Stewart Emily M 201801 Phd.Pdf

THE ROLE OF CHIRONOMIDS AS PALEOECOLOGICAL INDICATORS OF EUTROPHICATION IN SHALLOW LAKES ACROSS A BROAD LATITUDINAL GRADIENT

Emily Meaghan Stewart

A thesis submitted to the Department of Biology

In conformity with the requirements for

the degree of Doctor of Philosophy

Queen’s University

Kingston, Ontario, Canada

(January, 2018)

The aquatic larvae of chironomids (Diptera, Chironomidae) were historically classified according to lake trophic status, and taxa classified as “eutrophic” were labeled as such because of adaptations for surviving hypoxic or anoxic conditions in the hypolimnion of stratified eutrophic lakes. As such, sedimentary chironomid assemblages have been used to reconstruct production-related variables

(nutrients, chlorophyll-a), though this has been problematic, especially in shallow systems, because the response of chironomids to eutrophication is mediated through secondary environmental gradients including oxygen concentration, habitat, and food quality/quantity. In this thesis, eutrophic sewage ponds in the Canadian High Arctic were used to demonstrate that oxygen, not nutrients, is the primary control of chironomid species assemblages. The ability to explicitly test the influence of oxygen versus nutrients on chironomid distributions was made possible by the 24-hr daylight (continuous photosynthesis) and shallow, wind-mixed water column that resulted in oxygen concentrations that were decoupled from the effects of elevated nutrients and production. The subfossil chironomid assemblages were complacent during historical eutrophication, in contrast to marked changes in diatom assemblages, which have a direct physiological relationship with nutrients. Similarly, in shallow eutrophic ponds on islands in Lake

Ontario, chironomid assemblages did not appear to be governed by the large gradient in total phosphorus due to the presence or absence of waterbird nesting colonies, but rather by habitat and possible bird- mediated heavy metal pollution. In a subarctic lake that was formerly used for sewage disposal, chironomid assemblages were relatively unresponsive to eutrophication in comparison with the larger turnover in diatom species. However, periods of low oxygen observed in the temperate and subarctic sites may explain the higher (but still low) relative abundances of hypoxia-tolerant Chironomus species at these sites compared to the High Arctic ponds. Together, this research demonstrates the problems associated with classifying chironomids based on nutrient levels. Since it is uncommon to examine chironomid responses to eutrophication across latitude, this thesis offers a relatively unstudied perspective

ii of chironomid ecology, emphasizing that some of the assumptions of temperate chironomid ecology (with regards to eutrophication) may not necessarily hold true when applied at higher latitudes.

iii

Co-Authorship

Chapter 2 was co-authored by Reba McIver, Neal Michelutti, Marianne S.V. Douglas, and John P. Smol.

I collected 2011 samples, co-mentored Reba McIver with Neal Michelutti (supervised by John Smol), analysed chironomids, completed data analysis (including statistical work), and was the primary author of the paper. Reba McIver identified diatoms under the tutelage of Neal Michelutti. Marianne Douglas led

2011 field work, and together with John Smol, oversaw study design. Previous field seasons were also led by Marianne Douglas and John Smol. This chapter was published separately:

Stewart EM, McIver R., Michelutti N, Douglas MSV, Smol JP (2014) Assessing the efficacy of chironomid and

diatom assemblages in tracking eutrophication in High Arctic sewage ponds. Hydrobiologia 721:251–268

Chapter 3 was co-authored by Neal Michelutti, Christopher Grooms, Linda E. Kimpe, Jules M. Blais, and John P. Smol. I participated in study design and field work; analysed chironomids, sedimentary chlorophyll-a, and water chemistry; performed statistical analyses; and was the primary author of the chapter. Neal Michelutti aided with data analysis and interpretation, as well as (with Christopher Grooms) completed field work. Linda Kimpe and Jules Blais completed elemental analysis on the cores, and aided with the interpretation of core radiochronologies. John Smol oversaw study design and data interpretation.

