral microcrustacean community structure ...... 106 4.5.3.1 Microcrustacean density and diversity ...... 106 4.5.3.2 Multivariate analyses ...... 109 4.6 Discussion ...... 110 4.7 References ...... 116 4.8 Tables and Figures ...... 125

Chapter 5: Effects of low concentrations of silver nanoparticles on natural zooplankton communities in a small boreal lake

5.1 Preface ...... 135 5.2 Abstract ...... 136 5.3 Introduction ...... 137 5.4 Methods ...... 139 5.4.1 Study area ...... 139 5.4.2 Whole lake AgNP addition ...... 140 5.4.3 Collection and processing of lake water samples ...... 140 5.4.4 Zooplankton sampling and laboratory analyses ...... 141 5.4.5 Statistical analysis ...... 142 5.5 Results ...... 144 5.5.1 Physico-chemical variables ...... 144 5.5.2 Zooplankton biomass and Ag concentration ...... 144 5.5.3 Zooplankton diversity and species composition ...... 145 5.5.4 Community analyses ...... 147 5.6 Discussion ...... 148 5.7 References ...... 153 5.8 Tables and Figures ...... 161

ix Chapter 6: General Discussion ...... 173

6.1 Future Work ...... 176 6.2 References ...... 178

Appendix ...... 183

x List of Tables

Chapter 2 Table 2.1 Results of one-way ANOVA comparing the effects of chronic PVP-AgNP exposure (control, low, medium and high PVP-AgNP), dosing regimen (control, chronic and pulsed addition) and surface coatings (control, citrate and PVP) on benthic macroinvertebrate abundance, taxa richness and relative abundance of Chironomidae in fine sediment samples following an addition of AgNPs to freshwater mesocosms in Lake 239 at the IISD – Experimental Lakes Area...... 55 Table 2.2 Results of one-way ANOVA comparing the effects of chronic PVP-AgNP exposure (control, low, medium and high PVP-AgNP), dosing regimen (control, chronic and pulsed addition) and surface coatings (control, citrate and PVP) on benthic macroinvertebrate abundance, taxa richness and relative abundance of Chironomidae in stony sediment samples following an addition of AgNPs to freshwater mesocosms in Lake 239 at the IISD – Experimental Lakes Area...... 56 Table 2.3 Correlation coefficients (ρ) indicating the best subset of environmental variables related to patterns in macroinvertebrate community structure from the BIO- ENV procedure...... 57

Chapter 3 Table 3.1 Similarity percentages (SIMPER) analysis indicating which benthic macroinvertebrate taxa were primarily responsible for the observed dissimilarity (as indicated by ANOSIM results; R = 0.603, P = 0.001) in community composition over time (i.e. between years) in both our reference (Lake 221) and experimental lake (Lake 222)...... 93 Table 3.2 Results of the BIO-ENV analysis indicating which environmental variables were most strongly correlated with benthic macroinvertebrate assemblages...... 94

Chapter 4 Table 4.1 Results of the BIO-ENV analysis indicating which combination of environmental variables was best correlated with littoral microcrustacean community composition in our study...... 130

xi Chapter 5 Table 5.1 Similarity percentages (SIMPER) analysis indicating which zooplankton species were primarily responsible for the observed dissimilarity (as indicated by ANOSIM results) in community composition between our reference (Lake 221) and experimental (Lake 222) lake...... 163 Table 5.2 Results of the BIO-ENV analysis indicating which combination of environmental variables from a total of 10 were best correlated with zooplankton assemblage structure prior to and during a two-year (2014 – 2015) experimental addition of AgNPs to an experimental lake at the IISD – ELA...... 164

xii List of Figures

Chapter 2 Figure 2.1 Relative abundances of Chironomidae, Ephemeroptera-Plecoptera- Trichoptera (EPT) and other taxa across treatments in (A) fine sediment samples and (B) stony sediment samples following experimental AgNP exposure to freshwater mesocosms in Lake 239 at the IISD – Experimental Lakes Ares in summer 2012...... 60

Figure 2.2 Average benthic macroinvertebrate abundance (no. individuals -1 -1 mesocosm ) and taxa richness (no. taxa mesocosm ) for replicate chronic and pulse AgNP treatments in (A) and (C) fine sediment samples (n = 2), and (B) and (D) stony sediment samples (n = 2). Error bars are standard error...... 61 Figure 2.3 Non-metric multidimensional scaling (NMDS) ordination of benthic macroinvertebrate communities collected from fine and stony sediment samples across six treatment levels (n = 2 per treatment) following experimental AgNP exposure to freshwater mesocosms in Lake 239 at the IISD – Experimental Lakes Area in summer 2012...... 62 Figure 2.4 Non-metric multidimensional scaling (NMDS) ordination of benthic macroinvertebrate communities in (A) fine and (B) stony sediment samples in relation to environmental variables that were most strongly correlated with community structure. The direction and length of the environmental vectors are determined by the envfit() function in the vegan package...... 63

Chapter 3 Figure 3.1 Sampling locations along the littoral zone of the (A) reference (Lake 221) and (B) experimental lake (Lake 222). Centre buoy is marked with ×...... 98

Figure 3.2 Relative abundances (%) of Chironomidae, Ephemeroptera-Plecoptera- Trichoptera (EPT) and other benthic macroinvertebrate taxa across years in the (A) reference lake (Lake 221) and (B) experimental lake (Lake 222). AgNP additions to Lake 222 occurred throughout 2014 and 2015. Values are monthly averages across all sites (n = 3)...... 99

xiii Figure 3.3 Average monthly abundance (no. of individuals) of benthic macroinvertebrates in our reference (Lake 221; L221) and experimental lake (Lake 222; L222) at the IISD – Experimental Lakes Area over a three-year study period. Values are means of locations (n = 3) ± standard error...... 100 Figure 3.4 Average monthly taxa richness (no. of taxa) of benthic macroinvertebrates in our reference (Lake 221; L221) and experimental lake (Lake 222; L222) at the IISD – Experimental Lakes Area over a three-year study period. Values are means of locations (n = 3) ± standard error...... 100 Figure 3.5 Average monthly abundance (no. of individuals) of Chironomidae in rock bag samples in our reference (Lake 221; L221) and experimental lake (Lake 222; L222) at the IISD – Experimental Lakes Area over a three year study period. Values are means of locations (n = 3) ± standard error...... 101 Figure 3.6 Two-dimensional non-metric multidimensional scaling (NMDS) ordination based on the Bray-Curtis dissimilarity matrix calculated from benthic macroinvertebrate abundance data. The ordination illustrates differences in macroinvertebrate community composition in our experimental lake (Lake 222) and reference lake (Lake 221) over time prior to (2013) and during a two-year (2014 – 2015) experimental addition of AgNPs to Lake 222 at the IISD – ELA. Data points are sampling locations from 2013 to 2015; points that are closer together in ordination space are more similar in terms of macroinvertebrate community composition than those further apart. White symbols represent Lake 221; black symbols represent Lake 222. Ellipses represent ordination confidence intervals (95%)...... 102 Figure 3.7 Non-metric multidimensional scaling (NMDS) ordination of benthic macroinvertebrate communities in relation to environmental variables that were most strongly correlated with community structure...... 103

Chapter 4 -1 Figure 4.1 Temporal fluctuations in periphyton biomass (mg organic matter tile ) and -1 periphyton Ag concentrations (µg Ag tile ) in our experimental lake (Lake 222) during the years of AgNP exposure (2014 – 2015). Values are means of locations (n = 5) on each sampling date ± standard error...... 133

xiv Figure 4.2 Temporal fluctuations of diversity metrics (Shannon diversity (H’), species richness (S), and Pielou’s evenness (J’)) of littoral microcrustacean communities in the (A) reference lake (Lake 221) and (B) experimental lake (Lake 222) at IISD – ELA over a three year study period...... 134 -2 Figure 4.3 Average annual density (no. individuals cm ) of littoral microcrustaceans collected from tiles in our reference lake (Lake 221; L221) and experimental lake (Lake 222; L222) over a period of three years (2013 – 2015). Values are means of locations (n = 5) for each sampling date ± standard error...... 134

Figure 4.4 Relative abundances of littoral microcrustaceans (number of individuals cm-2) in the (A) reference lake (Lake 221) and (B) experimental lake (Lake 222) over the three-year duration of the experiment...... 135 Figure 4.5 Non-metric multidimensional scaling (NMDS) ordination displaying littoral microcrustacean species (based on the Bray-Curtis dissimilarity matrix) and fitted environmental variables that were best correlated to microcrustacean community structure (BIO-ENV procedure). The ordination is based on data collected from a reference (Lake 221) and experimental lake (Lake 222) at the IISD – ELA over a period of three years prior to (2013) and during AgNP addition (2014 – 2015)...... 136 Figure 4.6 Non-metric multidimensional scaling (NMDS) ordination based on the Bray- Curtis dissimilarity matrix of littoral microcrustacean abundances in our experimental (Lake 222) and reference lake (Lake 221). The ordination displays changes in community composition following a two-year (2014 – 2015) experimental addition of AgNPs to Lake 222 at the IISD – Experimental Lakes Area. Data points are sampling locations from 2013 to 2015. Black symbols represent Lake 221; white symbols represent Lake 222...... 137

Chapter 5 Figure 5.1 Temporal fluctuations in zooplankton biomass (mg dry weight) and -1 zooplankton Ag concentrations (ng Ag mg dry weight ) in our experimental lake (Lake 222) during the years of AgNP exposure (2014 – 2015). Values are means of sampling dates (June and August 2014: n = 3; July and August 2015: n = 4; July 2014 and June 2015: n = 5; October 2014 and 2015: n = 1) ± standard error...... 168

xv -1 Figure 5.2 Densities of the six most frequent zooplankton taxa (number of adults L ; mean ± standard error) recorded in our experimental (Lake 222) and reference lake (Lake 221) over a period of three years (2013 – 2015). Samples were collected from centre buoy throughout the ice-free season, and values are means of sampling dates (2013: n = 2; 2014 and 2015: n = 4)...... 169 Figure 5.3 Densities of immature stages of Copepoda (Cyclopoida and Calanoida -1 copepodites and nauplii) (number of individuals L ; mean ± standard error) recorded in our experimental (Lake 222) and reference lake (Lake 221) over a period of three years (2013 – 2015). Samples were collected from centre buoy throughout the ice-free season, and values are means of sampling dates (2013: n = 2; 2014 and 2015: n = 4)...... 170 -1 Figure 5.4 Average annual zooplankton density (no. individuals L ) in the pelagic zone of our reference (Lake 221; L221) and experimental lake (Lake 222; L222) from 2013 to 2015. Values are means of sampling dates (2013: n = 2; 2014 and 2015: n = 4) ± standard error...... 170 Figure 5.5 Temporal variation of diversity metrics (Shannon diversity (H’), species richness (S), and Pielou’s evenness (J’)) of zooplankton communities in the experimental (Lake 222) and reference lake (Lake 221) at IISD – ELA over a three year study period...... 171 Figure 5.6 Relative abundances of pelagic zooplankton families in the (A) experimental lake (Lake 222) and (B) reference lake (Lake 221) over the three-year duration of the experiment (2013 – 2015)...... 172 Figure 5.7 Non-metric multidimensional scaling (NMDS) ordination based on zooplankton diversity following a two-year (2014 – 2015) experimental addition of AgNPs to a lake at the IISD – Experimental Lakes Area. Data points are sampling dates from 2013 to 2015. White symbols represent the reference lake (Lake 221); black symbols represent the experimental lake (Lake 222)...... 173 Figure 5.8 Non-metric multidimensional scaling (NMDS) ordination with fitted environmental variables that were most strongly correlated to zooplankton community composition (BIO-ENV analysis). The ordination is based on species abundance data collected from a reference (Lake 221) and experimental lake (Lake 222) at the IISD –

xvi ELA over a period of three years prior to (2013) and during AgNP addition (2014 – 2015)...... 174

xvii List of Abbreviations and Symbols

Ag silver Ag+ silver ion AgNP(s) silver nanoparticle(s) ANOVA analysis of variance Chl-a chlorophyll a DAg dissolved silver DOC dissolved organic carbon e.g. Latin exempli gratia (for example) et al. Latin et alia (and others) EPT Ephemeroptera-Plecoptera-Trichoptera g gram GF/C glass fibre filter grade C GF/F glass fibre filter grade F H’ Shannon-Wiener diversity

HNO3 nitric acid ICP-MS inductively coupled plasma mass spectrometry i.e. Latin id est (in other words) IISD – ELA International Institute for Sustainable Development – Experimental Lakes Area J’ Pielou’s evenness kg kilogram L litre

LC50 lethal concentration required to kill 50 percent of the target population log10 logarithm base 10 L222 Lake 222 of the IISD – Experimental Lakes Area L221 Lake 221 of the IISD – Experimental Lakes Area m metre mg milligram ml millilitre

xviii no number ng nanogram nm nanometre NMDS Non Metric Multidimensional Scaling P-value test statistic for a hypothesis test pH -log H+ [mol] PVC polyvinyl chloride PVP polyvinylpyrrolidone S species richness sp. species spp. Latin species pluralis (multiple species) TAg total silver TDN total dissolved nitrogen TDP total dissolved phosphorus µg micrograms µm micrometres % percent µ micro ~ approximately < less than > greater than = equals

xix CHAPTER 1

General Introduction

1.1 Background

Nanoparticles are defined as materials with at least one dimension on a scale from

1 to 100 nm (ISO, 2008). Although naturally present within the environment in the form of colloids, clays or organic matter (Nowack & Bucheli, 2007), the development of nanotechnology has caused a rapid increase in the production and use of engineered nanoparticles over the past few decades. The unique physico-chemical properties of materials at the nanoscale differ from their larger counterparts, making them useful for a wide range of applications in various fields, including electronics, biomedicine, pharmaceuticals and cosmetics (USNTC, 2004; Navarro et al., 2008; Fabrega et al., 2011;

Reidy et al., 2013). This characteristic is due to an increased surface-to-volume ratio, which results in an excess of surface energy relative to that of the bulk material, rendering the nanoparticles highly reactive (Auffan et al., 2009). This is especially observable with silver, whose natural antimicrobial and antifungal properties become more effective when manipulated at the nanoscale, allowing it to act as a potent broad- spectrum biocide (Lara et al., 2011). As a result of this, engineered silver nanoparticles

(AgNPs) have found applications in industry, medicine and consumer goods, where their use is recognized as the most widespread among nanomaterials (Fabrega et al., 2011;

Luoma, 2008).

Risk assessments predict that AgNPs will enter the environment through the use, weathering and disposal of AgNP-containing products (Blaser et al., 2008), resulting in

AgNP concentrations within the part-per-billion (ppb) range in natural waters (Luoma,

1 2008; Maes et al., 2009). Because of its unique size-related characteristics, nanosilver can be toxic at very low concentrations, affecting not only bacteria (Sondi & Salopek-Sondi,

2004; Khaydarov et al., 2009; Das et al., 2012a,b), but also algae, plants, fungi (Navarro et al., 2008; Khaydarov et al., 2009; Das et al., 2014), aquatic invertebrates (Croteau et al., 2011; Asghari et al., 2012; Cleveland et al., 2012; Buffet et al., 2014; Bone et al.,

2015) and fish (Asharani et al., 2008). Although the majority of these studies have been conducted on single species using high concentrations of AgNPs under controlled laboratory conditions, with only little research examining the impacts of AgNP exposure on freshwater organisms in natural environments, these small-scale experiments can provide valuable insights into the mechanisms of AgNP toxicity.

The introduction of a contaminant such as nanosilver can alter the composition of freshwater communities, in turn affecting the structure and function of the entire lake ecosystem (Rhind, 2009; Johnston et al., 2014). Although much research has already been done examining the impacts of AgNPs on individual taxa in controlled laboratory environments, predictions based on these small-scale experiments are often unable to accurately assess what will occur at an ecosystem scale. The research presented in this thesis was conducted at the International Institute for Sustainable Development –

Experimental Lakes Area (IISD – ELA), located on the Canadian Shield near Kenora,

Ontario, Canada. This remote area is characterized by a number of relatively small, pristine lakes that have been used for ecosystem-scale experimentation and long-term monitoring for over 40 years (Enache et al., 2011).

2 Studying the responses of whole freshwater communities to AgNPs over longer time scales in a natural setting can help us better understand the impact AgNPs have on aquatic systems. This is especially important considering studies have shown that the distribution, bioavailability and toxicity of AgNPs largely depend upon the stability of the nanoparticles, which can be altered when the nanoparticles are released into the environment (Stebounova et al., 2011; Levard et al., 2012). Dissolved organic matter

(Lowry & Casman, 2009; Aiken et al., 2011), pH (Liu et al., 2012), ionic strength

(O’Brien & Cummins, 2009; Stebounova et al., 2011), presence of inorganic ligands

(Levard et al., 2012), as well as the size and shape of the nanoparticles (Stebounova et al.,

2011) are considered some of the most critical factors that can influence the properties of

AgNPs in the environment, therefore influencing their environmental behaviour and impact.

To increase the overall stability of the nanoparticles within environmental media, surface coatings can be added as stabilizers in the manufacturing process (Silva, 2011).

The most commonly used surface coatings for AgNPs are citrate, which electrostatically stabilizes the nanoparticles by forming a weak bond using the surface charge of the particles (Tolaymat et al., 2010), and polyvinylpyrrolidone (PVP), which creates a stronger bond with the particle core and sterically stabilizes the particles (Sun & Luo,

2005; Hassell et al., 2007; Tejamaya et al., 2012). However, the use of different coatings can also impact the fate and toxicity of AgNPs in the environment (Lu et al., 2010;

Tolaymat et al., 2010; Bone et al., 2012). For example, citrate-coated AgNPs have been shown to exhibit decreased stability and greater toxicity to freshwater organisms than

3 PVP-coated AgNPs due to the weaker citric bond formed with the nanoparticles themselves (Silva, 2011; Lead et al., 2014; Hou et al., 2017).

The main objective of this Ph.D. thesis is to explore the chronic effects of engineered AgNPs on lower trophic levels in freshwater systems. A particular focus is placed on microcrustaceans and benthic macroinvertebrates, as these communities play a pivotal role in aquatic food webs, acting as a structuring force for phytoplankton, and transferring energy and nutrients to higher trophic levels (Covich et al., 1999; Palmer,

1997; Vanni, 2002; Jeppesen et al., 2011). Being placed between top-down and bottom- up regulators results in zooplankton and zoobenthic assemblages reflecting changes in their physico-chemical environment, as well as alterations in the food web (Caroni &

Irvine, 2010; Hairston et al., 1995; Korosi & Smol, 2012), making them valuable indicators of ecosystem structure and function and ideal study organisms for the potential effects of AgNPs.

1.2 Objectives

The first part of my study uses a six-week mesocosm experiment to examine the effects of various AgNP concentrations, surface coatings and types of addition on the structure of natural benthic macroinvertebrate communities. In this chapter, I explore the effects of a range of AgNP concentrations (low, medium and high) as well as different surface coatings (citrate and PVP) on macroinvertebrate biodiversity metrics and community structure. I test the hypothesis that elevated concentrations of AgNPs will have a negative impact on benthic macroinvertebrate communities, decreasing overall taxa richness and abundances, and shifting community composition to one dominated by metal-tolerant taxa (e.g. Chironomidae, Ceratopogonidae, miscellaneous Diptera

4 [Zimmerman et al., 1993; Winner et al., 1980; Barbour et al., 1999; Beltman et al., 1999;

Loayza-Muro et al., 2010]). Further, I hypothesize that citrate-AgNPs will have a stronger adverse effect on benthic macroinvertebrate biodiversity metrics than PVP-

AgNPs. Finally, in this section I examine the impacts of a single high dose of PVP-

AgNPs on natural benthic macroinvertebrate communities. I hypothesize that a single pulsed addition of AgNPs to the mesocosms will alter macroinvertebrate community composition due to the inability of metal-sensitive taxa (e.g. Ephemeroptera, Plecoptera,

Trichoptera [Beltman et al., 1999; Qu et al., 2010]) to adapt to a sudden release of an elevated dose of AgNPs.

The second part of my study provides a more long-term analysis, examining the effects of low levels of AgNP exposure on natural littoral and pelagic microcrustacean and benthic macroinvertebrate communities over the course of three years, using a whole lake addition of AgNPs. In Chapter 3, I examine the chronic effects of AgNPs on benthic macroinvertebrate communities inhabiting stony sediments. As much as possible, I aim to incorporate multiple generations of macroinvertebrate taxa to obtain a broader picture of the impacts of AgNPs on overall community structure and dynamics over time. I hypothesize that exposure to AgNPs will lead to a decline in the overall density, diversity and taxonomic richness of these communities compared to a reference lake that was not exposed to AgNPs. Chapters 4 and 5 investigate the effects of AgNPs on natural littoral and pelagic microcrustacean communities. As littoral communities (i.e. organisms in the nearshore region associated with the surface of macrophytes or sediments) (Suthers &

Rissik, 2009) differ from pelagic communities (i.e. free floating organisms adapted to life in open waters) (Thorp & Covich, 1991; Suthers & Rissik, 2009) in species richness,

5 community composition, behaviour and functional role in the ecosystem (Chengalath,

1982; Julli, 1986; Thorp and Covich, 1991), it is crucial that we study these habitats and communities simultaneously and examine their responses to AgNP exposure. Here I hypothesize that a release of AgNPs into a freshwater lake will negatively impact metal- sensitive microcrustacean taxa (e.g. Holopedium gibberum, Daphnia spp., Sida crystallina, Bosmina longirostris [Yan & Strus, 1979; Marshall et al., 1981; Koivisto &

Ketola, 1995; Keller & Yan, 1998; Holt et al., 2003; Kennedy et al., 2010; Zhao & Wang,

2011; Palmer et al., 2013; Labaj et al., 2015]), which will be seen in a decrease in overall density, diversity and species richness in both littoral and pelagic habitats, as well as in altered community composition compared to a reference lake that was not exposed to

AgNPs.

Although our knowledge of engineered nanomaterials has greatly increased following their increased production and use, the environmental risk of the release of nanoparticles has yet to be fully understood. This thesis examines the potentially harmful effects of this newly emerging manufactured product on freshwater ecosystems, as well as the effects of a natural environment on the fate and bioavailability of AgNPs.

1.3 Novelty and Importance

My Ph.D research investigates the impact of silver nanoparticles on lower trophic levels in freshwater systems, a crucial and understudied area of ecotoxicology. With the increased use of AgNPs in consumer goods, industry and medicine, their ensuing release into the environment is inevitable (Benn & Westerhoff, 2008; Maes et al., 2009). It is critical, therefore, that we conduct studies in natural environments with environmentally

6 relevant levels of exposure to elucidate effects that cannot be anticipated in controlled laboratory studies. This whole ecosystem study will make a distinct contribution to understanding the risks associated with the release of AgNPs into the environment.

Moreover, the results of my research will aid decision-making by policymakers, as well as contribute significantly to the international and local scientific and regulatory community.

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14 CHAPTER 2

2.1 Preface

Title: The effects of silver nanoparticles on natural benthic macroinvertebrate communities in littoral mesocosms

Authors: Katarina A. Cetinic1, Paul Finigan2,3, Beth C. Norman2,4, Paul Frost2,

Marguerite A. Xenopoulos2

Contact Information:

1Environmental and Life Sciences Graduate Program, Trent University, Peterborough,

Ontario, Canada

2Department of Biology, Trent University, Peterborough, Ontario, Canada

3Present address: Fish and Wildlife Program, Fleming College, Lindsay, Ontario, Canada

4Present address: Department of Microbiology and Molecular Genetics, Michigan State

University, Michigan, USA

Corresponding author: Katarina A. Cetinic, [email protected], 705-748-1011 ext.

6467

Status: In preparation for submission

Author Contributions: KAC, BC, and MAX conceived and designed the experiments.

KAC conducted the experiments and processed the samples. KAC and MAX analyzed the data. KAC wrote the manuscript. MAX supervised the project. All authors provided critical feedback and helped with paper revisions.

Keywords: Silver nanoparticles, zooplankton, littoral microcrustaceans, benthic macroinvertebrates, Experimental Lakes Area, community composition, lake ecosystem

15 CHAPTER 2

The effects of silver nanoparticles on natural benthic macroinvertebrate

communities in littoral mesocosms

2.2 Abstract

Silver nanoparticles (AgNPs) are the most commonly used nanomaterial due to their antibacterial and antifungal properties. With the increased use of AgNPs and their subsequent release into the environment, there is a need to assess the effects of these nanomaterials on aquatic ecosystems. In this study, we examined the responses of benthic macroinvertebrates to AgNPs using replicated littoral mesocosms at the International

Institute for Sustainable Development – Experimental Lakes Area (Ontario, Canada). The effects of AgNP concentration (low, medium and high dose), surface coating (citrate- and polyvinylpyrrolidone [PVP]-coated AgNPs), and type of exposure (chronic and pulsed addition) were evaluated on invertebrates associated with two substrate types: fine sediments and stones. Chironomidae were the dominant taxon found across all treatments and accounted for more than 74% of benthic invertebrate community abundance in both substrates. Overall, total abundance of organisms was not affected by AgNP exposure across substrate types. Taxa richness (i.e. the total number of invertebrate taxa present; ranging from family to phylum), on the other hand, was impacted by surface coating in fine sediment samples, with exposure to citrate-coated AgNPs resulting in significantly higher richness than the control treatments. Furthermore, relative abundance of

Chironomidae was also significantly affected by chronic PVP-AgNP treatment and PVP-

AgNP dosing regimen, with high dose chronic PVP-AgNP treatments exhibiting the highest measured relative abundances of Chironomidae. Non-Metric Multidimensional

16 Scaling and analysis of similarity indicated no significant differences between macroinvertebrate community composition among treatments. Our findings highlight that, overall, at the temporal and spatial scales we investigated, AgNPs had little effect on the structure of benthic macroinvertebrate communities.