Chapter 4 was co-authored by Kathryn E. Hargan, Branaavan Sivarajah, Linda E. Kimpe, Jules M. Blais, and John P. Smol. I participated in study design and field work, processed for chironomids and chlorophyll-a, did statistical analyses, and was the primary author of the paper. Kathryn Hargan completed sterol analysis and interpretation under the supervision of Jules Blais. Branaavan Sivarajah completed diatom analysis. Jules Blais led field work and aided in sterol analysis. Linda Kimpe was responsible for core radiochronology. John Smol oversaw project design. The manuscript for this chapter is under review in the journal Arctic (submission number: 17-134).

Appendix A was co-authored by Kathryn E. Hargan, Neal Michelutti, Christopher Grooms, Linda E.

Kimpe, Mark L. Mallory, John P. Smol, and Jules M. Blais. I helped with study design, collected and analysed 210Pb data, analysed total lead and stable lead isotope data, and was the second author of the iv paper. Kathryn Hargan completed sterol analysis and was first author of the manuscript. Neal Michelutti and Christopher Grooms did field work. Linda Kimpe and Jules Blais oversaw sterol data analysis and interpretation, as well as collected total lead and stable lead isotope data. Mark Mallory helped with data analysis with respect to birds. John Smol aided in project concept design. Data were originally included as supplementary material in:

Hargan KE, Stewart EM, Michelutti N, Grooms C, Kimpe L, Mallory M, Smol JP, Blais JM (in review)

Incorporating sterols and stanols as biomarkers for tracking waterbird impacts to temperate ponds. Proc R

Soc Lond B Biol Sci: manuscript ID RSPB-2017-2669

Appendix B was co-authored by Neal Michelutti, Mina Vu, Christopher Grooms, Linda E. Kimpe, John

P. Smol, and Jules M. Blais. I took part in study design, co-mentored Mina Vu with Neal Michelutti

(supervised by John Smol), analysed data, performed statistical analyses, and was the primary author of the manuscript (in prep.). Mina Vu analysed diatoms with Neal Michelutti. Christopher Grooms did field work and collected chlorophyll-a data. Linda Kimpe and Jules Blais performed stable isotope analysis and helped with interpretation. John Smol partook in study design and overseeing data analysis and interpretation.

Acknowledgements

My sincerest thanks to my supervisor, John Smol, and co-supervisor, Marianne Douglas, without whom I would not be a paleolimnologist. John, thank you for taking me in as a 3rd year summer student and letting me stay for a defining 7 years of my life. You both fostered my love of limnology and introduced me to one of my favourite things: The Arctic. Thank you for this opportunity, and thank you for sharing your infectious passion for learning and exploring with your students. I could not have had a better start to my path as a scientist. Thank you to my committee members, Drs. Shelley Arnott and Scott

Lamoureux, for positive committee meeting experiences that churned out better ideas and led to stronger work.

Thank you to PEARL, current and former, you have always been a family to me. PEARL was and is full of my best friends - a support network and a collection of wonderful minds that I have the privilege to know. I would like to express an immense amount of gratitude to Neal Michelutti – you were stuck with me from the beginning and dealt with the terrible writing and numerous questions! Thank you for being a superior role model in science writing and interpretation; your patience and constructive feedback have always been helpful. I cannot name everyone who has made my time here memorable, but needless to say, if you were or are a part of PEARL, you are on that list! Finally, thank you to the Secchi table for consistently hosting some of the very best and funniest experiences I’ve ever had, as well as the most productive science conversations – especially during coffee party.

Thank you to my family, especially my poor parents and sister, who withstood a large portion of my graduate woes without (much) complaint. You are more important than I could ever say, and you deserve at least one of the letters of this PhD (the “P” perhaps). An enormous thank-you is awarded to my husband, Graham, you obviously deserve the other letters. Not only did you come into my life at the beginning of my Master’s, but a) you decided to stay with me through the PhD, despite the stress and craziness, b) you are absolutely my favourite person, and c) you gave me an amazing and supportive second family, as well. The love and support of my family through this was and still is immeasurable. vi