Keywords: Silver nanoparticles, mesocosms, benthic macroinvertebrates, Experimental

Lakes Area, community composition, lake ecosystem

2.3 Introduction

The development of nanotechnology has resulted in an increased production and application of engineered nanoparticles over the past few decades. At the nanoscale (1 –

100 nm in diameter; Katz et al., 2004), materials possess unique physico-chemical properties that differ from larger particles or ions (Navarro et al., 2008; Reidy et al.,

2013; USNTC, 2004). In the case of silver (Ag), its natural antimicrobial and antifungal properties are enhanced when engineered to the nanoscale (Lara et al., 2011). As a result, engineered silver nanoparticles (AgNPs) have applications in industry, medicine, and consumer goods, making them the most widespread nanomaterial (Fabrega et al., 2011;

Luoma, 2008). The manufacture and use of AgNPs in consumer products such as textiles, cosmetics and detergents is predicted to result in the release of AgNPs into the environment in concentrations within the ppb range in natural waters (Benn &

Westerhoff, 2008; Luoma, 2008; Maes et al., 2009).

The stability, and therefore toxicity of the nanoparticles can be altered when they are released into the environment and exposed to natural media (Stebounova et al., 2011;

Levard et al., 2012). The environmental behaviour and impact are affected by factors

17 such as pH, concentration of dissolved organic matter, ionic strength, presence of inorganic ligands and the physical characteristics of the nanoparticles themselves (e.g. size, shape and surface coating [Ju-Nam & Lead, 2008]). The majority of this research, however, has focused on the behaviour of AgNPs in the water column, and it is less clear how these factors relate to the behaviour of these nanoparticles in aquatic sediments, where other processes can come into play. For instance, bioturbation by benthic organisms may resuspend AgNPs that settle onto the surface of the sediments resulting in higher concentrations in overlying water (Amato et al. 2016) or changes in Ag complexes

0 (e.g. AgCl, Ag2S, Ag ) (Kaegi et al., 2011; Levard et al., 2012). Therefore, to better understand the environmental behaviour of AgNPs and the effects these nanoparticles have on benthic organisms, it is vital to examine their toxicity with associated aquatic sediments in natural settings.

Recent studies have shown that AgNPs can be toxic to a diverse group of aquatic organisms (Fabrega et al., 2011). Freshwater benthic communities may be especially susceptible to AgNPs, as these nanoparticles have been found to accumulate at the bottom of lakes (Lowry et al., 2012; Furtado et al., 2014). Despite this potential, there has been relatively little study of AgNPs and their effects on sediment-dwelling organisms in aquatic systems. The few field studies that have been conducted show that Nematoda

(Bone et al., 2015), Mollusca (Buffet et al., 2014) and Bivalvia (Cleveland et al., 2012) are highly sensitive to AgNP exposure in natural environments. Although research on the effects of AgNPs on other benthic invertebrate taxa is lacking, exposure to metals has generally been shown to have a toxic effect on certain taxa (e.g. Ephemeroptera,

Plecoptera, Trichoptera) (Beltman et al., 1999; Clements et al., 2000), suggesting that

18 they may be more metal-sensitive than others. On the other hand, some benthic taxa have been shown to increase in abundances in metal-contaminated environments (e.g.

Chironomidae, Ceratopogonidae, Diptera, Oligochaeta) (Zimmerman et al., 1993;

DeShon, 1995; Barbour et al., 1999; Beltman et al., 1999). It is important to note, however, that these effects depend on the metal and its concentration in the medium

(Bryan, 1971). Therefore, it is not yet clear whether these results from metal contamination studies will be useful to understand AgNP effects on benthic invertebrate communities exposed under environmental conditions.

We used a field-based experiment involving lake mesocosms at the International

Institute for Sustainable Development – Experimental Lakes Area (IISD – ELA)

(Ontario, Canada) to examine responses of benthic macroinvertebrates to AgNP exposure. Our first objective was to compare macroinvertebrate community responses across a gradient of PVP-coated AgNP additions. We hypothesized that benthic macroinvertebrate abundance and taxa richness would decrease with AgNP exposure in a dose-dependent manner, and that benthic communities exposed to AgNPs would experience a shift in community composition as a result of metal-tolerant taxa becoming dominant within the community. We also examined benthic macroinvertebrate responses between contrasting exposure regimens, i.e. between long-term press disturbances and a temporary pulse disturbance (hereinafter referred to as chronic and one-time pulsed addition, respectively). Acute AgNP exposure has been shown to negatively impact lower trophic levels. For example, Echavarri-Bravo et al. (2015) found that a single pulse exposure of AgNPs altered the functional diversity of benthic estuarine microbial communities, while Sørensen et al. (2016) reported acute immobility of Daphnia magna

19 following pulsed AgNP exposure. Further, Panacek et al. (2011) found that single pulsed exposure to AgNPs caused strong adverse effects on the development of Drosophila melanogaster (Diptera). We therefore hypothesized that pulse AgNP treatments would adversely affect benthic macroinvertebrate communities compared to control treatments, due to sudden exposure to high levels of AgNPs. Although Echavarri-Bravo et al. (2015) reported a complete recovery of the benthic microbial communities 120 hours after exposure, these results are unlikely to be seen in our study due to longer generations times of benthic macroinvertebrate organisms compared to bacteria. Lastly, we examined whether invertebrate community responses vary between similar exposure to different

AgNP surface coatings – citrate and polyvinylpyrrolidone (PVP). Surface coatings are used to stabilize nanoparticles and prevent agglomeration within a new medium (Sperling

& Parak, 2010). The most commonly used coatings for AgNPs are PVP and citrate

(Tolaymat et al., 2010). Citrate is electrostatically attracted to the nanoparticle core

(Sperling and Parak, 2010), and therefore more loosely attached to Ag than sterically- bound PVP, which may allow for easier dissolution, increased Ag+ release, and increased toxicity of citrate-coated AgNPs in water (Tejamaya et al, 2012). Given this, we hypothesized that citrate-AgNPs would have a stronger negative effect on benthic fauna than PVP-AgNPs, causing a greater decline in macroinvertebrate density and taxonomic richness, as well as a greater increase in relative abundances of Chironomidae compared to treatments exposed to PVP-AgNPs.

20 2.4 Materials and methods

2.4.1 Study area

To determine the effects of AgNPs on benthic macroinvertebrate community structure in a natural aquatic environment, we conducted a field mesocosm experiment in

Lake 239 (L239) of the International Institute for Sustainable Development –

Experimental Lakes Area (IISD – ELA). Lake 239 (49°39’ N, 93°43’ W) is an oligotrophic lake with a surface area of 56 ha, a mean depth of 10 m and a maximum depth of 30 m (Jansen and Hesslein, 2004). Epilimnetic means for environmental parameters measured from 1995 to 2011 were: dissolved organic carbon (DOC) 7.1 mg

L-1, dissolved oxygen (DO) 8.32 mg L-1, total phosphorus (TP) 5.56 µg L-1, total nitrogen

(TN) 300 µg L-1, pH 7.10, and conductivity (COND) 17.9 µS cm-1 (unpublished long- term ELA data).

2.4.2 Mesocosms

Twelve mesocosms were deployed in L239 during the summer of 2012.

Mesocosms were 2 m in diameter and used liners constructed of thick polyethylene

(Curry Industries, Winnipeg, Manitoba, Canada) that extended from ~20 cm above the surface to the lake sediments. Each mesocosm contained approximately 3000 L of lake water (Furtado et al, 2015).

The mesocosms were installed on 18 June 2012 with mesocosm depths varying between 1.5 m and 1.8 m on fine muddy substrate. Each mesocosm was fastened to a floating ring at the surface and was open to the sediments. Mesocosm liners were held to the sediments by sandbags, which minimized migration of mobile macrofauna. Following installation, all mesocosms were left undisturbed for more than 72 hours to allow for any

21 newly suspended material to settle prior to the addition of AgNPs. During this time, zooplankton collected from Lake 239 (0 – 20m depth) were added to the mesocosms in densities of approximately 310 individuals/L, and all leeches and fish were removed using nets and minnow traps.

2.4.3 Mesocosm treatments and sampling

We conducted three simultaneous AgNP experiments using these 12 mesocosms.

To test the effects of AgNP dose, we added PVP-AgNPs at four levels with two replicate mesocosms for each treatment level. Doses were selected to span a range of contamination at zero, low, medium and high doses and aimed to produce nominal concentrations of 0, 4, 16, and 64 µg L-1, respectively, by the end of six weeks. To reach these concentrations, 0, 0.89 ml, 3.56 ml and 14.24 ml of PVP-AgNP were added to the respective mesocosms every other day (“chronic addition”), starting on 23 June 2012, and continuing until 13 August 2012. A second experiment compared the effects of sustained exposure to two commonly applied surface coatings: PVP and citrate. To do so, citrate- coated AgNPs were added to two replicate mesocosms at the same level as the high-dose

PVP chronic mesocosms (14.24 ml every other day to reach a nominal concentration of

64 µg L-1). A third experiment examined the effects of dosing regimen, with PVP-AgNPs applied in a single dose of 240 ml to each of the final two replicated mesocosms, to achieve a nominal concentration of 64 µg L-1. The pulsed PVP-AgNP additions started on

11 July 2012, approximately two weeks after the start of the chronic AgNP additions.

After each addition, the mesocosms were carefully mixed using a vertical haul of a plastic disk (45 cm in diameter) to evenly distribute the AgNPs throughout the water column without disturbing the sediments.

22 Water quality parameters within the mesocosms were first sampled on 21 June

2012, three days before the first addition of AgNPs, and then weekly throughout the duration of the experiment. Samples were taken for total organic carbon (TOC), total dissolved nitrogen (TDN), nitrate (NO3-N), ammonium (NH4-N), total dissolved phosphorus (TDP), as well as total (TAg) and dissolved silver (DAg).

2.4.4 Characterization and fate of silver nanoparticles

The PVP and citrate-coated AgNPs used in the study were supplied from

NanoComposix, Inc. (San Diego, CA) in stock solutions of 1 mg Ag ml-1 nominal concentration. Nominal concentrations were confirmed by acid digestion and ICP-MS analysis at the Water Quality Centre at Trent University (Furtado et al., 2014).

Information provided by NanoComposix, Inc. indicated that PVP- and citrate-coated

AgNPs had an average diameter of 48.3 and 49.1 nm, a hydrodynamic diameter of 56.3 and 54.9 nm, and a zeta potential of -36.9 and -55.9 mV, respectively. Average AgNP diameter, hydrodynamic diameter and zeta potential were measured by the manufacturer via transmission electron microscopy (TEM) and dynamic light scattering (DLS) techniques.

Fate and characterization of the silver in the mesocosms have been described and published elsewhere (Furtado et al. 2014, 2015). For the current study we used three measures of Ag which were quantified using ICP-MS (X Series 2, Thermo Scientific).

For TAg analysis, 10 ml of 35 µm-filtered water was collected in 15 ml centrifuge tubes, acidified with 70% nitric acid (HNO3) to a total concentration of 4% and stored at 4ºC for further analysis. For DAg concentrations, water was filtered through a 1.2 µm polycarbonate filter, and then through a 0.2 µm polycarbonate filter prior to acidification,

23 and then acidified and stored similarly to TAg samples for later analysis. For sediment

Ag, two replicate cores were taken from different locations within each mesocosm at the end of the study. The cores were divided into three sections (0-2 cm, 2-4 cm and 4-6 cm), digested with concentrated HNO3, and analyzed using ICP-MS (X Series 2, Thermo

Scientific). For the purposes of this study, only Ag from the top layer of sediments (0-2 cm) was used. A detailed description of sample collection and laboratory analyses of

AgNPs within mesocosm water and sediments is given by Furtado et al. (2015).

2.4.5 Analysis of water quality parameters

For dissolved nutrients, 500 ml of pre-filtered water (35 µm) was filtered through a Whatman GF/F filter (0.7 µm pore size), and then subsequently through a 0.2 µm polycarbonate filter. Subsamples for TDP, NO3-N and NH4-N analysis were kept frozen at -20°C, while subsamples for TOC and TDN were stored in glass amber bottles and kept in the dark at 4°C. A detailed description of analytical methods for nutrient analysis can be found in Vincent et al. (2017). Temperature, dissolved oxygen, conductivity, pH, and concentration of dissolved nutrients (organic carbon, total nitrogen, ammonium, nitrate, and total phosphorus) were not significantly different between mesocosms over the course of the study (Furtado et al., 2015).

24 2.4.6 Benthic macroinvertebrate community collection and analysis

2.4.6.1 Stony sediment samples

As rocky substrates were absent from the mesocosms, we used precolonized rock bags deployed within each of the mesocosms to obtain samples of hard-bottom benthic macroinvertebrates (Resh, 1995). Cobbles of similar sizes (8 – 10 cm) were collected from a nearby sand bank and sewn into in 1-cm mesh bags. Thirty-six rock bags were placed in L239 on 14 June 2012, and allowed to colonize with native macrofauna over the course of 6 days. Three replicate rock bags were then gently placed at the bottom of each mesocosm on 20 June 2012, with care taken so as not to disturb the rock bags or soft sediments. On 10 August 2012, the rock bags were carefully lifted from the sediments and placed immediately in a collection net to avoid loss of individuals. All cobbles were gently scrubbed and invertebrates were preserved in 5% formaldehyde for enumeration and identification.

2.4.6.2 Fine sediment samples

A standard-size Ekman grab (6x6x6) was used to collect benthic macroinvertebrates from the fine substrate at the end of the study on 13 August, 2012.

Two replicate grab samples were taken from each mesocosm, within an area of approximately 3.14 m2. Each replicate consisted of one grab sample. Material collected in

Ekman grabs was transferred to 1L jars and stored in 70% ethanol until analysis.

2.4.7 Benthic macroinvertebrate enumeration and identification

Preserved stony and fine sediment samples were processed to determine abundance and community composition. All benthic samples were rinsed with tap water to remove excess silt or mud and sieved (ASTM no. 60; 250 µm). In stony sediment

25 samples, retained sieve material was processed in its entirety. Fine sediment samples were sub-sampled until 100 benthic macroinvertebrates were identified in each sample as described in Jones et al. (2004). If less than 100 individuals were present in the sample, all organisms were identified and enumerated. Organisms were sorted into 27 groups (a mix of phyla, classes, orders and families) as defined by Jones et al. (2004) using a Leica

S8 APO stereomicroscope (8:1 zoom with standard magnification of 10x – 80x).

2.4.8 Statistical analysis

Prior to analysis, average values were calculated for all variables that were measured within a single mesocosm (i.e. 2 sediment samples and 3 rock bags), in order to avoid pseudoreplication. Differences in benthic community structure between treatments were analyzed for total number of taxa (i.e. number of taxa per mesocosm), total number of individuals (i.e. number of individuals per mesocosm), and percent (%) contribution of the dominant taxon. For all calculations, a single (average) value was used per mesocosm. A one-way analysis of variance (one-way ANOVA) was used to compare total abundance of benthic macroinvertebrates, benthic macroinvertebrate taxa richness, and relative abundance of Chironomidae among treatments. To test differences between treatments for a) chronic PVP-AgNP concentrations (control, low, medium and high), b) exposure regimens (control, chronic addition and one-time pulsed addition), and c) surface coatings (control, citrate-AgNPs and PVP-AgNPs), separate one-way ANOVA analyses were performed, followed by Tukey’s HSD post hoc tests where applicable.

Prior to univariate analyses, environmental variables were log (x+1) transformed prior to analyses to satisfy the assumptions of normality.

26 Community composition was compared among treatments using multivariate statistical techniques. An analysis of similarity (ANOSIM) based on the Bray-Curtis dissimilarity matrix (Bray & Curtis, 1957) of benthic macroinvertebrate abundances was used to assess significant differences in community composition between AgNP treatments. Benthic macroinvertebrate abundance data were square root transformed and standardized (Wisconsin double standardization) prior to all multivariate analyses to reduce the effect of sample size for increased ordination quality (Oksanen, 2015). Non- metric Multidimensional Scaling (NMDS) analysis was used to visualize the differences in macroinvertebrate assemblages between treatments, allowing for the observation of dissimilarities among AgNP treatments based on the Bray-Curtis dissimilarity matrix

(Bray & Curtis, 1957) of benthic macroinvertebrate abundances. The BIO-ENV method

(Clarke & Ainsworth, 1993) was employed to determine which environmental variables

(TAg, TOC, TDP, TDN, chlorophyll-a, NO3-N, NH4-N) were important in structuring benthic macroinvertebrate assemblages. This method calculates Spearman rank correlation between Euclidean distances of environmental variables and Bray-Curtis similarities of macroinvertebrate abundance data to find the subset of environmental variables that best correlates with community structure. The community ecology package

‘vegan’ (Oksanen et al., 2015) was used for multivariate analyses. All statistical analyses were performed with the open source software package R (R Development Core Team,

2015), version 2.3-1.

27 2.5 Results

2.5.1 AgNPs and water chemistry

Water quality parameters did not vary significantly between treatments over the course of the study (Furtado et al., 2015). The average concentration for DOC across treatments was 8.71 mg L-1, while TDP and TDN measured 6.45 µg L-1 and 287.10 µg L-1,

-1 -1 respectively. Average concentrations of NH4 were 12.45 µg L , and NO3 6.95 µg L .

TAg measured in the water column was found to be dose-dependent, with average concentrations in the chronic PVP-AgNP treatments ranging from 1.09 µg L-1 in the controls to 19.27 µg L-1 in the high chronic PVP-AgNP treatments. High dose citrate treatments measured, on average, 19.90 µg L-1, whereas the highest average TAg concentrations (32.95 µg L-1) were recorded in the high pulse PVP-AgNP mesocosms

(Furtado et al., 2014). Analysis of sediment cores (0 – 2 cm) indicated that AgNPs accumulated in the mesocosm sediments, with concentrations increasing with AgNP loading rate (Furtado et al., 2014). On average, low chronic PVP-AgNP treatments contained (mean ± SD) 0.171 ± 0.009 µg Ag cm-3, medium chronic PVP-AgNP contained

0.659 ± 0.016 µg Ag cm-3, and high chronic PVP-AgNP treatments contained 0.886 ± 0.1

µg Ag cm-3 (Furtado et al., 2015).

2.5.2 Macroinvertebrate community composition in fine sediment samples

A total of 14 different taxa were identified in the fine sediment samples

(Appendix 1A). Benthic macroinvertebrate abundance varied on average between 211 and 322 individuals per treatment (Appendix 1A), with Chironomidae accounting for

74% to 87% of the total number of individuals within the benthic invertebrate community

(Fig. 2.1A). Other predominant taxa included Hydrachnida (up to 11%) and Nematoda

28 (up to 9%), while Gastropoda, Bivalvia, Ephemeroptera, Amphipoda, Turbellaria,

Oligochaeta, Culicidae and Ceratopogonidae were rarely found (<5% of the samples)

(Appendix 1A). Relative abundance of Chironomidae was significantly affected by the level of chronic PVP-AgNP dose (one-way ANOVA, F3,4 = 8.537, P = 0.033) and exposure regimen (one-way ANOVA, F2,3 = 16.73, P = 0.024) in fine sediment samples

(Table 2.1), with relative abundances of Chironomidae significantly higher in high dose chronic PVP-AgNP treatments than in controls (Tukey HSD, P = 0.029 and P = 0.022, respectively). We did not find the total abundance of macroinvertebrates to be significantly different between low, medium and high chronic PVP-AgNP treatments and the controls (Fig. 2.2A, Table 2.1). There was also no significant effect of AgNP exposure regimen (pulsed addition vs. chronic addition) or surface coating (high dose

PVP vs. high dose citrate) on benthic invertebrate abundance (Fig. 2.2A, Table 2.1).

Average taxa richness within the fine sediment samples ranged from 5.5 to 9 taxa per treatment (Fig. 2.2C). The low and high chronic PVP-AgNP treatments had the lowest average taxa richness (5.5 taxa), while the high dose pulse and citrate-AgNP treatments contained the most diverse community assemblages of all sites, with an average of 9 taxa identified (Fig. 2.2C). Surface coating was found to have an effect on macroinvertebrate taxa richness (one-way ANOVA, F2,3 = 12.70, P = 0.034, Table 2.1), with richness significantly higher in the high dose citrate-AgNP treatments compared to the high chronic PVP-AgNP treatments (Tukey’s HSD, P = 0.031).

2.5.3 Macroinvertebrate community composition in stony sediment samples

A total of 15 taxa were found in the stony sediment samples across all sites

(Appendix 1B). Average abundance varied from 28 to 48 individuals per treatment, with

29 Chironomidae as the most abundant taxa, making up between 75% and 82% of the individuals (Fig. 2.1B). Other frequently found taxa were Turbellaria (up to 9%),

Ephemeroptera (up to 7%), Gastropoda (up to 6%), Amphipoda (up to 5%). Anisoptera,

Oligochaeta, Hirudinea, Coleoptera, Zygoptera, Trichoptera, Decapoda, Hydra

(Coelenterata), Culicidae, and Ceratopogonidae were found only occasionally (<5% of the samples). Differences in relative abundances of Chironomidae did not depend on

AgNP treatment within stony sediment samples (Fig. 2.1B, Table 2.2), which may be due to the high variability between replicates. AgNP concentration did not affect benthic invertebrate abundance in chronic PVP-AgNP treatments (Fig. 2.2B, Table 2.2). Neither did AgNP exposure regimen (pulsed addition vs. chronic addition) or surface coating

(high dose PVP vs. high dose citrate) (Fig. 2.2B, Table 2.2).

Average taxa richness within the stony sediment samples ranged from 6.5 to 9 taxa per treatment. The low dose PVP-AgNP treatment contained the least diverse invertebrate community with an average of 6 taxa, while the communities in the control and high dose citrate-AgNP treatments were the most rich, with an average of 9 taxa,

(Fig. 2.2D). Taxa richness did not differ with chronic PVP-AgNP dose, surface coating or exposure regimen (Table 2.2).

2.5.4 Multivariate analyses

Multivariate analyses were carried out to determine differences in benthic community composition between mesocosm treatments, as well as examine relationships between environmental variables and patterns in community composition. Non-metric multidimensional scaling (NMDS) based on the community dissimilarity matrix revealed a distinct separation between macroinvertebrate communities collected from fine

30 sediment samples and those from rocky sediments (Fig. 2.3). However, no other groupings were observed within fine or stony sediment communities, indicating a high similarity in community composition across different AgNP treatments (control, low, medium and high PVP, citrate, and high pulse) (Fig. 2.3). ANOSIM results confirmed these observations, revealing significant differences between stony and fine sediment samples (ANOSIM, R = 0.968, P = 0.001), but revealing no differences in community composition between AgNP treatments within the fine sediments (ANOSIM, R = 0.139,

P = 0.202) or stony sediments (ANOSIM, R = -0.211, P = 0.795). The BIO-ENV procedure showed that of seven environmental variables, TAg, TDP and NO3-N were most strongly correlated with benthic community structure in fine sediment samples, whereas TOC and NO3-N best explained patterns in community structure in stony sediment samples (Fig. 2.4; Table 2.3).

2.6 Discussion

Overall, we found limited evidence to suggest widespread negative effects of environmentally relevant doses of AgNPs on benthic invertebrate abundance or community composition across both fine and stony sediment samples. We observed a significant effect of surface coating on macroinvertebrate taxa richness, as well as a strong effect of chronic PVP-AgNP treatment and exposure regimen on relative abundances of Chironomidae collected from fine sediment samples. However, AgNPs appeared to have no effect on benthic macroinvertebrate communities in stony substrates.

Similar findings were obtained through the BIO-ENV analysis, which showed that TAg played in important role in structuring macroinvertebrate communities in fine sediment

31 samples, whereas communities in stony sediment samples were not affected by TAg concentrations, and were instead impacted primarily by other physico-chemical variables measured in the mesocosms. These findings indicate that benthic macroinvertebrates dwelling in soft sediments may be more sensitive to AgNP exposure than those in stony sediments. The implications of these findings within freshwater ecosystems could be widespread, with alterations in benthic community composition modifying trophic interactions (Clements, 1999), as well as having a considerable impact on aquatic ecosystem function (Beltman et al., 1999; Clements, 1999).

Sediment can be a major sink for nanoparticles in aquatic environments (Lowry et al., 2012; Furtado et al., 2014), which renders benthic organisms susceptible to AgNP exposure. In this study, AgNPs showed a high affinity towards agglomeration and sedimentation in the mesocosms, accumulating in high concentrations (up to 5.6 µg g-1 dry weight) in the top layers of the sediments (Furtado et al., 2015). In addition to binding to particulate matter (Furtado et al., 2014), high concentrations of Ag in the sediments may have been due to adsorption by planktonic organisms and sedimentation following death, as bacterioplankton (Blakelock et al., 2016), phytoplankton and zooplankton (Vincent et al., 2017); all had measurable amounts of bound Ag. The concentration of Ag in the sediments in the chronic treatments was affected by surface coating, with concentrations 1.2 – 2.3 times higher in the high dose citrate-AgNP treatments compared to those in PVP-AgNP treatments (Furtado et al., 2015). As the citrate coating interacts with the nanoparticles through a relatively loose electrostatic bond (compared to sterically-bound PVP) (Sperling & Parak, 2010), we expected them to be more bioavailable and toxic to benthic invertebrate communities than PVP-coated

32 AgNPs. Despite higher concentrations of Ag found within the high dose citrate-AgNP treatments, there was no negative effect on benthic macroinvertebrate abundance, and taxa richness was significantly higher than in PVP-AgNP treatments across fine sediment samples. Our findings contrast with studies that have shown strong negative effects of citrate-AgNPs on both phytoplankton and zooplankton (Kennedy et al., 2010; Angel et al., 2013), and are instead similar to those of Park et al. (2015) who found no differential effect of surface coating on the benthic invertebrate Chironomus riparius

(Chironomidae), and Vincent et al. (2017) who similarly noted no effect on natural zooplankton communities.