Table of Contents

Abstract ...... ii Co-Authorship...... iv Acknowledgements ...... vi List of Figures ...... xi List of Tables ...... xiii List of Supplemental Figures ...... xiv List of Supplemental Tables ...... xv List of Abbreviations ...... xvi Chapter 1 Introduction and literature review ...... 1 Eutrophication ...... 1 Paleolimnology ...... 2 Chironomids as proxies for eutrophication ...... 4 Thesis objectives ...... 8 References ...... 9 Chapter 2 Assessing the efficacy of chironomid and diatom assemblages in tracking eutrophication in High Arctic sewage ponds ...... 20 Abstract ...... 20 Introduction ...... 21 Materials and methods ...... 24 Site description ...... 24 Water chemistry ...... 25 Epilithic diatom sampling ...... 26 Surface sediment and sediment core sampling ...... 26 Diatom processing and assessment...... 27 Chironomid head capsule collection and assessment ...... 27 Statistical analysis ...... 27 Results ...... 28 Water chemistry ...... 28 Epilithic diatom rock scrapes ...... 29 Sediment core diatom assemblages ...... 29 Surface sediment chironomid assemblages ...... 30 Sediment core chironomid assemblages ...... 30

vii

Discussion ...... 31 Limnological responses to changing sewage inputs ...... 31 Diatom response to eutrophication and recovery from the sediment cores ...... 33 Epilithic diatom response to continuing recovery ...... 34 Chironomid response to eutrophication and recovery ...... 35 Evidence of recent climate warming ...... 38 Conclusions ...... 40 References ...... 41 Chapter 3 Bird-mediated eutrophication and enrichment of heavy metals in shallow ponds on islands in Lake Ontario ...... 60 Abstract ...... 60 Introduction ...... 61 Site description...... 63 Bird-impacted ponds ...... 63 Non-impacted ponds ...... 64 Methods ...... 65 Water chemistry ...... 65 Sediment sampling and analysis ...... 66 Statistical analysis ...... 67 Results ...... 67 Water chemistry ...... 67 Trace element concentrations in water ...... 68 Sedimentary trace element chemistry ...... 70 Dissolved oxygen concentrations ...... 70 East Brother sediment core ...... 71 Little Galloo sediment core ...... 72 Pigeon sediment core ...... 72 Main Duck Pond 2 sediment core ...... 73 Calf sediment core ...... 73 Comparison of bird-impacted and non-impacted sediments ...... 74 Discussion ...... 75 Waterbird-mediated impacts on water chemistry ...... 75 Chironomid assemblage response to eutrophication and contamination ...... 78 Conclusions ...... 82 viii

References ...... 84 Chapter 4 A paleoenvironmental study tracking eutrophication, metal pollution, and climate change in Niven Lake (NT), Yellowknife’s first sewage lagoon ...... 105 Abstract ...... 105 Introduction ...... 106 Site description...... 109 Methods ...... 110 Water chemistry ...... 110 Sediment sampling and dating ...... 111 Sterols and stanols ...... 111 Elemental analysis and stable nitrogen isotopes ...... 112 Spectrally-inferred chlorophyll a ...... 113 Diatoms and chironomids ...... 113 Results ...... 114 Water chemistry ...... 114 Sediment chronology ...... 116 Sterols and stable nitrogen isotopes ...... 117 Diatoms and chironomids ...... 118 Discussion ...... 120 Tracking sewage using sterols and δ15N ...... 120 Effects of sewage, mining, and climate warming on Niven Lake ...... 122 Recovery in Niven Lake ...... 126 Conclusions ...... 127 References ...... 129 Chapter 5 General Discussion ...... 148 Chironomids as indicators of eutrophication in shallow freshwaters ...... 148 General patterns in shallow lake and pond ecology ...... 154 Effects of metal pollution on chironomid assemblages ...... 155 Summary ...... 157 References ...... 157 Appendix A Lake Ontario bird ponds radiometric dating profiles ...... 162 Appendix B Diatom data for Lake Ontario island ponds ...... 166 Appendix C Chapter 2 (Resolute Bay) raw counts ...... 172 Appendix D Chapter 3 (Lake Ontario bird pond) raw counts ...... 181 ix

Appendix E Chapter 4 (Niven Lake) raw counts ...... 189

List of Figures

Figure 1.1 The dynamics of eutrophication in deep stratifying lakes and shallow lakes ...... 18 Figure 1.2 Chironomid assemblages from bird-impacted ponds and a non-impacted pond at Cape Vera, on Devon Island (NU) ...... 19 Figure 2.1 Map showing the location of Resolute Bay on Cornwallis Island, with an inset of sewage and control ponds ...... 47 Figure 2.2 Photographs of the Resolute sewage ponds and reference sites ...... 48 Figure 2.3 Mean annual temperature for Resolute Bay, Nunavut, from 1948 to 2010 ...... 49 Figure 2.4 Limnological variables over time for the Resolute control ponds ...... 50 Figure 2.5 Limnological variables over time for the Resolute sewage ponds ...... 51 Figure 2.6 The epilithic diatom assemblages for all Resolute sewage ponds ...... 52 Figure 2.7 The diatom species assemblage changes of the the sediment cores for the sewage ponds (R-12 and R-13) and the control ponds (R-1 and R-2) ...... 53 Figure 2.8 The chironomid assemblage changes of the sewage pond cores (R-12 and R-13) and the control pond cores (R-2 and R-1) ...... 54 Figure 2.9 Principal Component Analysis of sedimentary chironomid assemblages and detrended correspondence analysis of sedimentary diatom assemblages in the Resolute ponds ...... 55 Figure 3.1 Bird-impacted and non-impacted study ponds on islands in eastern Lake Ontario ...... 91 Figure 3.2 Select water chemistry data from bird-impacted sites and non-impacted sites on islands in eastern Lake Ontario ...... 93 Figure 3.3 Total concentrations of trace elements in water from bird-impacted ponds and non-impacted sites on islands in eastern Lake Ontario...... 94 Figure 3.4 Dissolved oxygen concentrations and saturation in East Brother Pond, Main Duck Pond 2, and Calf Pond ...... 95 Figure 3.5 Chironomid assemblages of bird-impacted sites, East Brother Pond and Little Galloo Pond, as well as δ15N and VRS-chla profiles ...... 96 Figure 3.6 Chironomid assemblages of non-bird-impacted sites, Main Duck Pond 2 and Calf Pond, as well as δ15N and VRS-chla profiles ...... 97 Figure 3.7 Principal Component Analysis (PCA) of down-core chironomid species data for bird-impacted ponds, East Brother and Little Galloo, as well as non-impacted ponds, Main Duck Pond 2 and Calf ...... 98 Figure 4.1 Map showing the location of Niven Lake in Yellowknife (NT) ...... 141

Figure 4.2 Relative abundances (%) of sterols and stanols, as well as stable nitrogen isotopes (δ15N in ‰) in the sediment core from Niven Lake ...... 144 Figure 4.3 The sub-fossil diatom and chironomid assemblages of Niven Lake ...... 145

xii

List of Tables

Table 1.1 Possible negative consequences of freshwater eutrophication (Schindler and Smith 2009; Smith 2003) ...... 17 Table 3.1 Pond characteristics and water chemistry of bird-impacted sites and non-impacted sites on islands in eastern Lake Ontario ...... 92 Table 4.1 Sterol fecal-contamination ratios used to compare with sediments of Niven Lake ...... 140 Table 4.2 Modern water chemistry of Niven Lake and Frame Lake, as well as historical sampling data from 1990 ...... 142 Table 4.3 Dissolved oxygen concentration and saturation, specific conductance, and water temperature over depth in Niven Lake...... 143

xiii

List of Supplemental Figures

Supplemental Figure 2.1 210Pb, 214Bi, and 137Cs radioactivities in the Resolute sediment cores...... 56 Supplemental Figure 2.2 Dissolved oxygen concentration (mg/L) and saturation (%) for sewage ponds, R-12 and R-13...... 59 Supplemental Figure 3.1 Trace metal and element concentrations in surface sediments (0-0.5 cm) of cores from High Bluff (cormorant-impacted), Gull (gull-impacted), Little Galloo (mixed species impact), and Calf (non-impacted) ponds ...... 102 Supplemental Figure 3.2 The sedimentary chironomid remains retrieved from Pigeon Pond ...... 103 Supplemental Figure 3.3 Principal Component Analysis (PCA) of chironomid species data of sediment cores from two sewage ponds (R-12 and R-13) and a reference pond (R-2) near Resolute Bay, Cornwallis Island (NU). The temporal dissolved oxygen concentration (mg/L) and saturation (%) profiles of B. R-12 and C. R-13 ...... 104 Supplemental Figure 4.1 210Pb gamma spectrometry dating for the Niven Lake (NT) sediment core. .. 146 Supplemental Figure 4.2 Mean air temperatures for the City of Yellowknife (NT) ...... 147