Generally, in situ field studies examining the effects of heavy metals on benthic macroinvertebrates have reported strong detrimental effects on these communities, with effects ranging from reductions in overall density and biomass to decreases in species richness and changes in species composition (Plafkin et al., 1989; Beltman et al., 1999;

Maret et al., 2003; Jones et al., 2004; Qu et al., 2010). Although only few studies to date have examined the effects of AgNPs on whole benthic macroinvertebrate communities in natural freshwater ecosystems, a number of smaller-scale laboratory and mesocosm studies have found that individual species of benthic macroinvertebrates are highly sensitive to AgNP exposure (Croteau et al., 2011; Cleveland et al., 2012; Buffet et al.,

2014; Bone et al., 2015; Gambardella et al., 2015). Our findings therefore suggest that, under natural conditions, whole benthic macroinvertebrate communities may be less sensitive to AgNPs than to other metals.

Various factors could be contributing to a weaker toxic response of benthic invertebrates to AgNPs compared to those reported in other studies. Natural freshwater

33 ecosystems contain a large diversity of organic material in both the water column and sediments that can bind to silver ions (Ag+) (Fabrega et al., 2009). As the release of Ag+ from nanoparticles plays a significant role in the toxicity of AgNP exposure (Navarro et al., 2008; Fabrega et al. 2011; Levard et al 2012), the binding of these ions to organic matter in our field study may have reduced the bioavailability and toxicity of the AgNPs

(Angel et al., 2013). Additionally, AgNPs can be adsorbed to the surface of phytoplankton (Lei et al., 2017) and zooplankton (Asghari et al., 2012) or accumulated within the organisms themselves (McTeer et al., 2013; Tripathi et al., 2017). This accumulation of Ag in lower trophic levels or within fine particulate organic matter in the sediments can present an Ag-contaminated food source for higher-level predators such as benthic macroinvertebrates, which can effectively uptake Ag through their diet (Croteau et al., 2011; Cleveland et al., 2012; Oliver et al., 2014). In contrast, Wang et al. (2016) reported that, following internalization within algal cells, AgNPs underwent chemical transformation to silver sulphide (Ag2S), a form of silver that is biologically inert

(Hirsch, 1998). Furthermore, several studies have found that levels of accumulated contaminants in primary producers can be reduced through the development of new biomass (i.e. growth dilution) (Herendeen & Hill, 2004; Hill & Larsen, 2005). For example, Hill and Larsen (2005) showed that light and nutrients affected algal growth rates, in turn decreasing metal concentrations in the algae through growth dilution. This process can extend to higher trophic levels, reducing overall concentrations in predators

(Herendeen and Hill, 2004). If the benthic macroinvertebrates in our study were ingesting

Ag via trophic transfer, our lack of strong negative results suggests that perhaps biotransformation or growth dilution were at play.

34 Our ability to detect strong effects of AgNP exposure may have also been limited by the natural variability of benthic communities present within the littoral mesocosms, combined with the coarse taxonomic levels used in identification. As no significant results were observed at the higher taxonomic levels used in this study, these levels of identification may have been too coarse to detect any subtle effects of AgNPs. For example, different genera of Chironomidae larvae have been shown to exhibit varying levels of metal tolerance in aquatic environments. Swansburg et al. (2002) reported decreased genera richness and a shift in community composition towards metal-tolerant

Orthocladiinae, while metal-sensitive Tanytarsini (subfamily Chironominae) declined in abundance following exposure to metal mine drainage. Similar results were reported by

Mousavi et al. (2003), who found a higher relative abundance of Orthocladiinae in heavy metal contaminated lakes, while the least polluted lakes were dominated by the subfamilies Tanypodinae and Chironominae. The high proportion of Chironomidae found across all treatments in our study suggests that a more in-depth analysis (i.e. genus- or species-level identification) of these communities could give us a more accurate assessment of any subtle alterations in community structure (Arscott et al., 2006).

Additionally, the short preincubation times, combined with the limited taxonomic diversity of macroinvertebrates recorded within the mesocosms may have reduced our chances of detecting changes within the communities. The fairly low taxonomic richness would have limited the effect sizes on taxonomic richness and possibly diversity. It is also important to note that not all life stages of aquatic invertebrates were tested, as the number of generations per year (voltinism) varies among benthic taxa. The majority of freshwater benthic macroinvertebrates are univoltine (i.e. producing a single generation

35 per year; e.g. many Coleoptera, Ephemeroptera, Trichoptera, Amphipoda (Smock, 1988;

Naiman et al., 1998), so we were not able to capture a complete life cycle of these organisms within our experiment. If AgNP exposure had an effect on macroinvertebrate reproduction or egg development, we likely would not have been able to detect this due to the limited scope of our study.

Overall, this study shows that the results of acute toxicity tests of AgNPs might not be easily applied to freshwater ecosystems, as many factors in complex natural environments may mitigate the toxicity of AgNPs (Park et al., 2015). Further investigations on long-term community- and ecosystem-scale effects of AgNPs need to be done, as these studies will provide more insight on the effects of AgNP exposure on natural freshwater systems.

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44 2.8 Tables and Figures

Table 2.1 Results of one-way ANOVA comparing the effects of chronic PVP-AgNP exposure (control, low, medium and high PVP-AgNP), dosing regimen (control, chronic and pulsed addition) and surface coatings (control, citrate and PVP) on benthic macroinvertebrate abundance, taxa richness and relative abundance of Chironomidae in fine sediment samples following an addition of AgNPs to freshwater mesocosms in Lake 239 at the IISD – Experimental Lakes Area. df represents the degrees of freedom for the sources of variation. P-values in bold indicate a significant effect (P < 0.05).

df SS MS F P

Chronic PVP-AgNP exposure

Total abundance Treatment 3 1160 3868 0.159 0.919 Residuals 4 9742 24356 Taxa richness Treatment 3 2.625 0.875 1.556 0.331 Residuals 4 2.250 0.563 Chironomidae Treatment 3 0.007 0.003 8.537 0.033 abundance Residuals 4 0.001 0.0003

PVP-AgNP dosing regimen

Total abundance Treatment 2 8171 4085 0.135 0.878 Residuals 3 9047 30156 Taxa richness Treatment 2 3.583 1.792 1.654 0.328 Residuals 3 3.250 1.083 Chironomidae Treatment 2 0.007 0.003 16.73 0.024 abundance Residuals 3 0.001 0.0002

AgNP surface coating

Total abundance Treatment 2 5586 2793 0.088 0.918 Residuals 3 9470 31565 Taxa richness Treatment 2 10.58 5.292 12.70 0.034 Residuals 3 1.250 0.417 Chironomidae Treatment 2 0.011 0.005 4.466 0.126 abundance Residuals 3 0.004 0.001

45 Table 2.2 Results of one-way ANOVA comparing the effects of chronic PVP-AgNP exposure (control, low, medium and high PVP-AgNP), dosing regimen (control, chronic and pulsed addition) and surface coatings (control, citrate and PVP) on benthic macroinvertebrate abundance, taxa richness and relative abundance of Chironomidae in stony sediment samples following an addition of AgNPs to freshwater mesocosms in Lake 239 at the IISD – Experimental Lakes Area. df represents the degrees of freedom for the sources of variation. P-values in bold indicate a significant effect (P < 0.05).

df SS MS F P

Chronic PVP-AgNP exposure

Total abundance Treatment 3 377.9 126.0 0.465 0.722 Residuals 4 1083 270.8 Taxa richness Treatment 3 5.500 1.833 2.538 0.195 Residuals 4 2.889 0.722 Chironomidae Treatment 3 0.005 0.002 0.226 0.874 abundance Residuals 4 0.032 0.008

PVP-AgNP dosing regimen

Total abundance Treatment 2 444.0 222.0 1.225 0.408 Residuals 3 543.6 181.2 Taxa richness Treatment 2 4.000 2.000 1.009 0.462 Residuals 3 5.944 1.982 Chironomidae Treatment 2 0.003 0.002 0.131 0.882 abundance Residuals 3 0.038 0.013

AgNP surface coating

Total abundance Treatment 2 291.1 145.6 0.445 0.677 Residuals 3 981.8 327.3 Taxa richness Treatment 2 1.037 0.519 1.037 0.455 Residuals 3 1.500 0.500 Chironomidae Treatment 2 0.002 0.001 0.108 0.901 abundance Residuals 3 0.026 0.009

46 Table 2.3 Correlation coefficients (ρ) indicating the best subset of environmental variables related to patterns in macroinvertebrate community structure from the BIO- ENV procedure. Variables in bold indicate the best match.

Fine sediment samples Stony sediment samples

Best variable combination Rho (ρ) Best variable combination Rho (ρ)

TAg, TDP 0.074 TOC, NO3-N 0.022

TAg, TDP, NO3-N 0.091 TOC, NH4-N, NO3-N 0.020

TAg, TOC, TDP, Chl-a 0.057 TOC, TDP, NH4-N, NO3-N 0.005

TAg, TDP, Chl-a, NH4-N, TOC, TDP, TDN, NH4-N, 0.314 -0.057 NO3-N NO3-N

TAg, TOC, TDP, Chl-a, TAg, TOC, TDP, TDN, NH4- 0.012 -0.115 NH4-N, NO3-N N, NO3-N

TAg, TOC, TDP, TDN, TAg, TOC, TDP, TDN, Chl-a, -0.095 -0.230 Chl-a, NH4-N, NO3-N NH4-N, NO3-N

47 Figures and Captions

Figure 2.1 Relative abundances of Chironomidae, Ephemeroptera-Plecoptera-

Trichoptera (EPT) and other taxa across treatments in (A) fine sediment samples and (B) stony sediment samples following experimental AgNP exposure to freshwater mesocosms in Lake 239 at the IISD – Experimental Lakes Ares in summer 2012.

Figure 2.2 Average benthic macroinvertebrate abundance (no. individuals mesocosm-1) and taxa richness (no. taxa mesocosm-1) for replicate chronic and pulse AgNP treatments in (A) and (C) fine sediment samples (n = 2), and (B) and (D) stony sediment samples (n = 2). Samples were collected following AgNP addition to freshwater mesocosms in Lake 239 at the IISD – Experimental Lakes Area in summer 2012. Error bars are standard error. Note that the y-axes vary between panels.

Figure 2.3 Non-metric multidimensional scaling (NMDS) ordination of benthic macroinvertebrate communities collected from fine and stony sediment samples across six treatment levels (n = 2 per treatment) following experimental AgNP exposure to freshwater mesocosms in Lake 239 at the IISD – Experimental Lakes Area in summer

2012. Gray symbols represent fine sediment samples; black symbols represent stony sediment samples.

Figure 2.4 Non-metric multidimensional scaling (NMDS) ordination of benthic macroinvertebrate communities in (A) fine and (B) stony sediment samples in relation to environmental variables that were most strongly correlated with community structure.

The direction and length of the environmental vectors are determined by the envfit() function in the vegan package. Samples were collected from 12 freshwater mesocosms

48 following an experimental addition of AgNPs in summer 2012 at the IISD –

Experimental Lakes Area.

49 Chironomidae EPT Others A B 100% 100%

80% 80%

60% 60%

40% 40%

macroinvertebrates 20% 20% Relative abundance of benthic 0% 0%

Figure 2.1

50 FINE STONY A B 500 70 450 60 400 350 50 300 40 250 200 30 abundance 150 20 100 10 50

Average benthic macroinvertebrate 0 0

C D 9 8 8 7 7 6 6 5 5 4 4

richness 3 3 2 2 1 1 0 0 Average benthic macroinvertebrate taxa

Figure 2.2

51 Fine sediments + Stony sediments 0.5

0.0 +

+ NMDS2 + − 0.5

Control Low PVP Medium PVP

− 1.0 High PVP Non-metric fit R2 = 0.988 Citrate Linear fit R2 = 0.941 High Pulse Stress = 0.110

−1.0 −0.5 0.0 0.5 1.0 1.5

NMDS1

Figure 2.3

Ceratopogonidae A 0.4

Hydrachnidae TDP TAg 0.2 Bivalvia

Chironomidae 0.0 NMDS2

− 0.2 Gastropoda Nematoda Amphipoda

Non-metric fit R2 = 0.989 Linear fit R2 = 0.936 − 0.4 NO3−N Stress = 0.104

−0.5 0.0 0.5

NMDS1

Ephemeroptera B 0.2

TOC 0.1 Turbellaria Gastropoda 0.0 Chironomidae NMDS2 − 0.1 − 0.2 Non-metric fit R2 = 0.995 2 NO3−N Linear fit R = 0.974 Amphipoda Stress = 0.071 − 0.3

−0.4 −0.2 0.0 0.2 0.4

NMDS1

Figure 2.4

53 CHAPTER 3

3.1 Preface

Title: Responses of lower trophic levels to a whole lake addition of silver nanoparticles

Authors: Katarina A. Cetinic1, Michael Paterson2, Beth C. Norman3,4, Daniel C. Rearick1,

Paul Frost3, Marguerite A. Xenopoulos3

Contact Information:

1Environmental and Life Sciences Graduate Program, Trent University, Peterborough,

Ontario, Canada

2International Institute for Sustainable Development – Experimental Lakes Area,

Winnipeg, Manitoba, Canada

3Department of Biology, Trent University, Peterborough, Ontario, Canada

4Present address: Department of Microbiology and Molecular Genetics, Michigan State

University, Michigan, USA

Corresponding author: Katarina A. Cetinic, [email protected], 705-748-1011 ext.

6467

Status: In preparation for submission

Author Contributions: KAC, BC, and MAX conceived and designed the experiments.

Keywords: Silver nanoparticles, zooplankton, littoral microcrustaceans, benthic macroinvertebrates, Experimental Lakes Area, community composition, lake ecosystem

54 CHAPTER 3

Benthic macroinvertebrate community responses to an addition of silver

nanoparticles: A whole lake experimental approach

3.2 Abstract

To better understand the impacts of silver nanoparticles (AgNP) on benthic macroinvertebrate communities, we released environmentally relevant concentrations

(<10 µg L-1) of PVP-coated AgNPs to an entire lake at the Experimental Lakes Area

(Ontario, Canada). AgNPs were released into Lake 222 throughout the ice-free period in

2014 and 2015. Rock bags were placed along the littoral zone of the experimental (Lake

222) and reference (Lake 221) lakes to allow for the colonization of benthic macroinvertebrates, and removed on a monthly basis throughout summer and autumn in the year prior to, and during both years of silver addition. Although Ag was detected within the water column of the experimental lake at the time of collection, we observed no effect of AgNPs on benthic macroinvertebrate abundance, taxonomic richness (family to phylum) or community composition. Instead, we recorded a marked decline in taxa richness and overall abundance by the end of the study across both Lake 221 and Lake

222. Non-metric multidimensional scaling (NMDS) and analysis of similarity (ANOSIM) indicated significant differences in community composition across years in both lakes, suggesting that environmental factors, and not AgNP exposure, may be a factor influencing benthic macroinvertebrate community change across time. Overall, we found no measurable effects of AgNPs on benthic macroinvertebrates under environmentally

55 relevant conditions. Instead, these communities were influenced by seasonal dynamics and nutrient concentrations across both lakes.

Keywords: Silver nanoparticles, mesocosms, benthic macroinvertebrates, Experimental Lakes Area, community composition, lake ecosystem

3.3 Introduction

Silver nanoparticles (AgNPs) are recognized as the most commonly used nanomaterial in consumer products today (e.g. textiles, cosmetics, and detergents) due to their potent antimicrobial properties (Fabrega et al., 2011; Lara et al., 2011). Although current risk assessments predict fairly low concentrations of AgNPs entering surface waters (~ 0.116 – 0.764 ng/L) (Gottschalk et al., 2009) with the majority (>95%) being removed within wastewater treatment plants (Li et al., 2013), the increased production and use of AgNPs in consumer products may lead to higher volumes released into the environment (Sun et al., 2014).

The negative effects of AgNPs on lower trophic levels have been widely reported

(Asharani et al., 2008; Navarro et al., 2008; Khaydarov et al., 2009; Croteau et al., 2011;

Das et al., 2012; Das et al., 2014). Freshwater benthic communities (i.e. those that live in and on sediments) could be especially susceptible to the toxicity of AgNPs, as these nanomaterials have been shown to destabilize in natural media, settle out of the water column, and accumulate in the sediments of lakes (Lowry et al., 2012; Furtado et al.,

2014). Many benthic macroinvertebrate taxa including gastropods (Croteau et al., 2011), molluscs (Buffet et al., 2014), nematodes (Bone et al., 2015), crustaceans (Gambardella et al., 2015; Brittle et al., 2016), and echinoderms (Gambardella et al., 2015) have already

56 been found to be sensitive to AgNPs, with several mechanisms of toxicity suggested for benthic macroinvertebrate taxa (Croteau et al., 2011; Buffet et al., 2014; Bone et al.,

2015; Gambardella et al., 2015). The majority of studies on benthic macrofauna, however, describe the effects of AgNPs on individual taxa (Croteau et al., 2011; Buffet et al., 2014; Bone et al., 2015; Brittle et al., 2016), rather than assemblages or communities.

Additionally, only few studies have explored these effects under natural conditions

(Cleveland et al., 2012; Buffet et al., 2014) with no studies to date observing the long- term (i.e. multi-year) effects of AgNPs on whole freshwater benthic macroinvertebrate communities in situ.

Benthic macroinvertebrates are commonly used as indicators of environmental conditions due to their sensitivity to stressors and ability to provide responses that differ from natural variation (Plafkin et al., 1989). Specifically, exposure to metals (Hare et al.,

1992; Beltman et al., 1999) has been shown to alter benthic macroinvertebrate communities in predictable ways, decreasing taxonomic diversity (Plafkin et al., 1989;

Maret et al., 2003; Jones et al., 2004) and total abundance (Qu et al., 2010), and shifting community composition from metal-sensitive (e.g. Ephemeroptera) to more metal- tolerant taxa (e.g. Chironomidae) (Winner et al., 1980; Beltman et al., 1999; Chapter 2).

These studies suggest that changes in community-level responses can be used as reliable indicators of the impacts of metal contamination on natural benthic invertebrate communities. Additionally, as zoobenthic communities occupy a central position in aquatic food webs, influencing ecological processes such as nutrient cycling and energy flow (Covich et al., 1999; Palmer et al., 1997; Vanni, 2002), understanding the potential impact of AgNPs on macroinvertebrate communities is crucial, as changes in these

57 communities may have impacts on the structure and function of the entire lake ecosystem

(Covich et al., 1999).

Various physico-chemical factors in aquatic systems can influence the properties of nanoparticles, highlighting the importance of studying the effects of AgNPs in a natural setting. Studies have shown that the stability of AgNPs is highly sensitive to environmental factors such as pH (Liu et al., 2012), UV light (Poda et al., 2013), dissolved organic matter (Gunsolus et al., 2015), ionic strength (O’Brien & Cummins,

2009; Stebounova et al., 2010), and inorganic ligands (e.g. phosphates) (Xiu et al., 2011;

Levard et al., 2012). The release of AgNPs into natural environments can therefore have a strong impact on the properties of the nanoparticles (Stebounova et al., 2010; Levard et al., 2012), altering their bioavailability and toxicity (Park et al., 2015). Due to the complexity and variability of natural environments, ecologically relevant field-based studies are critical for assessing the effects of AgNPs on aquatic ecosystems.

The aim of this study was to identify changes in benthic macroinvertebrate communities in response to an addition of AgNPs in a natural environment. We hypothesized that an addition of AgNPs to a freshwater lake would negatively impact metal-sensitive benthic macroinvertebrate taxa (e.g. Ephemeroptera, Plecoptera,

Trichoptera) (Beltman et al., 1999; Qu et al., 2010), and cause a decline in overall macroinvertebrate abundance, diversity and taxa richness compared to a reference lake that was not exposed to AgNPs. To test this hypothesis, we conducted a whole ecosystem study at the Experimental Lakes Area in northwestern Ontario, Canada. Our approach incorporated an experimental lake to which we added low, environmentally relevant concentrations of AgNPs over the ice-free season of 2014 and 2015, and a reference lake,

58 which had no added AgNPs and was used to observe and compare natural spatial and temporal variations in benthic macroinvertebrate communities. Benthic macroinvertebrates were sampled regularly prior to and during the experiment from both lakes in order to examine whether any observed changes in benthic macroinvertebrate communities were a result of AgNP exposure or due to natural variation over time.

3.4 Methods

3.4.1 Study area

The whole lake experiment was conducted at the Institute for Sustainable

Development – Experimental Lakes Area (IISD – ELA) (49°40’N, 93°44’W), located on the Canadian Shield in Kenora District Ontario, Canada. The IISD – ELA is a field station composed of 58 lakes that have been used for ecosystem-scale experimentation and long-term monitoring of ecosystem processes since 1968 (Blanchfield et al., 2009).

The remoteness of the IISD – ELA allows for the study of environmental stressors at an ecosystem scale without the influences of point-source anthropogenic contamination. The two study lakes, Lake 221 (L221) and Lake 222 (L222), served as our reference and experimental lake, respectively. Both lakes are weakly stratified, oligotrophic and dimictic, and similar in size and depth, with L222 having a slightly larger surface area and maximum depth (16 ha, Zmax = 6 m) than L221 (9 ha, Zmax = 5.7 m).

3.4.2 Silver nanoparticles and whole lake dosing

Polyvinylpyrrolidone (PVP)-coated AgNPs were purchased as a powder from

NanoAmor (Houston, TX, USA). Information provided by the manufacturer indicated

59 that the AgNPs were spherical in shape with an average diameter of 30 – 50 nm. The

AgNPs were suspended in 0.8 µm-filtered lake water using a KADY® Mill Model L rotor-stator dispersion mill (Scarborough, ME, USA). Gum arabic (0.025%) was added as a stabilizer during the process of suspension to prevent agglomeration of the nanoparticles in the lake (Sperling & Parak, 2010). AgNP suspensions were made every other day and stored in dark PVC containers at 4°C prior to addition to the lake (Martin et al., 2017). Every second day during the ice-free season of 2014 and 2015, 12 L of the

AgNP suspension was added to the experimental lake (L222) using a metering pump from a point on the southwestern shore of the lake. The aim of this type of addition was to simulate point source exposure in surface waters impacted by wastewater discharges.

A total of 15 kg of AgNPs was added to L222 over the course of the study (2014 – 2015), with 9 kg added during the first year (14 June – 23 October 2014), and 6 kg added during the second year (15 May – 25 August 2015).

3.4.3 Collection and processing of lake water samples

Dissolved oxygen (DO), electrical conductivity and temperature were measured weekly in situ at centre buoy at 0.5 m intervals from the surface to the maximum depth of both lakes, using a hand-held multiparameter probe (YSI30-25FT, YSI Inc., Yellow

Springs Ohio, USA). Turbidity was measured using Secchi disk depth at centre buoy.

Epilimnetic temperatures were measured from centre buoy in L221 and L222 using

HOBO® temperature loggers (Onset Computer Corp.), which measured surface water temperatures every half hour each day in 2013 (2 June to 31 August), 2014 (31 May to 3

November), and 2015 (5 May to 10 November).

60 Water samples were collected throughout the ice-free season from centre buoy at two depths (epilimnion and hypolimnion) in 2013 and three depths (epilimnion, metalimnion and hypolimnion) in 2014 and 2015, which were determined based on stratification (i.e. vertical temperature profile) of the lakes at the time of sampling.

Sampling dates varied between years, with samples collected approximately every three weeks in 2013 (3 June to 20 October), every two weeks in 2014 (3 June to 20 October), and monthly in 2015 (11 May to 26 October). Approximately 8 L of lake water was collected from each depth per sampling date using a Van Dorn sampler and pre-screened through a 35 µm mesh filter to remove any larger plankton or debris. During the years of

AgNP addition (2014 – 2015), an additional 4 L were collected from each depth. Samples were analyzed for dissolved organic carbon (DOC) and total dissolved nitrogen (TDN), as well as particulate phosphorus (P), nitrogen (N) and carbon (C) in the seston fraction

(35 – 0.7 µm). For dissolved nutrients (DOC, TDN, TDP), 500 ml of pre-filtered (35 µm) water was run through a 0.7 µm Whatman GF/F filter, and subsequently through a 0.2 µm polycarbonate filter. The filtrate was collected and subsamples for DOC and TDN were stored in 125 ml glass amber bottles in the dark at 4ºC, whereas TDP subsamples were stored in 125 ml plastic bottles and frozen until further analysis. For Chl-a concentrations, 200 – 400 ml of pre-screened water was run through unashed Whatman

GF/F filters in duplicate, stored in aluminum foil and frozen. For seston C, N and P analysis, samples of pre-filtered (35 µm) water were run through ashed Whatman GF/F filters in duplicate, stored in polyethylene bags and dried at 60ºC. TDN, TDP and seston

P were measured by spectroscopy (APHA, 1992), DOC by persulfate digestion, and

61 seston C and N by dry combustion using an elemental CN analyzer (Vario EL III

Elementar) (Vincent et al., 2017).