xiv

List of Supplemental Tables

Supplemental Table 2.1 Chironomid species presence or absence in the surface sediments of the sewage ponds near Resolute Bay, Cornwallis Island, Nunavut...... 57 Supplemental Table 2.2 Species abbreviations used in PCA and DCA of Resolute sewage and control ponds ...... 58 Supplemental Table 3.1 Raw data for select water chemistry of bird-impacted and non-impacted ponds on islands in eastern Lake Ontario ...... 99 Supplemental Table 3.2 Raw data for trace element concentrations of bird-impacted and non-impacted ponds on islands in eastern Lake Ontario ...... 100 Supplemental Table 3.3 Trace element concentrations not presented in Figure 3.3 of bird-impacted and non-impacted ponds on islands in eastern Lake Ontario ...... 101

List of Abbreviations

BOD Biological oxygen demand CCME Canadian Council of Ministers of the Environment chl-a Chlorophyll a (water concentration, unless specified – see VRS-chla) Cond Specific conductance CONISS Constrained incremental sum of squares CRS Constant Rate of Supply DCA Detrended correspondance analysis DIC Dissolved inorganic carbon DOC Dissolved organic carbon HC/g dry sed Head capsules per gram dry sediment ISQG Interim sediment quality guideline ITR Intrinsic time resolution MAAT Mean annual air temperature PCA Principal Component Analysis PEL Probable effects level POC Particulate organic carbon PON Particulate organic nitrogen TN Total nitrogen TN-f Total filtered nitrogen TN-u Total unfiltered nitrogen TP Total phosphorus TP-f Total filtered phosphorus TP-u Total unfiltered phosphorus VRS-chla Visual reflectance spectroscopy chlorophyll a (sediment concentration)

Zmax Maximum depth δ15N stable nitrogen isotope ratio

xvi

Chapter 1

Introduction and literature review

Eutrophication

Cultural eutrophication remains the main threat to freshwater ecosystems worldwide. It results from over-fertilization of nutrients to waterbodies, resulting in excessive algal growth and a multitude of other problematic symptoms, summarized in Table 1.1 (Smith 2003; Smith and Schindler 2009).

Phosphorus inputs, largely from agricultural runoff and municipal sewage, represent the predominant driver of eutrophication in lakes and rivers (Schindler 1977). Long-term reductions of phosphorus loading have typically curbed freshwater eutrophication in Europe and North America (Schindler et al. 2016); however, recent anthropogenic warming in combination with nutrient inputs, has resulted in additional limnological effects (e.g. a switch to cyanobacterial populations), further complicating eutrophication management plans (Smol 2010).

Eutrophication has been well-studied over the past six decades, particularly in deep temperate lakes (Smith et al. 2006). One of the most apparent environmental changes associated with eutrophication is the increase in primary production, by either microscopic algae, macrophytes, or both. Large algal blooms can increase the biological oxygen demand in hypolimnetic waters (Figure 1.1), leading to undesirable side effects such as fish kills and reductions in biodiversity (Seehausen et al. 1997; Jacobson et al. 2010; 2017). However, shallow, polymictic lakes often follow different environmental trajectories relative to deep, stratified lakes (Bennion et al. 2010). In particular, the dynamics of large nutrient additions to shallow lakes is largely governed by macrophyte abundance, which can deter phytoplankton

(e.g. cyanobacteria) blooms by taking up nutrients, stabilizing sediments against resuspension, and providing habitat for grazers that help control blooms (Scheffer et al. 1993). However, once a threshold of nutrient additions is reached, a shallow system may rapidly shift from a clear water, macrophyte- dominated state to a turbid state dominated by planktonic algal blooms (Table 1.1; Figure 1.1), or in other

words, the lake undergoes an ecological regime shift (Scheffer and Jeppesen 2007). Furthermore, oxygen dynamics associated with eutrophication in shallow lakes depend on mixing and photosynthesis, and oxygen depletion may not occur, or may be temporary or localized, as these systems generally do not thermally stratify for sustained periods of time.