3.4.4 Ag analysis

To determine the presence of total (TAg) and dissolved AgNPs (DAg) in the water column, water samples were collected from five locations along a gradient near and distant to the point source (point source, 0.5 m, 5 m, centre buoy, outflow) (Fig. 3.1; Fig.

3.2), selected to observe the distribution of AgNPs from point source. For TAg analysis,

10 ml of pre-screened (35 µm) water was acidified with concentrated (70%) nitric acid

(HNO3) to a total concentration of 4% (v/v) and stored at 4ºC in 15 ml Falcon centrifuge tubes until further analysis. Samples for DAg (<0.2 µm) analysis were filtered through a

1.2 µm polycarbonate filter, and subsequently through a 0.2 µm polycarbonate filter prior to acidification with concentrated (70%) HNO3 to a final volume of 4%, and then stored similarly to TAg samples for later analysis. Water samples for TAg and DAg were heated at 70°C for six hours prior to analysis by inductively coupled plasma mass spectrometry

(ICP-MS). A detailed description of analytical methods can be found in Rearick et al.

(2017). For the purposes of this study, only TAg concentrations collected from the epilimnion of centre buoy are represented.

3.4.5 Benthic macroinvertebrate sampling and taxonomic analyses

To assess the responses of natural benthic macroinvertebrate communities to

AgNPs, rock bags, which served as artificial substrates, were deployed along the littoral zone of our reference (L221) and experimental lake (L222) early in the ice-free season

(early- to mid-May). To prepare the rock bags, cobbles of approximately the same size

62 were collected from the surrounding area and stored in 1-cm mesh bags. The rock bags were then placed along the littoral zone of the lakes, at three predetermined locations in

2013 (August), 2014 (June – October), and 2015 (June – October). Sites were selected to be at approximately equal distances of each other (Fig. 3.1) and at similar depths (~ 0.75 m). The habitat into which the rock bags were placed was chosen to be as similar as possible across sites, and consisted of a rocky bottom with cobbles of different sizes that allowed for the colonization of macroinvertebrates from the surrounding area. As we only had a single sampling date in the first year of our study (14 August 2013), a total of 9 rock bags were placed in each lake, with three replicate rock bags allocated to each of the three locations. In 2014 and 2015, a total of 36 rock bags were placed in each lake at the start of each season, with twelve rock bags placed at each location. After allowing native macrofauna to naturally colonize, three replicate rock bags were removed from each site in August 2013, and then in approximately monthly intervals (June, July, August and

October) in 2014 and 2015. Samples were then brought back to the laboratory at the

ELA, where the contents were gently scrubbed and preserved in 95% ethanol for further enumeration and identification. All samples were sieved through an ASTM no. 60 (250

µm) sieve prior to identification. Preserved samples were processed using the Ontario

Benthos Biomonitoring Network (OBBN) protocol (Jones et al., 2004) to determine benthic macroinvertebrate abundance and community composition. A Leica S8 APO stereomicroscope (8:1 zoom with standard magnification of 10x – 80x) was used to identify the organisms, and sort them into 27 taxonomic groups as defined in Jones et al.

(2004). All samples were processed in their entirety.

63 3.4.6 Statistical analysis

Benthic community structure was analyzed for taxonomic richness, total abundance of individuals and relative abundance of sensitive (Ephemeroptera [mayflies],

Plecoptera [stoneflies] and Trichoptera [caddisflies] = % EPT) and tolerant (%

Chironomidae) taxa. Linear mixed-effects (LME) models (lmer function in the R package

‘lme4’) were used to assess temporal variation in benthic macroinvertebrate community metrics. Treatment (i.e. reference vs. experimental lake) and year (2013 – 2015) were set as main fixed effects with interactions, while monthly sampling date was set as a crossed random effect, with site included as a nested random factor (nested within treatment). If an interaction effect was found to be significant (P < 0.05), Tukey’s all-pair comparisons was applied using the R package ‘emmeans’.

Changes in macroinvertebrate community structure and effects of environmental variables on patterns of community variation across years within L221 and L222 were examined using multivariate methods. Multivariate ordination techniques provide a more in-depth analysis, and go beyond observations of changes in total abundance or taxonomic richness. To examine differences in benthic macroinvertebrate community structure between our reference (L221) and experimental (L222) lake across years, an analysis of similarity (ANOSIM) was performed on the Bray-Curtis dissimilarity matrix

(Bray & Curtis, 1957) using the ‘vegan’ package (Oksanen et al., 2016) for the R software environment (R Development Core Team, 2016). ANOSIM is a non-parametric multivariate technique that uses a dissimilarity matrix to compare variation between samples by calculating a test statistic R as an index of relative within-group dissimilarity, with minimal statistical assumptions (Clarke, 1993; Oksanen et al., 2016). Prior to

64 analysis, benthic macroinvertebrate abundances were transformed and standardized

(square-root transformation and Wisconsin double standardization) to stabilize variances, and rare taxa (<5% frequency) were removed from analyses. A similarity percentages

(SIMPER) analysis was subsequently conducted to examine the primary macroinvertebrate taxa that best explained any differences detected with the ANOSIM.

A Non-Metric Multidimensional Scaling (NMDS) ordination using the Bray-

Curtis dissimilarity matrix was subsequently performed to visualize patterns of change in benthic macroinvertebrate communities between lakes and years. Samples collected from both pre-AgNP addition (2013) and during the AgNP addition (2014 and 2015) years were run in a combined ordination to allow for an assessment of change in our study lakes over time. To further examine any changes in stability detected within the communities, coefficients of variation (CV) were calculated for community metrics

(Pearson, 1896) and then statistically tested using the ‘cvequality’ package in R

(Marwick & Krishnamoorthy, 2018). To analyze environmental factors that may have influenced changes in benthic macroinvertebrate assemblages, a BIO-ENV analysis

(Clarke & Warwick, 2001) was performed using the bioenv function in the ‘vegan’ package (Oksanen et al., 2016). Within the bioenv function, Spearman correlation coefficients are calculated between the macroinvertebrate community dissimilarity matrix and Euclidean distances of scaled environmental parameters, where the stronger the correlation between the community matrix and environmental variable, the greater the influence of that variable (or combination of variables) on variations in community composition (Oksanen et al., 2016). All water chemistry variables were log10 transformed to improve the normality of the data.

65 3.5 Results

3.5.1 Benthic macroinvertebrates

Twenty-one taxa were found in both L221 and L222 across years (Appendix 2).

Chironomidae was the most abundant taxon across all sampling dates in both L221 and

L222 (with the exception of June 2015 in L221, where Hirudinea dominated the community with 38%), accounting for 28% to 87% of the benthic invertebrate community in L222, and 30% to 78% in L221 (Fig. 3.2). Other predominant taxa (>5% occurrence frequency) in both lakes included Hirudinea, Ceratopogonidae,

Ephemeroptera, Trichoptera, Miscellaneous Diptera, Amphipoda, Hydrachnidae,

Bivalvia, Anisoptera, Coelenterata, and Decapoda, while all other taxa were found only occasionally (<5% occurrence frequency) (Appendix 2).

Total abundance of individuals did not differ between treatments (LME models,

F1,48 = 0.225, P = 0.637); however, a significant main effect of year was exhibited (LME models, F2,48 = 59.942, P < 0.001), with abundances declining in the final year of the study (Tukey’s HSD, P < 0.05). In our reference lake, mean abundances dropped from

139 and 99 in the first and second year of the study, respectively, to an average of 20 individuals recorded in 2015 (Fig. 3.3). Similarly, in our experimental lake, mean abundances declined by 71% and 81% in 2015 compared to 2013 and 2014 (Fig. 3.3).

These shifts were reflected in the declines of most taxa present, with average annual abundances of individuals in L221 decreasing by 86% and 83% (Chironomidae), 94% and 88% (Ceratopogonidae), 85% and 70% (Hirudinea), 82% and 90% (Miscellaneous

Diptera), 81% and 32% (Trichoptera), 94% and 18% (Ephemeroptera), and 23% and 22%

(Amphipoda) in 2015 compared to 2013 and 2014, respectively. Similar trends were

66 observed in our experimental lake, where average abundances in 2015 were lower by

74% and 86% (Chironomidae), 95% and 92% (Ceratopogonidae), 62% and 35%

(Hirudinea), 62% and 86% (Miscellaneous Diptera), 35% and 46% (Trichoptera), 39% and 17% (Ephemeroptera), and 34% and 62% (Amphipoda) than in 2013 and 2014, respectively. Bivalvia and Hydrachnidae were among the few taxa that experienced increases in average annual abundances in both L221 and L222 by the end of the study.

Taxa richness was found to differ across years as well (LME models, F2,45 = 19.380, P <

0.001), with richness significantly lower in the final year of the study than in both prior years (Tukey’s HSD, P < 0.05; Fig. 3.4). No differences in taxa richness were observed between treatments (LME models, F1,45 = 1.540, P = 0.221).

Relative abundances of Ephemeroptera-Plecoptera-Trichoptera (EPT) and

Chironomidae did not differ significantly between treatments; however, a strong treatment main effect of year was observed for both EPT and Chironomidae (LME models, F2,42= 18.035, P < 0.001 and F2,45 = 23.257, P < 0.001), with relative abundances of Chironomidae exhibiting a significant decline in 2015 (Tukey’s HSD, P < 0.05; Fig.

3.2; Fig. 3.5), and EPT declining significantly in 2014 compared to other years (Tukey’s

HSD, P < 0.05; Fig. 3.2). Relative abundances of all other benthic macroinvertebrate taxa also varied significantly over time (LME models, F2,45 = 19.393, P < 0.001), increasing in

2015 across both lakes (Tukey’s HSD, P < 0.05), likely due to the decline in relative abundances of Chironomidae (Fig. 3.2; Fig. 3.5). Tests for the equality of coefficients of variation from multiple groups reported similar values between our reference and experimental lake for all community metrics, with no differences in variation detected between lakes (Table 3.1).

67 3.5.2 Multivariate analyses

To assess community change resulting from AgNP exposure, non-parametric multivariate analyses were performed. The dataset composed of 12 benthic taxa was used in the multivariate analyses. Ordination of the taxonomic data (double standardized) by

NMDS produced a 2-dimensional plot with a final stress value of 0.194, which is reasonable considering the dimensionality of the data. To help discern differences among clusters, 95% confidence ellipses were added to the ordination plot. The resulting ordination map showed high similarity between sites in L221 and L222 regarding taxonomic composition, seen as overlapping points on the map (Fig. 3.5). An increase in variability was observed across both lakes from 2013 to 2015 (Fig. 3.5), suggesting substantial alteration in the invertebrate communities in L221 and L222 over time. The

ANOSIM confirmed significant differences between sampling years across the full dataset (R = 0.603, P = 0.001), with no differences observed between our experimental and reference lake across years (R = 0.011, P = 0.240). Results of the similarity percentage (SIMPER) analysis identified three taxa in L221 (Chironomidae,

Ephemeroptera, Miscellaneous Diptera) and L222 (Chironomidae, Miscellaneous

Diptera, Ceratopogonidae) as key taxa responsible for the dissimilarity observed between years within each lake (Table 3.2).

To detect relationships between taxonomic structure and selected environmental variables for L221 and L222, Spearman correlation coefficients were calculated between the benthic macroinvertebrate community dissimilarity matrix and environmental parameters using the BIO-ENV procedure (Table 3.3). The subset of environmental variables that was best correlated with macroinvertebrate community composition (ρ =

68 0.431) was a combination of TDP, seston P, pH and temperature (Table 3.3; Fig. 3.7).

When examining maximum rank correlations within each lake individually, a strong correlation between benthic macroinvertebrate assemblages and TDP (ρ = 0.603) was found in Lake 221, whereas in Lake 222, a combination of TDP, DOC, seston N, seston P and pH best explained the patterns in community composition (ρ = 0.568). In contrast to our findings with littoral microcrustacean communities (Chapter 4), TAg was not strongly correlated with benthic macroinvertebrate community structure (Table 3.3; Fig.

3.7).

3.6 Discussion

We released AgNP to an entire lake at the Experimental Lakes Area to understand the effects of AgNPs on benthic macroinvertebrate communities. Our results indicate that benthic macroinvertebrate community metrics were not impacted by a chronic addition of low doses of AgNPs in a natural setting. Although we observed a significant decline in total benthic macroinvertebrate abundance and taxonomic richness over time in our experimental lake, these findings were also seen in our reference lake, indicating that benthic community dynamics were likely influenced by water quality parameters and seasonal trends in both lakes rather than AgNP exposure. Furthermore, relative abundances of metal-sensitive taxa (Ephemeroptera, Plecoptera and Trichoptera)

(Beltman et al., 1999) and metal-tolerant taxa (Chironomidae, Ceratopogonidae and

Miscellaneous Diptera) (Zimmerman et al., 1993; Barbour et al., 1999; Beltman et al.,

1999; Loayza-Muro et al., 2010) were seen to exhibit similar patterns of change in both lakes. These results contrast sharply with those reporting strong adverse effects of AgNPs

69 on lower trophic levels (Asharani et al., 2008; Griffitt et al., 2008; Khaydarov et al.,

2009; Croteau et al., 2011; Das et al., 2014), and on benthic macroinvertebrates in particular (Croteau et al., 2011; Buffet et al., 2014; Gambardella et al., 2015; Brittle et al.,

2016). However, as the majority of these studies have been conducted under laboratory conditions, in artificial media, and over short periods of time, these findings may not be surprising. Blinova et al. (2013) showed that AgNPs had a stronger toxic effect on aquatic crustaceans in artificial freshwater than in natural water. Similarly, Park et al.

(2015) reported that the negative effects of AgNPs on the growth of the benthic macroinvertebrate Chironomus riparius (Chironomidae) were mitigated in the presence of organic surface coatings and natural sediment.

Multivariate analyses revealed a high similarity in taxonomic composition across sites in L221 and L222, likely as a result of geographic proximity and similar benthic habitats, as well as similar chemical and physical characteristics of both lakes (Fig. 3.1,

Appendix 3). Relationships between macroinvertebrate communities, mean TAg concentrations, and mean concentrations of environmental variables were identified using

Spearman correlation coefficients between the community dissimilarity matrix and

Euclidean distances of scaled environmental factors. Despite our predictions, TAg was not a strong driver of benthic community structure; instead, TDP, seston P, temperature and pH were identified as variables with the strongest correlation with benthic community composition. From this it is clear that natural temporal changes in physico- chemical parameters were mainly responsible for benthic macroinvertebrate community dynamics (Hämäläinen et al., 2003; Mesa, 2012, and not exposure to AgNPs.

70 Our findings suggest that natural conditions may have mitigated the bioavailability and toxicity of AgNPs to benthic macroinvertebrates, resulting in a lack of

AgNP toxicity observed in our whole lake study. The stabilization of AgNPs as a result of binding to DOC, for example, has been shown to reduce the toxicity of AgNPs to lower-level aquatic organisms. In their study, Gao et al. (2009) showed that the overall toxicity of AgNPs to the microcrustacean Ceriodaphnia dubia was reduced in natural waters compared to artificial media, further noting that toxicity decreased with a decrease in ionic strength and an increase in DOC. Other studies have also reported on the mitigating effects of DOC on the biotoxicity of AgNPs. Blinova et al. (2013) reported reduced toxicity of AgNPs to the fairy shrimp (Thamnocephalus platyurus) in the presence of DOC. Similarly, the presence of natural organic matter reduced the strong toxic effect of AgNPs to the nematode Caenorhabditis elegans (Yang et al., 2014). The relatively high concentrations of DOC (8.45 – 15.98 mg L-1) and low ionic strength measured within our experimental lake could have therefore been responsible for the observed lack of toxicity to benthic macroinvertebrate communities (Gao et al., 2009;

Conine, 2017).

Our inability to observe any strong effects of an addition of AgNPs could also be explained by the lower level of taxonomic resolution used in this study. Although various studies have shown that the use of coarser taxonomic levels (i.e. family-level and higher) provides sufficient information for assessing changes in benthic macroinvertebrate communities as a result of anthropogenic impacts (Olsgard et al., 1997; Bailey, 2001;

Arscott et al., 2006), subtler effects may be detectable only at lower levels of identification (i.e. genus- or species-level) (Arscott et al., 2006). Resh (1979) and Resh

71 and Unzicker (1975) have similarly suggested that the use of species-level discrimination allows for a greater degree of distinction among communities, as species within the same genus may exhibit varying sensitivities (and therefore responses) to pollution that may not be seen at higher levels of taxonomic discrimination. The effects of low, environmentally relevant concentrations of AgNPs added to our experimental lake may have been too weak to detect a response at the coarse taxonomic levels used. Any responses in benthic macroinvertebrate larvae, if present, may only be seen at finer taxonomic scales (e.g. genus or species) (Bailey, 2001; Arscott et al., 2006).

As an increase in TDP was recorded across both lakes in 2015 (by 61% and 68% compared to 2014 in our experimental and reference lake, respectively; Appendix 3), this could explain the observed changes in benthic macroinvertebrate assemblages. Nutrient enrichment in freshwater systems has been associated with alterations in benthic macroinvertebrate community structure (Widbom & Elmgren, 1988; Steinman & Ogdahl,

2011) and decreases in taxa richness (Ortiz & Puig, 2007). Furthermore, Steinman &

Ogdahl (2011) reported an increase in macroinvertebrate density following a decrease in

TDP after alum treatment. These relationships are fairly complex, however, and effects of

TDP have been found to vary depending on benthic habitat (Brauns et al., 2007; O’Toole et al., 2008), presence of macrophytes (Langdon et al., 2010) and alkalinity (O’Toole et al., 2008). In contrast, other authors have reported increases in macroinvertebrate diversity (Rader & Richardson, 1992), densities (Rader & Richardson, 1992;

Blumenshine et al., 1997) and biomass (Rosemond et al., 2001) with increasing nutrient concentrations in freshwater environments, while others reported no observable effect of abiotic factors on benthic macroinvertebrate communities (Leppä et al., 2003). It is clear

72 that the responses of benthic macroinvertebrate communities to water chemistry parameters are not always predictable, and it is likely that the responses of benthic macroinvertebrates to TDP, if present, were indirect.

There may still be some unmeasured environmental variables that could explain the substantial decline in abundance and increase in variance within the benthic community. As a similar trend was also seen within the littoral microcrustacean communities of our experimental and reference lake, with overall densities of microcrustaceans declining rapidly in the final year of the study (Chapter 4), it is possible that a common factor affecting solely the littoral zone of the lakes may have been responsible for the steep decline in the abundances of both littoral macroinvertebrates and microcrustaceans. For example, physical alterations in littoral habitats (Heino, 2000) or increases in fish predation (Leppä et al., 2003; Miracle et al., 2006) heavily impact littoral community structure. Due to their shallowness, these habitats are easily impacted by weather patterns, such as water-level fluctuations or increases in wave action (Pabst et al.,

2008; Levin et al., 2012), processes that cause decreases in the number of benthic macroinvertebrates (Pabst et al., 2008) and algae (Hoagland & Peterson, 1990;

Kreuzinger-Janik et al., 2015). Further, Diehl (1992) found that predatory perch (Perca fluviatilis) caused large shifts in both the biomass and abundance of macroinvertebrates and microcrustaceans. Although the rise in temporal variance was not a result of AgNP exposure, as variability appeared to be equal between our reference and experimental lake, these changes should not be dismissed, as increases in temporal variability (i.e. decreases in stability) such as these may be an early warning sign of an imminent large-

73 scale shift within the community (Brock & Carpenter, 2006; Biggs et al., 2009; Eason et al., 2013).

In conclusion, our findings indicate that benthic macroinvertebrate communities are not sensitive to AgNP exposure under environmental conditions, but are rather influenced by seasonal dynamics and nutrient concentrations (Allan, 1995; Becerra-

Jurado et al., 2009). The cause for the marked increase in variability in community composition observed from 2013 to 2015 in both our experimental and reference lake, along with the substantial decline in macroinvertebrate abundances across both lakes in

2015 remains unknown and requires further investigation, as such changes in benthic community structure could have larger implications for the structure and function of freshwater benthic systems, releasing predatory pressure on lower trophic levels, and decreasing the availability of benthic macroinvertebrates as a food source for fish.

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84 3.8 Tables and Figures

Table 3.1 Summary of tests on the equality of coefficients of variation (CV) in benthic macroinvertebrate community metrics by treatment. Included are Feltz and Miller’s asymptotic test and Krishnamoorthy and Lee’s modified signed-likelihood ratio (M- SLRT) test.

Variable Test name Test statistic P

Asymptotic 0.018 0.894 Abundance M-SLRT 0.016 0.901

Shannon -Wiener Asymptotic 0.343 0.558 diversity M-SLRT 0.334 0.563

Asymptotic 0.746 0.388 Taxonomic richness M-SLRT 0.703 0.402

Asymptotic 0.214 0.644 Pielou’s evenness M-SLRT 0.199 0.656

85 Table 3.2 Similarity percentages (SIMPER) analysis indicating which benthic macroinvertebrate taxa were primarily responsible for the observed dissimilarity (as indicated by ANOSIM results; R = 0.603, P = 0.001) in community composition over time (i.e. between years) in both our reference (Lake 221) and experimental lake (Lake 222). Benthic macroinvertebrate taxa are ranked according to their average contribution (Contrib. (%)) to community dissimilarity. Average abundances of taxa within each year (Av. abund.), and percentage of cumulative similarity (Cum. (%)) are also included. A cut-off value of 80% cumulative similarity was applied.

Av. abund. Av. abund. Contrib. Cum. 2013 vs. 2014 (2013) (2014) (%) (%)

Chironomidae 91.44 69.61 10.3 36.9 Ephemeroptera 17.44 1.61 6.8 61.3 Miscellaneous Diptera 4.56 12.14 4.2 76.5

2013 vs. 2015 Av. abund. Av. abund. Contrib. Cum. (2013) (2015) (%) (%) Lake 221 Lake Chironomidae 69.61 11.67 49.7 71.4 Ephemeroptera 17.44 1.28 10.3 79.9

2014 vs. 2015 Av. abund. Av. abund. Contrib. Cum. (2014) (2015) (%) (%) Chironomidae 69.61 11.67 49.7 71.4

2013 vs. 2014 Av. abund. Av. abund. Contrib. Cum. (2013) (2014) (%) (%) Chironomidae 54.44 97.64 19.7 64.8

2013 vs. 2015 Av. abund. Av. abund. Contrib. Cum.

(2013) (2015) (%) (%) Chironomidae 54.44 13.17 36.4 62.6 Ceratopogonidae 6.56 0.25 5.5 72.1 Lake 222 Lake Miscellaneous Diptera 7.67 2.47 4.6 79.9

2014 vs. 2015 Av. abund. Av. abund. Contrib. Cum. (2014) (2015) (%) (%) Chironomidae 97.64 13.17 49.2 72.6

86 Table 3.3 Results of the BIO-ENV analysis indicating which environmental variables were most strongly correlated with benthic macroinvertebrate assemblages. Variables in bold indicate the best match.

Best variable combination Rho (ρ)

TDP 0.402

TDP, seston P 0.411

TDP, temperature, seston P 0.418

TDP, temperature, seston P, pH 0.431

TDP, temperature, seston P, pH, TAg 0.402

TDP, DOC, temperature, seston P, pH, TAg 0.369

TDP, DOC, temperature, seston N, seston P, pH, TAg 0.313

TDP, DOC, temperature, seston N, seston C, seston P, pH, TAg 0.266

TDP, TDN, DOC, temperature, seston N, seston C, seston P, pH, TAg 0.163

87 Figure Captions

Figure 3.1 Sampling locations along the littoral and pelagic zone of the (A) reference

(Lake 221) and (B) experimental lake (Lake 222) at the IISD – Experimental Lakes Area.

Samples were collected prior to (2013) and during AgNP addition (2014 – 2015). Centre buoy indicates the deepest part of the lake and is marked with ×. Point source indicates the location on the southwestern shore of Lake 222 from which AgNPs were added to the lake.

Figure 3.2 Relative abundances (%) of Chironomidae, Ephemeroptera-Plecoptera-

Trichoptera (EPT) and other benthic macroinvertebrate taxa across years in our reference lake (Lake 221) and experimental lake (Lake 222). AgNP additions to Lake 222 occurred throughout 2014 and 2015. Values are monthly averages across all sites (n = 3).

Figure 3.3 Average monthly abundance (number of individuals) of benthic macroinvertebrates in our reference (Lake 221; L221) and experimental lake (Lake 222;

L222) at the IISD – Experimental Lakes Area over a three-year study period. AgNP additions to Lake 222 began in 2014 and continued throughout 2015. Values are means of locations (n = 3) ± standard error.

Figure 3.4 Average monthly taxa richness (number of taxa) of benthic macroinvertebrates in our reference (Lake 221; L221) and experimental lake (Lake 222;

Figure 3.5 Average monthly abundance (number of individuals) of Chironomidae in rock bag samples in our reference (Lake 221; L221) and experimental lake (Lake 222; L222)

88 at the IISD – Experimental Lakes Area over a three year study period. AgNP additions to

Lake 222 occurred throughout 2014 and 2015. Values are means of locations (n = 3) ± standard error.

Figure 3.6 Two-dimensional non-metric multidimensional scaling (NMDS) ordination based on the Bray-Curtis dissimilarity matrix calculated from benthic macroinvertebrate abundance data. The ordination illustrates differences in macroinvertebrate community composition in our experimental lake (Lake 222) and reference lake (Lake 221) over time prior to (2013) and during a two-year (2014 – 2015) experimental addition of AgNPs to

Lake 222 at the IISD – ELA. Data points are sampling locations from 2013 to 2015; points that are closer together in ordination space are more similar in terms of macroinvertebrate community composition than those further apart. White symbols represent Lake 221; black symbols represent Lake 222. Ellipses represent ordination confidence intervals (95%).

Figure 3.7 Non-metric multidimensional scaling (NMDS) ordination of benthic macroinvertebrate communities in relation to environmental variables that were most strongly correlated with community structure. Samples were collected across an experimental (Lake 222) and reference lake (Lake 221) over the course of three years, prior to (2013) and during a two-year (2014 – 2015) experimental addition of AgNPs to

Lake 222 at the IISD – ELA.

89 A

Sampling sites: × Littoral microcrustaceans Benthic macroinvertebrates × Pelagic zooplankton Point source

Figure 3.1

90 Lake 221 100

20 macroinvertebrates [%]

Relative abundance of benthic 0 July July June June August August August October October

2013 2014 2015

Chironomidae EPT Others

Lake 222 100

20 macroinvertebrates [%] Relative abundance of benthic 0 July July June June August August August October October

2013 2014 2015

Chironomidae EPT Others

Figure 3.2

91 250 L221 L222 200

150

100 abundance

Average benthic macroinvertebrate 0 Aug Jun Jul Aug Oct Jun Jul Aug Oct

2013 2014 2015

Figure 3.3

10 9 L221 L222 8 7 6 5 4 taxa richness 3 2 1

Average benthic macroinvertebrate 0 Aug Jun Jul Aug Oct Jun Jul Aug Oct

2013 2014 2015

Figure 3.4

92 180 L221 L222 160 140 120 100 80 60 40 20

Average Chironomidae abundance 0 Aug Jun Jul Aug Oct Jun Jul Aug Oct 2013 2014 2015

Figure 3.5

93 Non-metric fit R2 = 0.962 Linear fit R2 = 0.844 Stress = 0.194 0.5 NMDS2 0.0

2013

− 0.5 Lake 221 2014 Lake 222 2015

−1.0 −0.5 0.0 0.5 1.0

NMDS1

Figure 3.6

Non-metric fit R2 = 0.989 TDP 2

0.4 Linear fit R = 0.938 Stress = 0.107 Bivalvia

P 0.2 Trichoptera pH Decapoda Ephemeroptera Hirudinea 0.0

Chironomidae Hydrachnidae NMDS2 − 0.2 Ceratopogonidae Temperature − 0.4 − 0.6

Coelenterata − 0.8 −0.5 0.0 0.5 1.0

NMDS1

Figure 3.7

94 CHAPTER 4

4.1 Preface

Title: Responses of lower trophic levels to a whole lake addition of silver nanoparticles

Authors: Katarina A. Cetinic1, Michael Paterson2, Beth C. Norman3,4, Daniel C. Rearick1,

Paul Frost3, Marguerite A. Xenopoulos3

Contact Information:

1Environmental and Life Sciences Graduate Program, Trent University, Peterborough,

Ontario, Canada

2International Institute for Sustainable Development – Experimental Lakes Area,

Winnipeg, Manitoba, Canada

3Department of Biology, Trent University, Peterborough, Ontario, Canada

4Present address: Department of Microbiology and Molecular Genetics, Michigan State

University

Corresponding author: Katarina A. Cetinic, [email protected], 705-748-1011 ext.

6467

Status: In preparation for submission

Author Contributions: KAC, MP, BC, and MAX conceived and designed the experiments. KAC conducted the experiments and processed the samples. KAC, MP and

MAX analyzed the data. KAC wrote the manuscript. MAX supervised the project. All authors provided critical feedback and helped with paper revisions.

Keywords: Silver nanoparticles, zooplankton, littoral microcrustaceans, benthic macroinvertebrates, Experimental Lakes Area, community composition, lake ecosystem

95 CHAPTER 4

Responses of natural littoral microcrustacean communities to a whole lake addition

of silver nanoparticles

4.2 Abstract

The rapid increase in the use of silver nanoparticles (AgNPs) has been causing growing concern regarding their environmental impacts. Although many studies have reported adverse effects of AgNPs on aquatic organisms, the majority of these studies have been conducted under laboratory conditions, making predictions of their effects in natural environments difficult. Freshwater benthic communities could be especially susceptible to the toxicity of AgNPs due to the potential of these nanoparticles to accumulate in lake sediments and periphyton. For the first time, the effects of AgNPs on benthic microcrustacean communities in littoral habitats have been investigated in situ.

To examine the response of these communities to AgNPs, we released environmentally relevant (~10 ug/L) concentrations of AgNPs into a lake (Lake 222; L222) at the

International Institute for Sustainable Development – Experimental Lakes Area (Ontario,

Canada) over a two-year period, and compared effects to natural variation observed within our reference lake (Lake 221; L221). A distance-based analysis of similarity

(ANOSIM) and non-metric multidimensional scaling (NMDS) were used to examine differences in microcrustacean assemblages between lakes over time to examine whether an addition of AgNPs caused significant alterations in microcrustacean communities.

Overall, AgNPs did not have a strong effect on littoral microcrustacean communities in terms of community composition, density, diversity or species richness. Instead,

96 fluctuations were observed across both lakes, indicating that biodiversity metrics were more strongly impacted by natural changes in environmental variables. A marked decline in the relative abundance of the benthic family Chydoridae was observed solely in L222, suggesting that this taxon may have been most sensitive to AgNP exposure. More long- term in situ studies on the impacts of AgNP exposure on microcrustacean communities are necessary, as declines in highly abundant taxa such as Chydoridae can have a strong effect on the structure and function of littoral ecosystems.

Keywords: Silver nanoparticles, littoral microcrustaceans, Chydoridae, Experimental Lakes Area, community composition, lake ecosystem

4.3 Introduction

The widespread use of nanoparticles (1 – 100 nm in diameter) (Katz et al., 2004) is causing growing concern regarding their potential impacts on natural ecosystems.

Silver nanoparticles (AgNPs) are the most prevalent nanoparticle in commercial use today due to their potent antibacterial and antifungal properties (Fabrega et al., 2011;

Lara et al., 2011). With the increased use of AgNPs likely to result in elevated concentrations in the environment (Maes et al., 2009), it is crucial that we investigate the effects these nanoparticles could have on aquatic systems.

Although much research has demonstrated the toxicity of AgNPs to major taxonomic categories of aquatic organisms (Asharani et al., 2008; Navarro et al., 2008;

Khaydarov et al., 2009; Croteau et al., 2011; Buffet et al., 2014; Das et al. 2014;

Gambardella et al., 2015; Brittle et al., 2016), the majority of these studies involve small- scale (e.g. microcosm and laboratory) experiments that are limited in their ability to

97 replicate the influence of environmental factors, making extrapolations of these findings to natural systems problematic. Ambient environmental factors have repeatedly been shown to have a strong influence on the properties of AgNPs, altering the behaviour, and therefore toxicity, of these nanoparticles (Aiken et al., 2011; Stebounova et al., 2010; Liu et al., 2012; Levard et al., 2012; Poda et al., 2013), and highlighting the importance of studying the effects of AgNPs in a natural setting.

Chronic metal contamination has often been shown to have strong adverse effects on pelagic microcrustaceans (i.e. zooplankton), causing decreases in total abundance, species richness and diversity, as well as shifts in community structure due to overall higher tolerances of Copepoda to metal-induced stress compared to Cladocera (Kerrison et al., 1988; Monteiro et al., 1995; Gagneten & Paggi, 2009). However, comparatively fewer studies researching the effects of AgNPs on littoral microcrustacean species have been conducted to date, both in the laboratory (Wang et al., 2015) and in situ (Vincent et al., 2017), than those on pelagic zooplankton (Zhao & Wang, 2011; Asghari et al., 2012;

Blinova et al., 2013). As littoral microcrustaceans often exhibit higher diversity and densities than those in the pelagic zone (Julli, 1986; Castilho-Noll et al., 2010), it is imperative we include these habitats when examining the effects of AgNPs on aquatic systems.

Periphyton, a complex assemblage of algae, bacteria, fungi, protozoa, meiofauna and detritus attached to submerged surfaces (Wetzel, 2001), plays an important functional role in freshwater systems. In addition to providing a food source and habitat for many benthic organisms (Likens, 2010), periphyton is also the dominant primary producer in the littoral zones of lakes, accounting for up to 98% of annual primary productivity

98 (Vadeboncoeur et al., 2003). Furthermore, littoral microcrustaceans that live within periphyton communities play a vital role in the transfer of energy within the benthic food web, feeding on bacteria and primary producers, and serving as an important food source for higher order consumers (e.g. benthic macroinvertebrates and larval fish) (Suthers &

Rissik, 2009). As a result, any changes in periphyton or littoral microcrustacean community composition in response to pollution or other stressors will likely have immediate and strong impacts on the structure and function of the whole freshwater ecosystem.

Many studies have demonstrated the capacity of periphyton to accumulate metals from the surrounding aquatic environment (Farag et al., 1998; Behra et al., 2002; Meylan et al., 2003; Serra et al., 2008), with AgNPs exhibiting a dose-dependent accumulation in periphyton (Furtado et al., 2014). This capacity of metals, and more specifically AgNPs, to accumulate in high levels within periphyton could pose a threat to benthic fauna inhabiting these communities (Farag et al., 1998; Haegerbaeumer et al., 2017), as their relatively sedentary lifestyle and feeding behaviour (i.e. scrapers that feed on biofilm)

(Thorp & Covich, 1991; Farag et al., 1998) may cause them to be continuously exposed to elevated concentrations of AgNPs.

This study evaluated the effects of low levels of AgNPs on natural periphytic littoral microcrustacean communities in a boreal lake. Environmentally relevant concentrations of AgNPs were added to an experimental lake at the IISD – Experimental

Lakes Area during the ice-free season of 2014 and 2015. We hypothesized that chronic

AgNP exposure to a freshwater lake would (i) reduce total microcrustacean abundance, species richness and diversity (Monteiro et al., 1995; Gagneten & Paggi, 2009), (ii) shift

99 relative abundances of more sensitive Cladocera in favour of more tolerant Copepoda

(Kerrison et al., 1988), and (iii) shift microcrustacean community structure towards genera tolerant to various metals (e.g. copper, nickel, cadmium, zinc), e.g.

Acanthocyclops spp., Chydorus spp. (Yan & Strus, 1979; Kerrison et al., 1988; Monteiro et al., 1995; Bossuyt & Janssen, 2005; Palmer et al., 2013). To address this hypothesis, we monitored five sites along the littoral zone of an experimental and reference lake before and during AgNP exposure to determine littoral microcrustacean community structure and dynamics across both lakes over time.

4.4 Methods 4.4.1 Study area

The whole lake study was carried out in the littoral zone of two lakes at the

International Institute for Sustainable Development – Experimental Lakes Area (IISD –

ELA) in Northwestern Ontario, Canada (49°40’N, 93°44’W). Lake 222 (L222) was selected as our experimental lake, whereas Lake 221 (L221) was used as our reference lake. L222 is slightly larger in area and deeper, with an area of 16 ha and a maximum depth of 6 m, while L221 has an area of 9 ha and a maximum depth of 5.7 m (Fig. 3.1).

Both lakes are oligotrophic and dimictic, and stratify in the summer months, forming a thermocline of around 2 m depth.

4.4.2 AgNP characterization and whole lake addition

The polyvinylpyrrolidone (PVP)-coated AgNPs (30 – 50 nm) were purchased in powder form by NanoAmor (Houston, TX, USA). AgNP suspensions were made every other day and stored in dark PVC containers at 4°C prior to addition to the lake. A detailed description of the methods used to suspend AgNPs, as well as the whole lake addition can

100 be found in Chapter 3 (3.3.2. Silver nanoparticles and whole lake dosing). AgNPs were added to our experimental lake (L222) from the southwestern shore during the ice-free season of 2014 and 2015. The amounts of AgNP added to the experimental lake were determined based on the volume of the lake, and aimed to reach approximately 10 µg L-1.

By targeting AgNP concentrations in the low µg L-1 range, these additions simulated realistic exposures in surface waters.

4.4.3 Collection and processing of lake water and AgNP samples

Physico-chemical parameters of water quality (dissolved oxygen, temperature, electrical conductivity and turbidity) were measured in situ at centre buoy on a weekly basis across both lakes using a hand-held multiparameter probe (YSI30-25FT, YSI Inc.,

Yellow Springs Ohio, USA). For nutrient and AgNP analysis, water samples were collected from centre buoy using a Van Dorn water sampler, and screened immediately through a 35 µm mesh filter to remove any macroplankton or debris. Collected water samples were analyzed for dissolved organic carbon (DOC), total dissolved nitrogen

(TDN) and total dissolved phosphorus (TDP), as well as particulate (35 – 0.7 µm) phosphorus (P), nitrogen (N) and carbon (C), and total silver (TAg) concentrations. A detailed protocol for the collection and laboratory analysis of lake water samples is found in Chapter 3 (3.3.3. Collection and processing of lake water samples).

4.4.4 Littoral microcrustacean sampling and taxonomic analyses

To assess the responses of natural benthic littoral microcrustacean communities to

AgNPs, square unglazed ceramic tiles (48.5 mm x 48.5 mm), which served as artificial substrates for the colonization of periphyton and microcrustaceans (Peterson and

Stevenson, 1990; Barbour et al., 1999), were deployed along the littoral zone of our

101 reference (L221) and experimental lake (L222) between 2013 and 2015 (Fig. 3.1).

Ceramic tiles were chosen in order to decrease the variability of littoral habitats by providing a standardized environment for all replicates in regard to water quality parameters (e.g. light, dissolved oxygen, temperature). Each year we glued 4 tiles to a wooden board and placed the boards in five predetermined locations early in the ice-free season (mid to late May). The tiles were left undisturbed for three to five months (mid-

May to mid-August and mid-October) to allow for the natural colonization of littoral benthic microcrustaceans (Peterson and Stevenson, 1990). On each sampling date, two replicate tiles were randomly removed from each location, with care taken not to disturb the surrounding tiles, and were placed in Whirl-Pak® bags. In the laboratory, the microcrustaceans were collected by gently scraping all contents off the surface of the tiles

(Barbour et al., 1999), and then preserved in glass jars with 4% sugar formalin for identification and enumeration. Where possible, individuals were identified to species, otherwise to genus. Identifications were made using Smith and Fernando (1978) for

Copepoda, and Thorp and Covich (2009) for Cladocera. A Motic BA 400 compound light microscope at 10x magnification was used to identify the organisms. All samples were enumerated and identified in their entirety.

4.4.5 Periphyton analysis

Periphyton was collected monthly from June to October during the years of AgNP addition (2014 – 2015). A separate set of 12 tiles was added to the same wooden board

(Fig. 3.1), and from there three replicate tiles were randomly selected on each sampling date from each site and carefully collected in Whirl-Pak® bags. In the laboratory, contents from the entire surface of each tile were scraped off and homogenized in a

102 blender prior to filtration. Subsamples of the homogenized mixture were filtered onto pre-weighed ashed GF/F filers (Whatman) in duplicate, placed in polyethylene bags, and dried for 24 hours at 60°C. The samples were subsequently ashed (oxidized) in a muffle furnace, and the ash residue (inorganic content) was re-weighed. Organic matter content

(ash-free dry mass) on the tiles was determined by the difference in weight pre- and post- ashing.

4.4.6 Statistical analysis

Littoral microcrustacean community structure was analyzed for taxonomic composition, Cladocera and Copepoda densities, and biodiversity metrics including species richness (S [Whittaker, 1972]), diversity (H’ [Shannon and Weaver, 1949) and evenness (J [Pielou, 1967]). Temporal variation in microcrustacean densities and diversity metrics, as well as changes in periphyton biomass and Ag concentrations were assessed using linear mixed-effects (LME) models. These models are useful when analyzing nested data that includes fixed and random effects, in which observations from the same location are highly correlated due to measurements being repeated over time

(Bates et al., 2015). Treatment (i.e. lake) and year were set as fixed effects with interactive terms, while the random effects structure was month set as a crossed random effect, and sampling location nested within lake. The LME models were calculated using the lmer function within the lme4 package (Bates, 2010). To analyze for shifts in the stability within the community as a result of AgNP exposure, we calculated coefficients of variation (CV [Pearson, 1896]) for biodiversity metrics and microcrustacean density.

Feltz and Miller’s (1996) asymptotic test and Krishnamoorthy and Lee’s (2014) modified signed-likelihood ratio test were subsequently used to test for significant differences

103 between multiple CVs using the ‘cvequality’ package in R (Marwick & Krishnamoorthy,

2018).

To determine the relationship between AgNPs and periphyton development and microcrustacean density, we performed a repeated measures correlation for periphyton biomass and densities of Cladocera and Copepoda over time. A repeated measures correlation was conducted as samples taken from the same location over time were not independent of each other (i.e. repeated measures design). All repeated measures correlations were performed using the ‘rmcorr’ package in R (Bakdash & Marusich,

2017).

Differences in community composition over time, as well as ambient environmental factors influencing these changes were examined using multivariate methods. To visualize differences in microcrustacean community structure across years

(2013 – 2015) within each lake, a non-metric multidimensional scaling (NMDS) ordination was performed, with community dissimilarity assessed through relative distances of points on the NMDS plot. The NMDS ordination used the Bray-Curtis dissimilarity index on species abundance data that were8 transformed and standardized

(square-root transformation and Wisconsin double standardization) prior to analysis to equalize variances. To determine if observed differences in microcrustacean species composition between treatments (lakes) and sampling dates were significant, an analysis of similarity (ANOSIM) (Clarke, 1993; Oksanen et al., 2016) was performed on the

Bray-Curtis dissimilarity matrix.

To examine which environmental variables (periphyton Ag, TAg, DOC, TDN,

TDP, seston C, seston N, seston P, Chl-a, pH, temperature) were most related to littoral

104 microcrustacean species composition, a BIO-ENV analysis (Clarke & Ainsworth, 1993) was performed using the ‘vegan’ package in R (Oksanen et al., 2016). The BIO-ENV procedure calculates Spearman correlation coefficients using the microcrustacean community dissimilarity matrix and scaled environmental parameters to determine which subset of environmental variables correlates best with the community structure. For visualization of these results, environmental vectors were fit onto an NMDS ordination plot, with the direction of the arrow indicating correlation with the respective axis, and the length of the arrow indicating the strength of the correlation. All univariate and multivariate analyses were performed using the R software environment (R Development

Core Team, 2017).

4.5 Results

4.5.1 Water quality parameters

Concentrations of TDP exhibited significant temporal variation in L221, with higher values measured in 2015, while average TDN concentrations varied over time in both lakes (Appendix 4). pH values were significantly higher in our experimental lake

(L222) than in our reference lake (L221) (Appendix 4), with an average measured value of 7.86 in L222 and 6.50 in L221. Differences in pH were also observed between years in

L222. Epilimnetic temperature and concentrations of other dissolved and particulate nutrients did not vary significantly in the lakes over the course of the study (Appendix 4).

Annual means for all measured water quality parameters are presented in Appendix 3.

105 4.5.2 Periphyton biomass and Ag accumulation

After four weeks of colonization, periphyton had developed on the artificial substrates in L221 and L222, with an average biomass (organic matter content) of 10.66 mg tile-1 and 11.72 mg tile-1, respectively (Fig. 4.1). Exposure to AgNPs did not affect the development of periphyton biomass on the tiles (LME models, F1,8 = 0.001, P = 0.982).

In fact, seasonal variations in periphyton biomass in L221 were strongly correlated with fluctuations observed in L222 (repeated measures correlation, r = 0.45, CI [0.14, 0.69], P

= 0.005). Ag was detected in the periphyton within the first month of sampling, with highest levels recorded in August 2014 and 2015 (mean ± SE; 8.29 ± 1.71 µg Ag tile-1 and 7.72 ± 2.07 µg Ag tile-1, respectively), and lowest in June and October 2015 (mean ±

SE; 2.67 ± 1.09 µg Ag tile-1 and 2.54 ± 0.50 µg Ag tile-1, respectively) (Fig. 4.1). Due to high variability in the data, however, these seasonal fluctuations were not significant

(LME models, F3,32 = 2.402, P = 0.086). Copepoda densities exhibited a strong positive correlation with concentrations of periphyton Ag in L222 (repeated measures correlation, r = 0.438, CI [0.01, 0.73], P = 0.037), whereas densities of Cladocera were negatively correlated with periphyton Ag; however, these correlations were not significant (repeated measures correlation, r = -0.137, CI [-0.54, 0.31], P = 0.534).

4.5.3 Littoral microcrustacean community structure

4.5.3.1 Microcrustacean density and diversity

The littoral microcrustacean communities of L221 and L222 were fairly diverse with a total of twenty-seven species identified throughout the duration of the study, with

18 species (67%) corresponding to Cladocera, and 9 (33%) to Copepoda. Both our experimental and reference lake had all species present, with the exception of Pleuroxus

106 procurvus (found only in L222), and Leptodiaptomus spp. (found only in L221)

(Appendix 5). The most abundant species of Cladocera across L221 and L222 were Sida crystallina, Bosmina longirostris, Chydorus cf. brevilabris, and Alona cf. rustica, whereas Copepoda were dominated mainly by Tropocyclops spp., Acanthocyclops spp. and Cyclopoida copepodites. Copepoda were dominated by large numbers of Cyclopoid copepodites throughout the seasons, while adults were mainly represented by

Acanthocyclops spp., Tropocyclops spp., Eucyclops spp. and Microcyclops spp.

Cladocera also numerically dominated the community across all sampling dates in both lakes (with the exception of October 2015 in L222, where Copepoda dominated the community with 69%), accounting for 53% to 83% of the microcrustacean community in

L221 and 56% to 80% in L222. To examine differences in microcrustacean densities in more detail, we used nested linear mixed-effects models. Cladocera densities exhibited a significant main effect of treatment and year (LME models, F1,53 = 12.776, P < 0.001;

F2,53 = 4.147, P = 0.021, respectively), with abundances overall higher in L221, and a decline exhibited between 2014 and 2015 (Tukey’s HSD, P < 0.05). On the other hand, densities of Copepoda were not affected by treatment or time.

Temporal variations in Shannon diversity, species richness, and evenness in L221 and L222 are presented in Figure 4.2. Whereas only slight temporal changes were seen for evenness (LME models, F2,54 = 0.455, P = 0.637), Shannon diversity and species richness exhibited larger fluctuations. Particularly noticeable was the decline in richness in October 2015 in our experimental lake, where the average number of littoral microcrustacean species in the samples, otherwise ranging from 16 to 23, dropped to 5

(Fig. 4.2). Mixed models revealed a strong effect of year for both diversity and species

107 richness (LME models, F2,45 = 14.561, P < 0.001 and F2,53 = 11.971, P < 0.001), with a strong decline observed in 2015 compared to prior years for both community metrics

(Tukey’s HSD, P < 0.05). Species richness also exhibited a significant main effect of treatment (LME models, F1,53 = 12.908, P < 0.001), with a higher number of species found in our reference lake (Fig. 4.2); however, no interaction over time was observed for biodiversity metrics. Feltz and Miller’s asymptotic test on the equality of coefficients of variation from multiple groups confirmed these results, with variations in species richness marginally differing over time across both lakes (test statistic = 5.760, P = 0.056). No differences in variability between lakes (i.e. as a result of AgNP treatment) were observed for any other community metric (Table 4.1).

Mean annual densities of littoral microcrustaceans were consistently higher in our control lake (0.024 – 0.059 individuals cm-2) than in our experimental lake (0.018 – 0.029

-2 individuals cm ) (LME models, F1,53 = 9.304, P = 0.004), with maximum mean annual values measured in 2013 for both lakes (Fig. 4.3). From August 2013 to 2015, densities exhibited relatively consistent seasonal variations, ranging between 0.05 and 0.07 individuals cm-2 in L221, and 0.024 and 0.035 individuals cm-2 in L222. In October 2015, however, there was a noticeable decline, where mean densities dropped by more than 10- fold in both study lakes (Fig. 4.3). Mixed effects models confirmed a significant main effect of year on microcrustacean density (F2,53 = 4.240, P = 0.020), with densities significantly lower in the final year of the study than in prior years (Tukey’s HSD, P <

0.05).

Chydoridae, Cyclopoida and Bosminidae were among the most abundant taxa in both lakes. Although no differences were detected in Chydoridae densities between

108 treatments (LME models, F1,54 = 1.706, P = 0.197), this family exhibited a significant main effect of year (LME models, F2,54 = 3.832, P = 0.028), as well as a significant interaction effect between treatment and year (LME models, F2,54 = 3.812, P = 0.028), with densities in our experimental lake declining by 72%, from 0.374 ind cm-2 in 2013 to

0.106 ind cm-2 in 2015 (Tukey’s HSD, P < 0.05). Furthermore, Bosminidae exhibited a significant treatment main effect (LME models, F1,53 = 7.164, P = 0.010), with densities overall lower in our experimental lake. A strong effect of year was also observed (LME models, F2,53 = 7.439, P = 0.001), with populations of Bosmina longirostris exhibiting a marked increase in density between 2013 and 2014, followed by a decline between 2014 and 2015 (Tukey’s HSD, P < 0.05; Fig. 4.4). This decline was particularly noticeable in our reference lake, where densities dropped 188-fold from 0.400 ind cm-2 in 2014 to

0.002 ind cm-2 in 2015 (Fig. 4.4). Sididae exhibited a significant main effect of year

(LME models, F2,53 = 3.179, P = 0.049), increasing in average annual density by over

80% by the end of study in both lakes (Fig. 4.4). All other microcrustacean taxa in the lakes exhibited high temporal stability, with no distinct differences observed between treatments or over time (Fig. 4.4).

4.5.3.2 Multivariate analyses

Environmental variables including DOC, TDN, TDP, seston C, seston N, seston

P, epilimnetic temperature, Chl-a, TAg, periphyton Ag and pH were used in the BIO-

ENV procedure to determine which variables (or combination of variables) best explained patterns in community assemblages. The BIO-ENV procedure showed that the best combination of environmental variables with the highest correlation with littoral microcrustacean community structure (ρ = 0.452) was a combination of TAg,

109 temperature, pH and Chl-a (Table 4.2; Fig. 4.5). When examining each lake separately, taxonomic composition in our reference lake was most strongly correlated with DOC,

TDP, pH and Chl-a (ρ = 0.268), whereas in our experimental lake, microcrustacean community structure was best described by a combination of TDP, seston P, temperature and TAg (ρ = 0.586). However, despite the strong correlation between community structure and AgNP (i.e. TAg) concentrations, the NMDS ordination (Fig. 4.6), with an acceptable stress value of 0.137, showed that littoral microcrustacean assemblages were primarily structured by natural variation in the lakes, as no distinct groupings between treatments were observed. These observations were confirmed by ANOSIM analyses, revealing no differences between assemblages in our experimental and control lake

(ANOSIM, R = 0.006, P = 0.424). Furthermore, no significant temporal variation was observed within treatments (ANOSIM, R = 0.081, P = 0.202).

4.6 Discussion

This field study is the first to investigate the effects of AgNP exposure on the structure and composition of a natural littoral microcrustacean community. Both littoral and pelagic microcrustacean communities can be strongly impacted by increased levels of metals in the environment (Havens, 1991; Yan et al., 2004; Gagneten, 2010; Palmer et al., 2013; Labaj et al., 2015). Some of the most consistent responses to metal exposure are decreases in species richness, diversity, biomass and density (Gagneten, 2010; Palmer et al., 2013; Moreira et al., 2016). Our study, however, found that these biodiversity metrics were unaffected by exposure to AgNPs. Instead, noticeable declines in species richness and density were observed across both our experimental and reference lake by

110 the end of the study, suggesting that natural processes, such as seasonal changes in temperature, dissolved nutrients, or food availability, may have impacted these changes.

The lack of significant differences in the coefficients of variation between our experimental and reference lake further confirms that exposure to AgNPs did not trigger the observed changes. However, the increases in temporal variability (i.e. decreases in stability) of species richness observed across lakes confirmed by Feltz and Miller’s asymptotic test should indeed be taken into consideration, as it is possible that larger variations in biodiversity metrics and community structure such as these could be an early indicator of an impending regime shift, as increases in variance within systems have been shown to precede larger-scale changes (Brock & Carpenter, 2006; Biggs et al.,

2009; Eason et al., 2013).

Another possible explanation for the sudden decrease in littoral microcrustacean species richness and densities such as the one seen in our study could have been physical disturbance, such as an increase in wind or wave action. Turbulence from wave action can detach parts of the periphytic community from substrates and disperse them throughout the water column, causing large reductions in algal densities (Hoagland &

Peterson, 1990; Kreuzinger-Janik et al., 2015). Another possible explanation for the rapid decline in littoral microcrustacean densities could have been a decrease in the quantity or quality of available food, a common limiting factor in the development of zooplankton

(Kerfoot, 1980). However, as there were no observable decreases in chlorophyll concentrations or periphyton biomass compared to prior years, this explanation may not be likely. Biological interactions such as predation or competition could have also impacted littoral microcrustacean communities in both lakes. An increase in predation

111 pressure by predatory Cladocera or Copepoda, larger invertebrates (e.g. Chaoborus spp.), or zooplanktivorous fish can have a strong impact on shaping microcrustacean community structure, as these predators can be selective with their prey items, causing species that are less frequently preyed upon to become more abundant within the community (Lynch, 1979; Kerfoot, 1980). Furthermore, different stages in the life cycle of microcrustacean species may have exhibited varying sensitivities to certain environmental variables (Levin et al., 2012). It is also possible that there are other unmeasured variables that had a strong impact on the dynamics of the littoral communities in these lakes.

To identify any subtler effects of low-level AgNP exposure that may not be easily detectable through changes in biodiversity metrics, we examined shifts in the abundances of species in more detail, as these are considered to be a good indicator of changes within the community, even if species composition remains the same (Levin et al., 2012). While densities of most microcrustacean families present within the lakes did not exhibit any significant temporal changes, densities of Bosminidae and Sididae exhibited greater variations; however these were found to be synchronous over time across both lakes. The family Chydoridae (consisting of nine species), on the other hand, was the only taxon to exhibit a significant decline following AgNP addition solely in our experimental lake, which was contrary to our expectations. Densities of Chydoridae, otherwise dominant in

L222 both in density and species richness, declined rapidly in the final year of AgNP addition, with all nine species of Chydoridae completely disappearing from the community in October 2015. These shifts in Chydoridae were not observed in our reference lake, where relative abundances of this taxon remained stable over time,

112 suggesting that Chydoridae may have been negatively affected by the presence of

AgNPs, and replaced by metal-tolerant Cyclopoida (Yan & Strus, 1979; Kerrison et al.,

1988; Monteiro et al., 1995), as Cyclopoida became the most dominant taxon in the community in L222 in 2015. Our data contrast sharply with studies in freshwater systems that have found Chydoridae to be highly tolerant to metal contamination (Yan et al.,

2004; Bossuyt & Janssen, 2005; Muyssen et al., 2005; Labaj et al., 2015), and are instead similar to Koivisto et al. (1996) who reported steep declines in Chydorus sphaericus biomass following chronic cadmium exposure.

The selective decline in Chydoridae may have been due to the fact that

Chydoridae mainly include species that feed by crawling on substrata and scraping off attached algae (Fryer, 1968), while other taxonomic groups that were abundant in the community (e.g. Sididae, Bosminidae, Cyclopoida) mainly include filter-feeding and predatory species (Kerfoot, 1980). As AgNPs accumulated in high concentrations within the periphyton, Chydoridae may have therefore been exposed to higher levels of AgNPs than filter-feeding or predatory taxa. Our results are in support of this, as filter-feeding

Sididae, Daphniidae, Bosminidae varied synchronously across both lakes and did not appear to be affected by AgNPs. The ecological implications of a lower tolerance of

Chydoridae are that prolonged AgNP exposure could contribute to the dominance of cyclopoid copepods in littoral communities. This shift in microcrustacean community structure could cause an increase in primary production, due to the release of grazing pressure by common epiphyte scrapers (i.e. Chydoridae) (Fryer, 1968). Further,

Chydoridae represent a crucial link in the transfer of energy and nutrients from the littoral to the pelagic zone (Kerfoot, 1980), and can account for up to 98% of the cladoceran diet

113 of freshwater fish (Adamczuk, 2014). Declines in the relative abundance of Chydoridae therefore have the potential to impact higher trophic levels and alter predator-prey interactions. Moreover, the relative increase in the abundance of cyclopoid copepods in the community could cause a further decline in Cladocera in the littoral zone, as adult

Cyclopoida are highly effective predators on Cladocera (Kerfoot, 1980; Soto & Hurlbert,

1991).

In conclusion, AgNPs did not appear to have a strong effect on overall littoral microcrustacean community structure in a freshwater lake. Although an increase in variability was observed over time across lakes, suggesting that larger community shifts may be imminent within these freshwater systems, these changes did not appear to be triggered by AgNPs. The decrease in overall density and richness of all species within the

Chydoridae family recorded in the final year of AgNP addition within our experimental lake may however indicate that this taxon was particularly AgNP-sensitive. This decline in Chydoridae was followed by an increase in the proportion of Cyclopoida within the community, suggesting that community composition fluctuated in response to natural changes in water quality parameters across both lakes, but that metal exposure in L222 also impacted microcrustacean assemblages, allowing for the proliferation of metal- tolerant Copepoda (Yan and Strus, 1979; Kerrison et al., 1988; Monteiro et al., 1995) within the community. Shifts in microcrustacean assemblages such as these may have a larger impact on ecosystem function over time, through reductions in grazing rates or altered predator-prey interactions (Koivisto, 1997). Despite the long-term in situ exposure to AgNPs, microcrustacean community responses became apparent only in the final year

114 of AgNP addition, suggesting that a longer exposure time may be necessary to detect any larger shifts in microcrustacean assemblage structure.

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124 4.8 Tables and Figures

Table 4.1 Summary of tests on the equality of coefficients of variation (CV) in littoral microcrustacean community metrics by treatment. Included are Feltz and Miller’s asymptotic test and Krishnamoorthy and Lee’s modified signed-likelihood ratio (M- SLRT) test.

Variable Test name Test statistic P Asymptotic 0.006 0.941 Density M-SLRT -0.015 1

Shannon -Wiener Asymptotic 0.925 0.336 diversity M-SLRT 0.868 0.352

Asymptotic 2.235 0.135 Species richness M-SLRT 2.130 0.145

Asymptotic 0.139 0.709 Pielou’s evenness M-SLRT 0.118 0.731

125 Table 4.2 Results of the BIO-ENV analysis indicating which combination of environmental variables was best correlated with littoral microcrustacean community composition in our study. Variables in bold indicate the best match, and are presented as vectors in Figure 4.5.

Best variable combination Rho (ρ)

TAg, temperature, pH, Chl-a 0.452

Seston C, TAg, temperature, pH, Chl-a 0.418

TDP, seston C, TAg, temperature, pH, Chl-a 0.377

TDP, seston C, TAg, periphyton Ag, temperature, pH, Chl-a 0.317

DOC, TDP, seston C, TAg, periphyton Ag, temperature, pH, Chl-a 0.271

DOC, TDN, TDP, seston C, TAg, periphyton Ag, temperature, pH, Chl-a 0.224

DOC, TDN, TDP, seston C, seston P, TAg, periphyton Ag, temperature, pH, Chl-a 0.171

DOC, TDN, TDP, seston C, seston P, seston N, TAg, periphyton Ag, 0.122 temperature, pH, Chl-a

126 Figure Captions

Figure 4.1 Temporal fluctuations in periphyton biomass (mg organic matter tile-1) and periphyton Ag concentrations (µg Ag tile-1) in our experimental lake (Lake 222) during the years of AgNP exposure (2014 – 2015). Values are means of locations (n = 5) on each sampling date ± standard error. Note the differently scales axes.

Figure 4.2 Temporal fluctuations of diversity metrics (Shannon diversity (H’), species richness (S), and Pielou’s evenness (J’)) of littoral microcrustacean communities in the reference lake (Lake 221) and experimental lake (Lake 222) at IISD – ELA over a three year study period.

Figure 4.3 Average annual density (no. individuals cm-2) of littoral microcrustaceans collected from tiles in our reference lake (Lake 221; L221) and experimental lake (Lake

222; L222) over a period of three years (2013 – 2015). AgNP additions to Lake 222 occurred throughout 2014 and 2015. Values are means of locations (n = 5) for each sampling date ± standard error.

Figure 4.4 Relative abundances of littoral microcrustaceans (number of individuals cm-

2) in our reference lake (Lake 221) and experimental lake (Lake 222) over the three-year duration of the experiment. AgNP additions to Lake 222 occurred throughout 2014 and

2015.

Figure 4.5 Non-metric multidimensional scaling (NMDS) ordination displaying littoral microcrustacean species (based on the Bray-Curtis dissimilarity matrix) and fitted environmental variables that were best correlated to microcrustacean community structure (BIO-ENV procedure). The ordination is based on data collected from a reference (Lake 221) and experimental lake (Lake 222) at the IISD – ELA over a period

127 of three years prior to (2013) and during AgNP addition (2014 – 2015). Abbreviations for littoral species: Bos long – Bosmina longirostris, Alo quad – Alona quadrangularis/affinis, Alo rus – Alona cf. rustica, Alo int – Alona intermedia, Aln nana

– Alonella nana, Chy brev – Chydorus cf. brevilabris, Chy pig – Chydorus piger, Cer spp

– Ceriodaphnia spp., Acaleb cur – Acantholeberis curvirostris, Oph grac – Ophryoxus gracilis, Poly ped – Polyphemus pediculus, Sid crys – Sida crystallina, Dis leei –

Disparalona leei, Acacyc spp – Acanthocyclops spp., Tropo spp – Tropocyclops spp.,

Euc spp – Eucyclops spp., Micro spp – Microcyclops spp., Cyc cop – Cyclopoida copepodites.

Figure 4.6 Non-metric multidimensional scaling (NMDS) ordination based on the Bray-

Curtis dissimilarity matrix of littoral microcrustacean abundances in our experimental

(Lake 222) and reference lake (Lake 221). The ordination displays changes in community composition following a two-year (2014 – 2015) experimental addition of AgNPs to Lake

222 at the IISD – Experimental Lakes Area. Data points are sampling locations from

2013 to 2015. Black symbols represent Lake 221; white symbols represent Lake 222.

128 30 12 L221 L222 25 10

) Ag ) -1 -1 tile tile 20 8 μg Ag 15 6

10 4 Organic matter (mg Periphyton Ag ( 5 2

0 0 Jun-14 Jul-14 Aug-14 Oct-14 Jun-15 Jul-15 Aug-15 Oct-15

Figure 4.1

129 Lake 221 Lake 222

18 16 14 12 10 8 6

Species richness (S) 4 2 0

2.5

1.5

0.5 Shannon-Wiener diversity (H')

1.2 1

0.8

0.6

0.4 Pielou's evenness (J')

0.2

Figure 4.2

130

Figure 4.3

131 Lake 221 100% Harpacticoida Cyclopoida 80% Calanoida 60% Sididae Polyphemidae 40% Macrothricidae Daphniidae microcrustacean taxa 20%

Relative abundance of littoral Chydoridae

0% Bosminidae 2013 2014 2015

Lake 222 100% Harpacticoida Cyclopoida 80% Calanoida

60% Sididae Polyphemidae 40% Macrothricidae Daphniidae microcruxstacean taxa 20%

Relative abundance of littoral Chydoridae

0% Bosminidae 2013 2014 2015

Figure 4.4

132 Poly ped Non-metric fit R2 = 0.981 Sid crys Linear fit R2 = 0.914 Acaleb cur Stress = 0.137

0.4 Temperature Alo rus Alo quad

Alo int Cer spp

0.2 Oph grac Micro spp Tropo spp pH NMDS2 0.0

Chy pig TAg Dis leei

− 0.2 Bos long Acacyc spp Aln nana

− 0.4 Chlorophyll a

−0.5 0.0 0.5 1.0

NMDS1

Figure 4.5

133 Non-metric fit R2 = 0.935 Linear fit R2 = 0.690

0.6 Stress = 0.256 0.4 0.2 0.0 NMDS2 − 0.2

− 0.4 2013 2014 2015

− 0.6 Lake 221 Lake 222

−1.0 −0.5 0.0 0.5 1.0

NMDS1

Figure 4.6

134 CHAPTER 5

5.1 Preface

Title: Responses of lower trophic levels to a whole lake addition of silver nanoparticles

Authors: Katarina A. Cetinic1, Michael Paterson2, Beth C. Norman3,4, Daniel C. Rearick1,

Paul Frost3, Marguerite A. Xenopoulos3

Contact Information:

1Environmental and Life Sciences Graduate Program, Trent University, Peterborough,

Ontario, Canada

2International Institute for Sustainable Development – Experimental Lakes Area,

Winnipeg, Manitoba, Canada

3Department of Biology, Trent University, Peterborough, Ontario, Canada

4Present address: Department of Microbiology and Molecular Genetics, Michigan State

University

Corresponding author: Katarina A. Cetinic, [email protected], 705-748-1011 ext.

6467

Status: In preparation for submission

Author Contributions: KAC, MP, BC, and MAX conceived and designed the experiments. KAC conducted the experiments and processed the samples. KAC, MP and

MAX analyzed the data. KAC wrote the manuscript. MAX supervised the project. All authors provided critical feedback and helped with paper revisions.

Keywords: Silver nanoparticles, zooplankton, littoral microcrustaceans, benthic macroinvertebrates, Experimental Lakes Area, community composition, lake ecosystem

135 CHAPTER 5

Effects of low concentrations of silver nanoparticles on natural zooplankton

communities in a small boreal lake

5.2 Abstract

The rapid increase in the use of silver nanoparticles (AgNPs) in commercial products has prompted the need for field-based studies examining the potential impacts these nanoparticles could have on natural ecosystems. Lower trophic levels have been shown to be highly sensitive to AgNPs in laboratory settings; however only few studies have investigated these impacts in situ. This study evaluated the effects of AgNP exposure (< 10 µg L-1) on freshwater zooplankton communities in Lake 222 (L222) relative to a reference lake (Lake 221; L221) at the IISD – Experimental Lakes Area over the course of two years. Overall, we found little evidence to suggest that natural zooplankton communities were impacted by AgNP exposure at environmentally relevant concentrations. No changes were observed in zooplankton density, diversity, species richness or evenness over the course of the study. Furthermore, multivariate analyses detected significant differences in community composition between treatments (i.e. experimental and reference lake) and over time; however, these differences were apparent from the start of the experiment, suggesting that natural processes, and not AgNP exposure, were primarily responsible for the observed differences. The results of this study indicate that zooplankton communities were not sensitive to chronic exposure of low levels of AgNPs in a boreal lake, demonstrating the importance of conducting large-

136 scale studies that incorporate natural seasonal dynamics, as AgNP toxicity may be mitigated in a natural environment.

Keywords: Silver nanoparticles, pelagic microcrustaceans, zooplankton, Experimental

Lakes Area, community composition, lake ecosystem

5.3 Introduction

Recent advances in nanotechnology have led to the increased use of nanoparticles in industry and consumer products (Vance et al., 2015). Silver nanoparticles (AgNPs) are the most frequently used nanoparticles in commercial use today due to their potent antimicrobial properties (Fabrega et al., 2011; Lara et al., 2011; Vance et al., 2015).

Studies have repeatedly shown AgNPs to be highly toxic to a broad range of aquatic organisms (e.g. bacteria, algae, zooplankton, macrophytes, benthic macroinvertebrates and fish) (Navarro et al., 2008; Asghari et al., 2012; Bian et al., 2013; Buffet et al., 2014;

Das et al. 2014; Gambardella et al., 2015; Brittle et al., 2016), and with the increased use of AgNPs projected to lead to elevated discharges into aquatic environments (Maes et al.,

2009), it is vital we investigate the effects these nanomaterials could have on natural ecosystems.

The responses of zooplankton to AgNP exposure have been extensively studied; however, the majority of these experiments have been conducted on model species (i.e.

Daphnia spp.) under controlled laboratory conditions (Zhao & Wang, 2011; Asghari et al., 2012). These short-term single-species toxicity tests are crucial for understanding the mechanisms of AgNP toxicity; however, it is difficult to predict how these findings will translate on a larger scale. One of the reasons for this is that ambient environmental

137 variables (e.g. dissolved organic carbon, pH, UV radiation, ionic strength, and the presence of ligands) have been shown to chemically transform and alter the toxicity of

AgNPs (Lowry & Casman, 2009; Aiken et al., 2011; Stebounova et al., 2010; Xiu et al.,

2011; Liu et al., 2012; Levard et al., 2012; Poda et al., 2013).

Studies on the impacts of chronic metal contamination on zooplankton communities have reported strong decreases in total zooplankton abundance, species richness and diversity, as well as higher tolerances of Copepoda to metal-induced stress compared to Cladocera (Yan and Strus, 1979; Kerrison et al., 1988; Monteiro et al., 1995;

Gagneten & Paggi, 2009). Furthermore, zooplankton species have exhibited differential sensitivity to metal contamination, with certain species found to be more sensitive than others (e.g. Holopedium gibberum, Diaphanosoma birgei, Daphnia spp.) (Yan & Strus,

1979; Marshall et al., 1981; Keller & Yan, 1998; Holt et al., 2003; Labaj et al., 2015).

Zooplankton seem to be particularly sensitive to the effects of AgNPs (Zhao & Wang,

2011; Asghari et al., 2012; Hoheisel et al., 2012; Blinova et al., 2013), with lethal median

-1 concentrations (LC50) found to be as low as 2 µg L (Kennedy et al., 2010). Aside from mortality, exposure to AgNPs causes a number of sublethal effects on zooplankton species, such as growth inhibition (Zhao & Wang, 2011), reduced fecundity (Hook and

Fisher, 2001), abnormal swimming behavior (Asghari et al., 2012), and reduced predator avoidance (Pokhrel & Dubey, 2012). However, only a few studies have investigated the effects of AgNPs on natural plankton communities (Conine, 2017; Tsiola et al., 2017;

Vincent et al., 2017), with no studies to date examining the longer-term impacts of chronic in situ AgNP exposure on zooplankton community structure. Zooplankton communities represent an important trophic link between lower-level primary producers

138 and higher-level consumers, shaping phytoplankton assemblages and regulating their productivity (Gołdyn & Kowalczewska-Madura, 2008), while serving as a food source for aquatic invertebrates and fish (Lynch, 1979). Any changes in zooplankton community structure can therefore have far-reaching effects on the structure and function of the entire freshwater system.

The objective of our study was to examine the effects of long-term AgNP exposure on natural zooplankton communities in a freshwater lake. We hypothesized that

AgNP addition to a freshwater lake would adversely affect metal-sensitive zooplankton taxa, subsequently leading to a decline in zooplankton density, diversity and species richness, and altered community composition compared to a reference lake that was not exposed to AgNPs. To address this hypothesis, environmentally relevant concentrations of AgNPs (~10 µg L-1) were released into an experimental lake at the International

Institute for Sustainable Development – Experimental Lakes Area (IISD – ELA) over the course of two years (2014 – 2015). Changes in zooplankton assemblages in our experimental lake were compared with community changes in our reference lake, to determine if variations came as a response to AgNP exposure, or if they were due to natural seasonal variation over time.

5.4 Methods

5.4.1 Study area The whole lake study was conducted using an experimental lake (Lake 222 – abbreviated L222) and a reference lake (Lake 221; abbreviated L221) at the IISD

Experimental Lakes Area in Northwestern Ontario, Canada (49°40’N, 93°44’W) over the course of three years (2013 – 2015). Both lakes are dimictic and oligotrophic, and

139 relatively small in size (L222: 16 ha, Zmax = 6 m; L221: 9 ha, Zmax = 5.7 m; Fig. 3.1). The lakes were selected based on their geographic proximity and their similar chemical and physical characteristics, so as to minimize inter-system variability. Prior to the start of the experiment, zooplankton species recorded in L222 included Epischura lacustris,

Mesocyclops edax, Tropocyclops extensus, Orthocyclops modestus, Daphnia pulex,

Ceriodaphnia lacustris, Holopedium gibberum, Diaphanosoma birgei, and Chaoborus sp. (IISD – ELA databases). Pelagic zooplankton species were not previously recorded in

L221.

5.4.2 Whole lake AgNP addition

Whole lake AgNP additions took place in L222 during the ice-free season of 2014 and 2015. Polyvinylpyrrolidone (PVP)-coated AgNPs were purchased from NanoAmor

(Houston, TX, USA) as a powder, and were mixed into suspension prior to addition to the lake. A detailed description of the suspension method can be found in Chapter 3 (3.3.2

Silver nanoparticles and whole lake dosing) and in Martin et al. (2017). Twelve litres of newly made AgNP suspension (nominal concentration 5 g L-1) was added to L222 from a point source on the southwestern shore of the lake every other day from 14 June to 23

October 2014 and from 15 May to 25 August 2015. These additions resulted in overall low, environmentally relevant concentrations (under 10 µg L-1) of total Ag (TAg) across the whole lake.

5.4.3 Collection and processing of lake water samples

Both lakes were sampled for water quality, thermal profiles, Ag concentration, and zooplankton composition during the ice-free season 2013 – 2015. Surface water temperatures were measured daily from centre buoy in L221 and L222 using HOBO®

140 temperature loggers (Onset Computer Corp.). Water samples for dissolved (< 0.2 µm) organic carbon (DOC), total dissolved nitrogen (TDN) and phosphorus (TDP), particulate

(35 – 0.7 µm) carbon (C), nitrogen (N) and phosphorus (P), chlorophyll-a (Chl-a), and total Ag (TAg) were collected from centre buoy. Water samples were filtered if necessary, then appropriately stored until further analysis. The collection and laboratory analysis of whole lake nutrients and total Ag (TAg) concentrations followed the same protocol as seen in Chapter 3 (3.3.2. Silver nanoparticles and whole lake dosing; 3.3.3.

Collection and processing of lake water samples).

5.4.4 Zooplankton sampling and laboratory analyses

Zooplankton were collected using an 80 µm zooplankton net from the deepest part of the lake (centre buoy) in August and October 2013, and then in biweekly intervals in 2014 (3 June – 31 July), and monthly intervals in 2015 (11 May – 17 August) (Fig.

3.1). Two samples were collected on each sampling date, with each sample consisting of three vertical tows taken from a depth of 4 m. In the laboratory, one sample was randomly split using a zooplankton wheel, with half of the split sample preserved in a glass jar with 4% sugar formalin for further identification and enumeration. The other half of the split sample was combined with the second full sample, vacuum-filtered through an 80 µm mesh screen, and placed in labeled weigh boats in a drying oven at

60°C for zooplankton carbon, nitrogen and phosphorus content. Zooplankton were identified and counted using a Motic BA 400 compound light microscope (40x magnification). Taxonomic keys that were used included Smith & Fernando (1978) for

Copepoda, and Thorp & Covich (2009) for Cladocera. Where possible, individuals were

141 identified to species, otherwise to genus. All zooplankton samples were processed in their entirety.

To determine Ag concentrations in zooplankton, samples were collected weekly in 2014 (5 June – 21 August) and 2015 (12 May – 24 August), with an additional sampling date in October 2013 – 2015. A single sample consisting of three vertical tows from 4 m was collected on each sampling date, and then dried at 60°C using the method described above. The dried samples were preserved in 70% nitric acid and heated at

120°C for two hours prior to Ag analysis by ICP-MS (ICP-MS X-Series 2, Thermo

Scientific, Nepean, ON). A detailed description of analytical methods for zooplankton Ag analysis can be found in Rearick et al. (2017).

5.4.5 Statistical analysis

Community metrics that were analyzed included: total densities of individuals, species richness (S [Whittaker, 1972]), Shannon-Wiener diversity index (H’ [Shannon and Weaver, 1949]), Pielou’s evenness (J’ [Pielou, 1967]), relative abundance of zooplankton families, and relative abundances of Copepoda and Cladocera species.

Temporal variations in community metrics were assessed as a repeated measures design

(with a single replicate per sampling date) using linear mixed-effects models (LME) within the ‘lme4’ package in R. Treatment (i.e. reference and experimental lake) and the interaction between treatment and year were treated as fixed effects, while sampling dates were included as crossed random effects. Adherence to model assumptions was assessed by examining the dispersion and normality of model residuals. Environmental variables were log (x+1) transformed prior to analyses to meet the assumption of normality. P- values were calculated with an analysis of variance using the Satterthwaite approximation

142 for degrees of freedom with the ‘lmerTest’ package in R. If significant differences were detected between groups, Tukey’s all-pair comparisons were run using the ‘emmeans’ package in R to further examine those differences. The effects of measured physico- chemical parameters (TAg, pH, nutrients, chlorophyll-a, and temperature) on zooplankton biodiversity metrics were also examined using LMEs. To find the most parsimonious models for each biodiversity measure, a simple exhaustive search algorithm

(Morgan & Tatar, 1972) was used within the ‘leaps’ package in R (Lumley & Miller,

2004) to select LME models with the smallest sum of squares (McLeod & Xu, 2017).

For the visual analysis of temporal changes in zooplankton species composition across lakes, non-metric multidimensional scaling (NMDS) (Clarke, 1993) was conducted on untransformed zooplankton abundance data using Bray-Curtis dissimilarity

(Bray and Curtis, 1957) as the distance measure. Any taxa represented in <1.0% of total abundance across samples were considered rare and were excluded from all multivariate analyses. An analysis of similarity (ANOSIM) (Clarke, 1993) was used to identify differences in community composition between treatments and over time. The ANOSIM was performed on the Bray-Curtis dissimilarity matrix using the anosim function within the R software package ‘vegan’ (Oksanen et al., 2017). To identify the species responsible for any major differences between treatments detected by ANOSIM, we conducted a similarity percentage analysis (SIMPER) using the ‘primer’ package in R

(Clarke & Gorley, 2006).

To test for correlations between environmental variables and patterns in zooplankton assemblages (NMDS axes), we conducted the BIO-ENV analysis (Clarke &

Ainsworth, 1993) using the envfit function within the ‘vegan’ package in R (Oksanen et

143 al., 2017). For the visualization of these results, NMDS ordinations were plotted with fitted environmental vectors and zooplankton species (Oksanen et al., 2017).

5.5 Results

5.5.1 Physico-chemical variables

Annual means for environmental variables during the three sampling seasons are presented in Appendix 3. Mixed effects models revealed that pH differed significantly between lakes over the course of the study (Appendix 4), with concentrations overall higher in our experimental lake. pH varied over time in our experimental lake, with values significantly lower in 2014 than in other years (Tukey’s HSD, P < 0.05; Appendix

4). TDP and TDN also exhibited a significant treatment by year interaction (Appendix 4), with TDN concentrations lower in 2013 compared to other years in both L221 and L222, and concentrations of TDP exhibiting a significant increase in L221 by the end of the study (Tukey’s HSD, P < 0.05; Appendix 3). Concentrations of TAg varied between treatments and over time (Appendix 4), with TAg concentrations significantly higher in

L222 in both addition years than in 2013 (Tukey’s HSD, P < 0.05). Although mean TAg also varied between addition years, increasing from an average of 3.04 µg L-1 in 2014 to

9.09 µg L-1 in 2015 (Appendix 3), these differences were not significant, likely due to the large variation in TAg concentrations between sampling dates (Tukey’s HSD, P > 0.05).

5.5.2 Zooplankton biomass and Ag concentration

We detected low levels of Ag in the zooplankton early in the season (0.55 µg Ag g dry mass-1) with Ag levels increasing rapidly and reaching 20-fold higher levels within the next two weeks (10.46 µg Ag g dry mass-1) (Fig. 5.1). The highest Ag values in

144 zooplankton were recorded mid-July in both 2014 (56.18 µg Ag g dry mass-1) and 2015

(35.88 µg Ag g dry mass-1), while the lowest values were measured in the spring months

(May and early June), likely due to biomass dilution (Herendeen and Hill, 2003) as we also found zooplankton biomass to be highest in the spring (Fig. 5.1). Zooplankton biomass decreased in the summer months (July and August), likely due to the decline in edible phytoplankton following intensive grazing by herbivorous zooplankton (Horne and

Goldman, 1994). Despite these changes, mixed models revealed that differences in zooplankton Ag did not vary significantly between addition years (F1,17 = 0.001, P =

0.978), with differences between months only marginally significant (F7,17 = 2.583, P =

0.052), likely due to high variability in the data (Fig. 5.1). Zooplankton biomass, on the other hand, exhibited significant differences between months (F7,17 = 3.891, P = 0.010), with highest levels measured in May 2015 (Tukey’s HSD, P < 0.05; Fig. 5.1).

5.5.3 Zooplankton diversity and species composition

In total, 33,460 individuals were collected from 24 taxa over the course of the study (Appendix 6). The most abundant zooplankton species across both lakes were

Holopedium gibberum, Bosmina longirostris, Tropocyclops spp., Cyclopoida nauplii and

Cyclopoida copepodites (Fig. 5.2; Fig. 5.3). Taxa that were highly abundant only within the reference lake (L221) included Leptodipatomus spp., Diaphanosoma birgei,

Calanoida nauplii and Calanoida copepodites, whereas Daphnia pulex was dominant in the experimental lake (L222) (Fig. 5.2; Fig. 5.3). The average density (mean ± SE) of zooplankton recorded on each sampling date was 4.14 (±1.52) individuals L-1 in L221, and 3.59 (±2.08) individuals L-1 in L222. Zooplankton density was invariant over time

145 and with treatment as indicated by linear mixed-effects models (F2,14 = 0.262, P = 0.773;

F1,14 = 0.325, P = 0.577) (Fig. 5.4).

Zooplankton community metrics (Shannon diversity, species richness and

Pielou’s evenness) varied slightly between sampling dates (Fig. 5.5), but did not exhibit any significant changes across years (i.e. prior to and during AgNP addition) in either lake (LME models, P > 0.05). Furthermore, no marked differences in diversity, richness or evenness were observed as a result of AgNP treatment. The results of the linear mixed effects models showed that zooplankton species richness across treatments was affected by TDP and TAg concentrations (LME models, F1,11 = 16.251, P = 0.002 and F1,11 =

6.017, P = 0.032, respectively), whereas no environmental factors influenced diversity or evenness (LME models, P > 0.05). Repeated measures correlation did not reveal any correlation between zooplankton dry weight and zooplankton Ag (r = -0.360, P = 0.600)

(Fig. 5.1).

The zooplankton families Holopedidae, Bosminidae, Daphniidae and Cyclopoida dominated the zooplankton community in L222, accounting for, on average, 10% to 32% of the community (Fig. 5.6). In L221, the pelagic zooplankton community was dominated by Calanoida, Cyclopoida and Holopedidae, with, on average, 15% to 41% of the community (Fig. 5.6). Of these, only relative abundances of Daphniidae and Calanoida exhibited significant differences between treatments (LME models, F1,11 = 9.199, P =

0.012; F1,11 = 89.023, P < 0.001, respectively), with Daphniidae more abundant in L222, and Calanoida more abundant in L221 (Fig. 5.6). Relative abundances of Daphniidae also exhibited a significant treatment by year interaction (LME models, F2,11 = 5.185, P =

0.031), experiencing a 14-fold increase in L222 in the first year of AgNP addition (LME

146 models, Tukey’s HSD, P < 0.05). Although relative abundances of other zooplankton taxa varied over time in both lakes (Fig. 5.6), these changes were not significant.

Mixed models found that both Copepoda and Cladocera differed significantly between treatments (LME models, F1,11 = 6.499, P = 0.027 and F1,11 = 7.101, P = 0.022, respectively), with Copepoda dominating the zooplankton community in L221, both numerically (accounting for 62% to 79% of the community), and in species richness (12 species) across all years, and Cladocera numerically dominating the community in L222 in the year prior to and during the first year of AgNP addition (with 64 and 74%, respectively) (Appendix 6). In 2015, however, Copepoda became dominant in L222 with annual relative abundance reaching 55%.

5.5.4 Community analyses

Multivariate analysis showed that pelagic zooplankton community composition differed significantly between treatments (ANOSIM, R = 0.472, P = 0.001) and over time

(ANOSIM, R = 0.092, P = 0.037). The NMDS ordination reflected these results, visually exhibiting distinct groupings of data from L221 and L222 (Fig. 5.7). Analysis based on

SIMPER identified Calanoida copepodites, Bosmina longirostris, Holopedium gibberum,

Tropocyclops spp., Diaphanosoma birgei, and Leptodiaptomus minutus as the principal species contributing to over 50% of the average dissimilarity observed between treatments (Table 5.1). Further, when examining which species were responsible for differences detected between years, SIMPER identified seven key species (B. longirostris, Tropocyclops spp., H. gibberum, D. pulex, D. birgei, Cyclopoida copepodites, Calanoida nauplii) responsible for variations in zooplankton community structure in L222, and six species (D. birgei, H. gibberum, L. minutus, Tropocyclops spp.,

147 Calanoida copepodites, Calanoida nauplii) in L221. The BIO-ENV analysis identified the combination of DOC, seston C, pH and TAg as the subset of environmental variables that was best correlated with zooplankton community structure (BIO-ENV, weighted

Spearman rank correlation, ρ = 0.333; Table 5.2; Fig. 5.8). BIO-ENV analyses of each lake individually revealed that a combination of TDP, temperature and pH best explained patterns in zooplankton assemblages in L221 (ρ = 0.371), whereas assemblages in L222 were most strongly correlated with TDP, chlorophyll-a, seston C and seston N (ρ =

0.266).

5.6 Discussion

In this study, we analyzed the effects of AgNP exposure on the zooplankton community of a freshwater boreal lake. Overall, we found little evidence to suggest that natural zooplankton communities were negatively impacted by an addition of AgNPs.

Metal contamination has been shown to cause decreases in zooplankton biomass, species richness and density (Sharma et al., 2000; Gagneten and Paggi, 2009; Vincent et al.,

2017); however, our results showed only limited indication of any effects of AgNPs on these community metrics. Multivariate analyses revealed significant dissimilarities in zooplankton community composition between our experimental (Lake 222) and reference lake (Lake 221); however, these differences were apparent from the start of the experiment (i.e. prior to AgNP addition), suggesting that AgNP exposure was likely not the main cause for the observed dissimilarities, but instead natural variation between the lakes played the dominant role. This was confirmed by the BIO-ENV procedure, which indicated that a combination of environmental variables (dissolved organic carbon,

148 particulate carbon, and pH) and TAg concentrations was primarily responsible for the observed patterns in zooplankton community structure. These results suggest that, in contrast to our predictions, biodiversity metrics and zooplankton community composition were largely unaffected by AgNP exposure in natural waters (Gao et al., 2009; Blinova et al., 2013; Cupi et al., 2016).

Studies have shown that zooplankton species (e.g. Daphnia magna, Moina macrocopa) readily ingest and accumulate AgNPs (Hook & Fisher, 2001; McTeer et al.,

2013; Yoo-iam et al., 2014), which can have strong adverse effects on growth (Zhao &

Wang, 2011) and reproduction (Hook & Fisher, 2011; Zhao & Wang, 2011; Pakrashi et al., 2017). In our study, bacterioplankton, phytoplankton and zooplankton all accumulated Ag following exposure (Rearick, 2017), indicating that AgNPs were bioavailable to organisms in the pelagic zone. The levels of zooplankton Ag measured in our study were comparable to those observed in Zhao & Wang (2011), where average concentrations of Ag in Daphniidae ranged around 20 µg Ag g dry weight-1; however, they were several times higher than those reported in other laboratory studies. Khan et al.

(2015) measured concentrations of Ag in D. magna, and found that accumulated levels ranged around 0.5 µg Ag g dry weight-1, while Hook & Fisher (2001) reported even lower concentrations of accumulated Ag in the calanoid copepods Acartia tonsa and A. hudsonica and the cladocerans Simocephalus spp. and Ceriodaphnia dubia. It is interesting to note that, when normalized to milligrams of carbon (C) to provide biomass- specific Ag content, Rearick (2017) found that levels of Ag within zooplankton were much lower (< 0.15 µg Ag mg C-1) than those measured in bacterioplankton (0.27 – 16.82

µg Ag mg C-1) and phytoplankton (0.17 – 6.45 µg Ag mg C-1). Despite the high

149 bioavailability of AgNPs to all lower trophic levels in our experimental lake (Rearick,

2017), no pelagic zooplankton taxa appeared to be negatively affected. Similarly, no strong negative effects were observed on bacterioplankton (Rearick, 2017) or phytoplankton (Rearick, 2017; Conine et al., 2018) following exposure and accumulation of Ag in our experimental lake. In fact, contrary to our predictions, relative abundances of the family Daphniidae were found to increase during the years of AgNP addition in our experimental lake. Our results differ from numerous studies that have shown this taxon to be extremely sensitive to AgNP exposure in laboratory tests (Griffitt et al., 2008;

Zhao & Wang, 2011; Asghari et al., 2012; Hoheisel et al., 2012; Völker et al., 2013), and are instead similar to Sørensen et al. (2016), who reported no effect on Daphnia magna mortality, body length or molting following acute exposure to 10 – 50 µg L-1 AgNPs.

Furthermore, as part of this whole lake study, Conine (2017) found that algal particles present within the lake aided in the removal of AgNPs from the water column into the sediments, mitigating any toxic effect of AgNPs on Daphnia spp. The marked increase in

Daphniidae observed in Lake 222 could, however, be explained by the timing of our sampling. In 2013, only August and October were sampled, months in which recorded densities of D. pulex were lower across all years, likely due to increased predation by young-of-year perch (Wu & Culvez, 1994), whereas in 2014 and 2015, sampling took place starting in Spring, when much higher densities were recorded (May – July).

Metal contamination has also been shown to impact the size structure of zooplankton communities. For example, Vincent et al. (2017) found that, following

AgNP exposure, larger crustacean taxa (e.g. Sida crystallina) declined in densities, and smaller-bodied crustacean taxa (e.g. Bosmina longirostris) exhibited a higher tolerance to

150 metal contamination and increased in abundance. Gagneten (2010) reported similar findings, with larger sized zooplankton declining in abundances after exposure to heavy metal contamination. Although densities of B. longirostris in our study did not differ between lakes, immature (i.e. smaller-bodied) Copepoda did appear to dominate the community by the end of the experiment. This shift in community structure in Lake 222 from Cladocera-dominated to Copepoda-dominated by the end of the experiment could be an indication that Cladocera are more sensitive to AgNP exposure than Copepoda in a natural environment. These results are consistent with numerous laboratory and field- based studies, which show Cladocera to be more sensitive to environmental contamination and other anthropogenic stressors (Roch et al., 1985; Gulati et al., 1988;

Kerrison et al., 1988; Yan et al., 2004; Gagneten & Paggi, 2009), but contrast with the mesocosm study conducted by Vincent et al. (2017) in which Cladocera dominated the zooplankton community regardless of AgNP treatment. If the observed shift is indeed a result of AgNP exposure, prolonged exposure could have a profound effect on ecosystem function, leading to the proliferation of phytoplankton due to the removal of grazers (e.g.

Holopedium gibberum, Bosmina longirostris) from the system, as well as increased predatory pressure on Daphnia pulex as one of the larger-bodied Cladocera remaining in high densities (Kerfoot, 1980).

The lack of a strong negative response of zooplankton communities to AgNP exposure could be due to the natural water chemistry of the lake mitigating the toxic effects of AgNPs. For example, Seitz et al. (2015) reported that the toxicity of AgNPs to

Daphnia magna was reduced in media with higher pH and the presence of dissolved organic matter. Other studies have also reported on the mitigating effects of DOC on the

151 biotoxicity of AgNPs. Blinova et al. (2013) and Cupi et al. (2015) both found that elevated concentrations of DOC reduced overall AgNP toxicity to D. magna. The interaction between AgNPs and ligands such as DOC can increase the stability of AgNPs, due to the capacity of these molecules to bind to toxic Ag+ ions, decreasing their dissolution into the water column and subsequently reducing the toxic effect of AgNPs

(Bianchini & Wood, 2008; Gao et al., 2009; Levard et al., 2012). It is therefore possible that the high pH levels and concentrations of dissolved organic matter measured in our experimental lake increased AgNP stability and mitigated its toxic effects to zooplankton

(Gao et al., 2009; Seitz et al., 2015).

There is a substantial amount of research on the effects of metal contamination on aquatic systems. However, due to the multivariate nature of zooplankton communities, it is difficult to determine distinct patterns of change in community composition. In our study, no diversity metrics were directly related to AgNP concentration, but were instead largely dependent on ambient physico-chemical characteristics. Due to the central position of microcrustaceans within the aquatic food web, any changes in zooplankton community structure could have far-reaching effects on the structure and function of the entire ecosystem. It is therefore important that in situ studies on the effects of AgNPs continue, preferably over longer time spans.

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160 5.8 Tables and Figures

Table 5.1 Similarity percentages (SIMPER) analysis indicating which zooplankton species were primarily responsible for the observed dissimilarity (as indicated by ANOSIM results) in community composition between our reference (Lake 221) and experimental (Lake 222) lake. Zooplankton species are ranked according to their average contribution to dissimilarity between assemblages in Lake 221 and Lake 222. Average abundances within each lake and percentage of cumulative similarity are also included.

Average Average Average Cumulative abundance abundance contribution similarity (L221) (L222) (%) (%) ANOSIM, R = 0.472, P = 0.001

Calanoida copepodites 506.0 13.3 16.7 23.6 Bosmina longirostris 172.4 327.2 10.2 38.0

Holopedium gibberum 281.8 324.7 9.4 51.4

Tropocyclops spp. 126.3 190.6 7.6 62.2

Diaphanosoma birgei 128.8 76.1 4.9 69.2

Leptodiaptomus minutus 139.8 3.0 4.1 75.0

161 Table 5.2 Results of the BIO-ENV analysis indicating which combination of environmental variables from a total of 10 were best correlated with zooplankton assemblage structure prior to and during a two-year (2014 – 2015) experimental addition of AgNPs to an experimental lake at the IISD – ELA. Variables in bold indicate the best match, and are presented as vectors in Fig. 5.8.

Best variable combination Rho (ρ)

Seston C, pH, TAg 0.332

DOC, seston C, pH, TAg 0.333

DOC, seston C, seston P, pH, TAg 0.332

TDP, DOC, seston C, seston P, pH, TAg 0.314

TDP, DOC, temperature, seston C, seston P, pH, TAg 0.299

TDP, DOC, temperature, seston N, seston C, seston P, pH, TAg 0.281

TDP, TDN, DOC, temperature, seston N, seston C, seston P, pH, TAg 0.255

TDP, TDN, DOC, temperature, seston N, seston C, seston P, 0.184 Chl-a, pH, TAg

162 Figure Captions

Figure 5.1 Temporal fluctuations in zooplankton biomass (mg dry weight) and zooplankton Ag concentrations (ng Ag mg dry weight-1) in our experimental lake (Lake

222) during the years of AgNP exposure (2014 – 2015). Values are means of sampling dates (June and August 2014: n = 3; July and August 2015: n = 4; July 2014 and June

2015: n = 5; October 2014 and 2015: n = 1) ± standard error. Note the differently scales axes.

Figure 5.2 Densities of the six most frequent zooplankton taxa (number of adults L-1; mean ± standard error) recorded in our experimental (Lake 222) and reference lake (Lake

221) over a period of three years (2013 – 2015). AgNP additions to Lake 222 began in

2014 and continued through 2015. Samples were collected from centre buoy throughout the ice-free season, and values are means of sampling dates (2013: n = 2; 2014 and 2015: n = 4). Note the differently scaled y-axes.

Figure 5.3 Densities of immature stages of Copepoda (Cyclopoida and Calanoida copepodites and nauplii) (number of individuals L-1; mean ± standard error) recorded in our experimental (Lake 222) and reference lake (Lake 221) over a period of three years

(2013 – 2015). AgNP additions to Lake 222 began in 2014 and continued through 2015.

Samples were collected from centre buoy throughout the ice-free season, and values are means of sampling dates (2013: n = 2; 2014 and 2015: n = 4). Note the differently scaled y-axes.

163 Figure 5.4 Average annual zooplankton density (no. individuals L-1) in the pelagic zone of our reference (Lake 221; L221) and experimental lake (Lake 222; L222) from 2013 to

2015. AgNP additions to Lake 222 occurred throughout 2014 and 2015. Values are means of sampling dates (2013: n = 2; 2014 and 2015: n = 4) ± standard error.

Figure 5.5 Temporal variation of diversity metrics (Shannon diversity (H’), species richness (S), and Pielou’s evenness (J’)) of zooplankton communities in the experimental

(Lake 222) and reference lake (Lake 221) at IISD – ELA over a three year study period.

AgNP additions to Lake 222 occurred throughout 2014 and 2015.

Figure 5.6 Relative abundances of pelagic zooplankton families in our experimental lake

(Lake 222) and reference lake (Lake 221) over the three-year duration of the experiment

(2013 – 2015). AgNP additions to Lake 222 occurred throughout 2014 and 2015.

Figure 5.7 Non-metric multidimensional scaling (NMDS) ordination based on zooplankton diversity following a two-year (2014 – 2015) experimental addition of

AgNPs to a lake at the IISD – Experimental Lakes Area. Data points are sampling dates from 2013 to 2015. White symbols represent the reference lake (Lake 221); black symbols represent the experimental lake (Lake 222).

Figure 5.8 Non-metric multidimensional scaling (NMDS) ordination with fitted environmental variables that were most strongly correlated to zooplankton community composition (BIO-ENV analysis). The ordination is based on species abundance data collected from a reference (Lake 221) and experimental lake (Lake 222) at the IISD –

ELA over a period of three years prior to (2013) and during AgNP addition (2014 –

2015). Abbreviations for zooplankton species: Daph pul – Daphnia pulex, Trop spp –

Tropocyclops spp., Hol gib – Holopedium gibberum, Bos long – Bosmina longirostris,

164 Lept min – Leptodiaptomus minutus, Diap birg – Diaphanosoma birgei, Meso edax –

Mesocyclops edax, Cer spp – Ceriodaphnia spp., Cal cop – Calanoida copepodites,

Cyc Cop – Cyclopoida copepodites, Cal naup – Calanoida nauplii, Cyc naup –

Cyclopoida nauplii.

165

Figure 5.1

166 3 3

L221 L222 -1

L -1

L 2.5 2.5

2 spp. 2

1.5 1.5 B. longirostris

1 Tropocyclops 1

No. of 0.5 0.5 No. of

0 0 2013 2014 2015 2013 2014 2015

2.5 0.8

0.7 -1 -1 L 2 L 0.6

1.5 0.5 0.4

H. gibberum 1 0.3 Daphnia pulex 0.2 No. of 0.5 No. of 0.1 0 0 2013 2014 2015 2013 2014 2015

1 1.2 -1 -1 L L 0.9 0.8 1 0.7 spp. 0.8 0.6 0.5 0.6 0.4 0.3 0.4 Diaphanosoma birgei Leptodiaptomus 0.2 0.2 0.1 No. of No. of 0 0 2013 2014 2015 2013 2014 2015

Figure 5.2

167 0.8 2 L221 copepodites L221 nauplii 1.6 0.6 L222 copepodites

L222 nauplii

-1 -1 L

L 1.2 0.4 0.8 nauplii nauplii

0.2 0.4 No. of Calanoida copepodites and No. of Cylopoida copepodites and 0 0 2013 2014 2015 2013 2014 2015

Figure 5.3

7 L221 L222 6 ) -1

L 5

2 (no. individuals

Average zooplankton density 1

0 2013 2014 2015

Figure 5.4

168 L221 L222

2.5

1.5

1 Shannon diversity (H') 0.5

18 16 14 12 10 8 6 Species richness (S) 4 2 0

0.8

0.6

0.4

Pielou's evenness (J') 0.2

0 02-Aug 20-Oct 03-Jun 17-Jun 01-Jul 31-Jul 11-May 18-Jun 16-Jul 17-Aug 2013 2014 2015

Figure 5.5

169 Lake 221 1 Diptera

0.8 Cyclopoida

Calanoida 0.6 Sididae

Chydoridae 0.4 Daphniidae

0.2 Bosminidae Holopedidae Relative abundance of zooplankton taxa 0 2013 2014 2015

Lake 222 1 Diptera

Cyclopoida 0.8 Calanoida

0.6 Leptodoridae Sididae

0.4 Chydoridae

Daphniidae 0.2 Bosminidae

Relative abundance of zooplankton taxa Holopedidae 0 2013 2014 2015

Figure 5.6

170

Non-metric fit R2 = 0.963 0.4 Linear fit R2 = 0.801 Stress = 0.192 0.2 0.0 NMDS2 − 0.2

2013 2014 2015

− 0.4 Lake 221 Lake 222

−0.5 0.0 0.5

NMDS1

Figure 5.7

171 2 Non-metric fit R = 0.964 Cer spp Linear fit R2 = 0.820 Stress = 0.190 0.6 0.4

Diap birg

0.2 Daph pul Hol gib TAg

NMDS2 Bos long

0.0 Cyc naup Cal cop Cyc cop − 0.2 DOCLept min Trop spp

Seston C − 0.4

−0.5 0.0 0.5

NMDS1

Figure 5.8

172 CHAPTER 6

General Discussion

To better understand the effects of AgNPs on lower trophic levels under natural conditions, I conducted a six-week mesocosm study and a longer three-year whole lake study at the IISD – Experimental Lakes Area. The general aim of my thesis was to investigate the effects of AgNP exposure on the structure and composition of natural littoral microcrustacean, zooplankton, and benthic macroinvertebrate communities in boreal lakes. Overall, we found only limited evidence to suggest any negative impacts of

AgNPs on natural lower trophic level communities.

The results of our mesocosm study (Chapter 2) showed that benthic macroinvertebrate communities were not highly sensitive to AgNP exposure. I found no effect of AgNPs on benthic macroinvertebrates in stony sediments. As predicted, I detected a higher relative abundance of metal-tolerant Chironomidae (Zimmerman et al.,

1993; Winner et al., 1980; Barbour et al., 1999; Beltman et al., 1999; Loayza-Muro et al.,

2010) in high dose chronic PVP-AgNP treatments compared to control treatments across fine sediment samples; however, these changes were relatively subtle, and no larger differences in benthic macroinvertebrate community structure were detected between treatments.

Our whole lake ecosystem study (Chapters 3 – 5) resulted in similar findings.

Here I demonstrated that chronic exposure to environmentally relevant concentrations of

AgNPs had little to no impact on freshwater microcrustacean and macroinvertebrate communities in a boreal lake. My results show that any changes detected between or within lakes over the course of the experiment did not come as a result of AgNP

173 exposure, but were rather influenced primarily by natural variation in the lakes. The littoral communities of the lakes, both benthic macroinvertebrates (Chapter 3) and microcrustaceans (Chapter 4), exhibited a significant decline in densities and richness in the final year of the whole lake study across both lakes. This synchronous shift across lakes suggests that communities in the littoral zone are likely being influenced by a regional disturbance such as an alteration in weather patterns (Pabst et al., 2008; Levin et al., 2012) or an increase in fish predation (Leppä et al., 2003; Miracle et al., 2006), rather than exposure to AgNPs.

Although a considerable decline in abundances was observed across all taxa present in the littoral zone, the microcrustacean family Chydoridae exhibited a significant drop in relative abundances compared to other taxa solely in our experimental lake, indicating that this taxon may perhaps be more sensitive to AgNP exposure than other microcrustacean taxa in the littoral zone. This may have been due to the feeding behaviour of Chydoridae (i.e. epiphyte scrapers [Fryer, 1968; Thorp & Covich, 1991]), which differs from other taxa in the littoral zone (i.e. filter-feeders [Masclaux et al.,

2012]). These data contrast with studies that have found filter-feeding Bosmina longirostris (Koivisto & Ketola, 1995; Labaj et al., 2015), Sida crystallina (Palmer et al.,

2013) and Daphnia spp. (Kennedy et al., 2010; Zhao & Wang, 2011; Asghari et al., 2012;

Hoheisel et al., 2012) to be highly sensitive to metal exposure, suggesting that filter- feeders in the littoral zone may be less sensitive to the impacts of AgNP than scrapers.

Although these effects were relatively subtle and only observed at the end of our experiment, it is possible that a longer exposure time may be necessary to detect any larger shifts in littoral microcrustacean assemblage structure.

174 The results of my research differ from small-scale laboratory studies reporting a strong toxic effect of AgNPs to freshwater organisms (Griffitt et al., 2008; Khaydarov et al., 2009; Croteau et al., 2011; Buffet et al., 2014; Das et al., 2014; Baptista et al., 2015;

Gambardella et al., 2015; Brittle et al., 2016), and are instead similar to other long-term field-based studies conducted at the IISD – ELA. A study conducted by Conine et al.

(2018) examining the chronic impacts of low levels of AgNPs to natural phytoplankton in a boreal lake found little impact on these communities. Furthermore, Blakelock et al.

(2016) and Rearick (2017) detected no strong effect of chronic AgNP exposure on natural bacterioplankton communities. Zooplankton communities appeared to respond similarly, with only subtle changes in community structure detected following a six-week addition of AgNPs to freshwater mesocosms at the IISD – ELA (Vincent et al., 2017). These studies, combined with the findings of my thesis, suggest that natural communities may be more strongly driven by other internal or external mechanisms, such as natural temporal variability in community composition or seasonal changes in temperature or nutrient concentrations.

Not all field-based studies on the impacts of AgNPs have observed similar results, however. In a marine mesocosm study, Tsiola et al. (2018) observed a significant shift in bacterial community composition, as well as alterations in bacterial and viral processes, suggesting that AgNPs may ultimately have a negative impact on marine food web function. Strong adverse effects were also observed in natural estuarine plankton communities exposed to AgNPs, with declines recorded in growth rates, grazing rates and photosynthetic efficiency (Baptista et al., 2015). These findings could be attributed to the concentrations of AgNPs used in these studies, which were much higher than those used in our experiments and unlikely to be environmentally relevant. Another possible

175 explanation for the discrepancy between AgNP exposure experiments conducted at the

IISD – ELA and other field studies could be the natural water chemistry of the lakes mitigating any toxic effects of AgNPs. For example, DOC has been shown to stabilize the nanoparticles and decrease their toxicity at concentrations as low as 4 mg L-1 (Kennedy et al., 2012). As measured levels of DOC were relatively high in our experimental lake

(8.45 – 15.98 mg L-1), it is possible that this may have reduced the toxicity of AgNPs to freshwater organisms.

This dissertation has helped further our understanding of how freshwater communities respond to AgNP exposure over longer time scales under natural conditions.

Overall, my results indicate that benthic macroinvertebrate, littoral microcrustacean and zooplankton communities are not sensitive to long-term exposure of low levels of AgNPs under environmental conditions, but are rather primarily driven by natural seasonal dynamics and physico-chemical parameters within the lake (Allan, 1995; Becerra-Jurado et al., 2009).

6.1. Future Work

Despite the increased use and potential release of AgNPs into the environment, investigations into the effects of AgNPs on natural communities are still quite rare. The results of my thesis demonstrate the importance of including abiotic variables such as temperature, pH and nutrient concentrations when studying the effects of AgNPs on natural communities, as these factors can have a significant impact both on the behaviour and distribution of the nanoparticles themselves (Kittler et al., 2010; Liu & Hurt, 2010;

Stebounova et al., 2010; Liu et al., 2012; Rearick et al., 2017), as well as on community dynamics in freshwater systems (Kerfoot, 1980; Thorp & Covich, 1991). Although laboratory studies are extremely useful for determining the mechanisms of AgNP

176 toxicity, the discrepancy between our findings and these small-scale studies emphasizes the importance of conducting whole ecosystem studies that incorporate natural ecosystem and community variability for quality risk assessment, as these factors may play a strong role in the bioavailability and toxicity of AgNPs in natural environments. It is therefore vital to continue conducting field-based research that includes a wide range of environmental variables across different geographic regions in order to better understand the impacts AgNPs are having on various types of freshwater systems.

Examining the long-term impacts of different nanoparticle sizes, shapes and surface coatings on natural freshwater communities was beyond the scope of this thesis.

However, more research is necessary in this area, as differences in the physical characteristics of the nanoparticles will likely result in considerable contrasts in the toxicological effects of AgNPs compared to the findings observed in this study. Although all AgNP research conducted at the IISD – ELA came to the conclusion that AgNPs are not toxic to natural communities (from bacterioplankton to fish) (Blakelock et al., 2016;

Rearick, 2017; Vincent et al., 2017; Conine et al., 2018; Martin, unpublished), for a fuller understanding of any threats AgNPs may pose on freshwater systems, more studies are required analyzing different types of AgNPs across different environments.

177 6.2 References

Allan, J. D. (1995). Stream ecology: Structure and function of running waters. Chapman and Hill: New York. 388pp.

Baptista, M. S., Miller, R. J., Halewood, E. R., Hanna, S. K., Almeida, C. M., Vasconcelos, V. M., … Lenihan, H. S. (2015). Impacts of silver nanoparticles on a natural estuarine plankton community. Environmental Science & Technology, 49(21), 12968-12974. doi:10.1021/acs.est.5b03285

178

Conine, A. L., Rearick, D. C., Paterson, M. J., Xenopoulos, M. A., & Frost, P. C. (2018). Addition of silver nanoparticles has no long-term effects on natural phytoplankton community dynamics in a boreal lake. Limnology and Oceanography Letters, 3(4), 311-319. doi:10.1002/lol2.10071

Kerfoot, W. C. (1980) Evolution and Ecology of Zooplankton Communities. University Press, New England. 793pp.

179 Khaydarov, R. R., Khaydarov, R. A., Estrin, Y., Evgrafova, S., Scheper, T., Endres, C., & Cho, S. Y. (2009) Silver nanoparticles: Environmental and human health impacts. In I. Linkov; J. Steevens (eds.) Nanomaterials: Risks and Benefits, 287- 297. Netherlands: Springer. doi:10.1007/978-1-4020-9491-0_22

Kittler, S., Greulich, C., Diendorf, J., Köller, M., & Epple, M. (2010). Toxicity of silver nanoparticles increases during storage because of slow dissolution under release of silver ions. Chemistry of Materials, 22(16), 4548-4554. doi:10.1021/cm100023p

Leppä, M., Hämäläinen, H., & Karjalainen, J. (2003). The response of benthic macroinvertebrates to whole-lake biomanipulation. Hydrobiologia, 498(1-3), 97- 105. doi:10.1023/A:1026224923481

Liu, J., & Hurt, R. H. (2010). Ion release kinetics and particle persistence in aqueous nano-silver colloids. Environmental Science & Technology, 44(6), 2169-2175. doi:10.1021/es9035557

Liu, J., Wang, Z., Liu, F. D., Kane, A. B., & Hurt, R. H. (2012). Chemical transformations of nanosilver in biological environments. ACS Nano, 6(11), 9887-9899. doi:10.1021/nn303449n

Masclaux, H., Bec, A., & Bourdier, G. (2012). Trophic partitioning among three littoral microcrustaceans: relative importance of periphyton as food resource. Journal of Limnology, 71(2), 28. doi:10.4081/jlimnol.2012.e28

180 Pabst, S., Scheifhacken, N., Hesselschwerdt, J., & Wantzen, K. M. (2008). Leaf litter degradation in the wave impact zone of a pre-alpine lake. Hydrobiologia, 613(1), 117-131. doi:10.1007/s10750-008-9477-y

Thorp, J. H. & Covich, A. P. (1991) Ecology and Classification of North American Freshwater Invertebrates. Academic Press, San Diego. 991pp.

Zhao, C., & Wang, W. (2010). Biokinetic uptake and efflux of silver nanoparticles in Daphnia magna. Environmental Science & Technology, 44(19), 7699-7704. doi:10.1021/es101484s

181 Zimmerman, M. C. (1993). The use of the biotic index as an indicator of water quality. Pages 85-98, in Tested studies for laboratory teaching, Volume 5 (C.A. Goldman, P.L. Hauta, M.A. O’Donnell, S.E. Andrews, and R. van der Heiden, eds). Proceedings of the 5th Workshop/Conference of the Association for Biology Laboratory Education (ABLE), 115 pp.

182 7. Appendix

Appendix 1.A Mean abundance (no. individuals/mesocosm) for benthic invertebrate taxa recorded in fine sediment samples in our mesocosm study. Values are averages (n = 2 grab samples) of the abundance values obtained within each mesocosm, with standard error of the mean in parentheses. C-01 – Control 1, C-02 – Control 2, LD-01 – Low drip 1, LD-02 – Low drip 2, MD-01 – Medium drip 1, MD-02 – Medium drip 2, HD-01 – High drip 1, HD-02 – High drip 2, P-01 – Plug 1, P-02 – Plug 2, Cit-01 – High dose citrate 1, Cit-02 – High dose citrate 2.

C- C- LD- LD- MD- MD- HD- HD- P- P- Cit- Cit- Taxa 01 02 01 02 01 02 01 02 01 02 01 02

Chironomidae 235 110 248 276 171 265 84 429 167 191 227 175 (7) (21.5) (20.5) (24) (85.5) (64.5) (21.5) (192) (8) (11) (21) (7.5) Anisoptera 2 1 0 0 0 0 0 0 1 0 0 1 (1.5) (1) (1) (1) Ephemeroptera 5 0 0 0 2 5 0 0 2 3 6 2 (5) (1) (3) (1.5) (3) Gastropoda 15 7 8 7 0 0 5 0 9 0 22 6 (7.5) (3.5) (2.5) (0.5) (1.5) (5) (16) (2) Amphipoda 6 0 0 0 2 2 1 0 4 0 8 12 (2.5) (1.5) (2.5) (0.5) (2) (1.5) (4) Turbellaria 6 2 2 2 1 0 0 0 1 0 0 0 (6.5) (0.5) (1.5) (1.5) (1) (1) Oligochaeta 6 1 0 0 1 1 0 1 0 1 3 1 (0.5) (1) (1) (1.5) (1) (1) Nematoda 11 15 15 0 13 11 2 17 2 4 0 0 (11.5) (5) (9.5) (3) (6) (0.5) (12) (2) (4.5) Hydrachnida 13 2 19 50 22 38 5 19 0 10 8 0 (12.5) (1.5) (9) (2) (16) (24.5) (2) (2.5) (10) (1.5) Culicidae 0 0 0 0 1 0 0 0 0 0 0 0 (1) Ceratopogonidae 1 0 6 12 0 0 1 23 3 4 14 3 (1.5) (3.5) (2.5) (0.5) (20) (1) (2.5) (1) (1) Zygoptera 0 0 0 0 0 0 0 0 0 1 0 0 (1) Trichoptera 0 0 0 0 0 0 0 0 2 0 3 8 (1) Bivalvia 0 0 0 0 0 0 0 0 9 7 17 4 (9) (7) (17) (2.5) Total 300 138 298 347 213 322 98 489 200 221 308 212 (22.5) (28.5) (46.5) (23.5) (101) (94.5) (16.5) (225) (3) (4.5) (5) (9)

183 Appendix 1.B Mean abundance (no. individuals/mesocosm) for benthic invertebrate taxa recorded in stony sediment samples in our mesocosm study. Values are averages (n = 3 rock bags) of the abundance values obtained within each mesocosm, with standard error of the mean in parentheses. C-01 – Control 1, C-02 – Control 2, LD-01 – Low drip 1, LD-02 – Low drip 2, MD-01 – Medium drip 1, MD-02 – Medium drip 2, HD-01 – High drip 1, HD-02 – High drip 2, P-01 – Plug 1, P-02 – Plug 2, Cit-01 – High dose citrate 1, Cit-02 – High dose citrate 2.

C- C- LD- LD- MD- MD- HD- HD- P- P- Cit- Cit- Taxa 01 02 01 02 01 02 01 02 01 02 01 02

Chironomidae 51 26 44.3 11.3 30.3 37.7 30.3 16 23.7 20 45.3 13.7 (8.1) (8) (4.3) (1.9) (3.5) (10.1) (2.6) (1.5) (7.4) (0.6) (6.2) (3.4) Anisoptera 0.7 0 0 0 0 0.7 0.3 0.7 0.7 0 0 0.3 (0.3) (0.3) (0.3) (0.7) (0.3) (0.3) Ephemeroptera 1.3 1.7 0 0.3 5.7 0.7 2.3 0.3 3.7 0 0.3 1 (0.9) (0.9) (0.3) (2.7) (0.7) (0.7) (0.3) (1.2) (0.3) (1) Gastropoda 2 1.3 3 0.7 1 2.3 2.3 1.3 1 0.7 0.7 0.7 (0.6) (0.9) (1.7) (0.3) (0.6) (0.3) (0.9) (0.3) (0.6) (0.3) (0.3) (0.3) Amphipoda 1 3.3 0.3 0 0.7 0 0.7 0.3 0.3 0 1.3 1.3 (0.6) (2.8) (0.3) (0.3) (0.7) (0.3) (0.3) (0.9) (0.7) Turbellaria 2.3 2 2 1.7 2.3 1 3 2.3 4.7 0.3 3 2 (0.7) (0.6) (1.2) (0.9) (0.3) (1.7) (1.5) (3.2) (0.3) (1) (0.6) Oligochaeta 0.3 0.3 0.7 0.3 0.3 0 1 0.3 0.3 0.7 0 0.3 (0.3) (0.3) (0.3) (0.3) (0.3) (0.6) (0.3) (0.3) (0.7) (0.3) Hirudinea 0 0 0 0 0 0 0 0 0 0 0.3 0 (0.3) Coleoptera 0 0 0 0.3 0 0 0.3 0 0 0 0 0.3 (0.3) (0.3) (0.3) Zygoptera 0 0 0 0 0.7 0 0 0 0 0 0.3 0 (0.7) (0.3) Trichoptera 0.3 0.3 0 0.3 1.3 0 0 0 0 0 0.3 0 (0.3) (0.3) (0.3) (0.9) (0.3) Decapoda 0.3 0 0 0 0 0 0 0 0 0 0 0 (0.3) Hydra 0 0.3 0 0 0 0 0 0 0 0 0 0 (0.3) Culicidae 0 0 0 0 0 0 0.3 0 0 0 0.7 0.3 (0.3) (0.3) (0.3) Ceratopogonidae 0 0.7 0 0 0 0 0 0 0.3 0 0 0 (0.3) (0.3) Total 59.3 36 50.3 15 42.3 42.3 40.7 21.3 34.7 21.7 52.3 20 (3.4) (1.7) (2.9) (0.7) (2) (2.5) (2) (1.1) (1.6) (1.3) (3) (0.9)

184

Appendix 2. Benthic macroinvertebrate taxa found in our experimental (Lake 222) and reference lake (Lake 221) across years.

Lake 222 Lake 221

2013 2014 2015 2013 2014 2015

Chironomidae + + + + + + Hirudinea + + + + + + Misc. Diptera + + + + + + Ceratopogonidae + + + + + + Ephemeroptera + + + + + + Anisoptera + + + + + + Trichoptera + + + + + + Amphipoda + + + + + + Hydrachnidae + + + + + + Coelenterata + + + + + Turbellaria + + Nematoda + + + + + Bivalvia + + + + + + Decapoda + + + + + Coleoptera + + + Gastropoda + + + + Tabanidae + + + Hemiptera + Megaloptera + + + + + Plecoptera + + Zygoptera + + +

Total number of taxa 14 20 16 14 18 16

185 Appendix 3. Seasonal (ice-free water) mean values for water chemistry variables in our experimental (Lake 222) and reference lake (Lake 221) across years.

Year DOC TDN TDP Seston P Seston N Seston C TAg pH Chl-a Temperature (mg/L) (µg/L) (µg/L) (µg/L) (µg/L) (mg/L) (µg/L) (°C)

2013 10.49 326.94 6.00 5.41 109.68 0.75 - 8.14 3.38 22.81 222

2014 11.31 454.82 4.83 5.52 99.28 0.82 2.51 7.07 4.61 18.93 Lake 2015 11.10 449.93 7.76 5.53 98.50 0.83 7.56 8.37 4.98 19.16

201320 12.31 350.44 4.09 9.03 200.58 1.14 - 6.49 5.82 23.01 2014 12.66 500.63 4.58 8.67 131.28 1.36 - 6.38 4.99 18.70

Lake 221 Lake 2015 11.86 441.66 7.67 11.26 164.77 1.50 - 6.64 6.42 18.95

186 Appendix 4: Results of generalized linear models for the relationships of measured environmental variables to lake and sampling date (nested within lake). P-values in bold indicate a significant effect (P < 0.05).df represents the degrees of freedom for the sources of variation.

df MS F P

DOC (adj R2 = 0.4272)

Lake 1 10.4159 16.1071 0.001

Lake:Date 2 0.3430 0.5305 0.5983

Residuals 16 0.6467

TDN (adj R2 = 0.0204)

Lake 1 1612.4 0.2676 0.6120

Lake:Date 2 9421.4 1.5636 0.2397

Residuals 16 6025.5 TDP (adj R2 = 0.4743) Lake 1 0.2745 0.1086 0.7466

Lake:Date 2 23.0360 9.1138 0.0029

Residuals 14 2.5276

Seston P (adj R2 = 0.7655)

Lake 1 106.617 45.8982 < 0.0001

Lake:Date 2 22.211 9.5619 0.0019

Residuals 16 2.323

Seston N (adj R2 = 0.2859)

Lake 1 0.0441 8.1845 0.0113

Lake:Date 2 0.007 1.2115 0.3237

Residuals 16 0.005

Seston C (adj R2 = 0.6616)

Lake 1 2.1461 36.0055 < 0.0001

Lake:Date 2 0.1234 2.0704 0.1586

Residuals 16 0.0596

187

Appendix 4 (continued)

df MS F P

Temperature (adj R2 = -0.1771)

Lake 1 0.0225 0.0043 0.9485

Lake:Date 2 1.1369 0.2188 0.8062

Residuals 14 5.1959

Chlorophyll -a (adj R2 = 0.4677)

Lake 1 232.289 6.6659 0.0201

Lake:Date 2 226.960 6.5130 0.0085

Residuals 16 34.847 pH (adj R2 = 0.5416)

Lake 1 8.0096 23.8480 0.0001

Lake:Date 2 0.2693 0.8019 0.4657

Residuals 16 0.3359

188 Appendix 5. Littoral microcrustacean species recorded at sampling sites in our experimental (Lake 222) and reference lake (Lake 221) across years.

Lake 222 Lake 221

Taxa 2013 2014 2015 2013 2014 2015

Cladocera Bosminidae Bosmina longirostris + + + + + +

Chydoridae Acroperus cf. harpae + + + + + Alona intermedia + + + + + Alona cf. rustica + + + + + + Alona quadrangularis/affinis + + + + + + Alonella nana + + + + + + Chydorus cf. brevilabris + + + + + + Chydorus piger + + + + + + Disparalona leei + + + + + + Pleuroxus procurvus +

Daphniidae Ceriodaphnia spp. + + + + + +

Macrothricidae Acantholeberis curvirostris + + Ilyocryptus sordidus + + + + + + Ophryoxus gracilis + + + + + + Streblocerus serricaudatus + + + +

Polyphemidae Polyphemus pediculus + + + +

Sididae Sida crystallina + + + + + + Latona setifera + + + + +

189 Appendix 5 (continued)

Lake 222 Lake 221

Taxa 2013 2014 2015 2013 2014 2015

Copepoda Calanoida Copepodites + + + + + Leptodiaptomus spp. + +

Cyclopoida Copepodites + + + + + + Acanthocyclops spp. + + + + + + Eucyclops spp. + + + + + + Macrocyclops spp. + + + + + + Microcyclops spp. + + + + + + Tropocyclops spp. + + + + + +

Harpacticoida Copepodites + + +

Total number of species 22 24 22 22 24 24

190 Appendix 6. Pelagic zooplankton species recorded in our experimental (Lake 222) and reference lake (Lake 221) across years.

Lake 222 Lake 221

Taxa 2013 2014 2015 2013 2014 2015

Cladocera Bosminidae Bosmina longirostris + + + + + +

Chydoridae Acroperus cf. harpae + Alona affinis + Alona intermedia + +

Daphniidae Ceriodaphnia spp. + + + + Daphnia galeata mendotae + + Daphnia pulex + + + + + +

Holopediidae Holopedium gibberum + + + + + +

Sididae Diaphanosoma birgei + + + + + +

Leptodora kindtii +

Leptodoridae Cyclopoida Copepodites + + + + + + Nauplii + + + + + + Acanthocyclops spp. + + + Diacyclops spp. + + + + Eucyclops spp. + + Macrocyclops spp. + Mesocyclops spp. + + + + + + Orthocyclops spp. + + + + + + Tropocyclops spp. + + + + + +

191 Appendix 6 (continued)

Lake 222 Lake 221

Taxa 2013 2014 2015 2013 2014 2015

Copepoda Calanoida Copepodites + + + + + + Nauplii + + + + + + Epischura lacustris + + + + + Leptodiaptomus spp. + + + + +

Diptera Chaoborus spp. + + + + +

Total number of species 17 16 16 17 16 20

192