This thesis focuses on long-term changes linked to eutrophication in shallow lakes and small ponds. Specific causes of eutrophication may be known, as in the case of historical sewage disposal to freshwater systems (e.g. Chapters 2 and 3), yet the pre-disturbance conditions in a lake may not have been adequately recorded, making interpretations of the trajectory of eutrophication difficult. When monitoring records are absent, paleoecological proxy methods can be used to infer past environmental conditions, placing current or future eutrophication-related changes in meaningful context. In this thesis, I use paleolimnological methods to reconstruct past environmental conditions in various lake settings, tracking the magnitude and ecological consequences of eutrophication over time.

Paleolimnology

Paleolimnology uses the biological, chemical, and physical information preserved in lake sediments to infer past environmental changes (Smol 2008). One of the first applied uses of paleolimnology was in eutrophication research, specifically putting recent limnological conditions into the longer context of natural variability (e.g. Edmonson 1974). Recent lake sediments (< 150 years old) can be dated using 210Pb (Appleby 2001), and dated sediment records can then be assessed for a variety of paleolimnological proxies, including biological remains (e.g. subfossil algae, invertebrates), biogeochemical signatures (e.g. stable isotopes), and physical parameters (e.g. particle size) (reviewed in

Smol 2008). In this thesis, biological proxies, in conjunction with geochemical and biochemical indicators, were used to investigate various scenarios of eutrophication, with particular emphasis on determining how interacting environmental parameters affect biological assemblages during and after shallow lake eutrophication.

Paleolimnological reconstructions have been instrumental in determining the nature, timing, and magnitude of lake eutrophication scenarios. Even though phosphorus from lake water is stored in sediments, using direct measures of sedimentary phosphorus concentration or flux to track trends in nutrient inputs is unreliable because of the mobile and labile nature of phosphorus in sediments after it is deposited (Ginn et al. 2012). Nor does this approach provide any information on the ecological consequences associated with excess nutrients. Increases in overall primary production that often occur with eutrophication can be tracked using sedimentary concentrations of chlorophyll a (and its primary degradation products), as inferred using visual spectroscopy (Wolfe et al. 2006; reviewed by Michelutti and Smol 2016). Furthermore, specific causes of eutrophication can be tracked with biochemical markers, such as sedimentary sterols and stanols (Leeming et al. 1996), which are often used to track human sewage inputs to lakes via increases in coprostanol (reviewed by Korosi et al. 2015). Stable nitrogen isotope analysis (δ15N) has been commonly paired with the use of sterols and stanols (Vane et al. 2010), since the heavier stable isotope of nitrogen (15N) in enriched up food chains and can track nutrients sourced from higher trophic levels. Thus, not only has δ15N been used to track human sewage inputs, but also the nutrient-rich waste inputs from gregarious animals such as migrating salmon (Finney et al. 2000) or colonial seabirds (Blais et al. 2005).

Although more labour intensive, the remains of biological organisms that respond directly or indirectly to the effects of nutrient inputs can be used to infer past eutrophication, which can also provide information about the biological changes associated with nutrient enrichment. Reconstructions using biological proxies rely on the relationships of freshwater biota to nutrients (e.g. phosphorus), which can be determined by relating modern water chemistry to modern spatial species distributions in order to determine individual environmental optima and tolerance ranges using statistical methods such as weighted-averaging (reviewed by Juggins and Birks 2012). However, for these paleolimnological reconstructions to be reliable, the relationship between the indicator organism and the inferred variable

(e.g. phosphorus) must not change over time or in confluence with other environmental factors (e.g.

temperature, water depth) (Birks 1995). Unfortunately, this is rarely the case, since biological indicators have complex ecologies affected by multiple interacting environmental factors. Therefore, studying the relative relationships of paleo-indicators to various environmental conditions is essential for reliable paleoecological reconstructions.

Diatoms (Bacillariophyta) are an example of a commonly used biological proxy for investigations of eutrophication, including the quantitative reconstruction of total phosphorus concentrations (Hall and

Smol 2010). Diatom species distributions can often be characterized based on their optima for lake nutrient concentrations since they respond directly to changes in, for example, phosphorus. Here, diatoms are used as a proxy for determining lake eutrophication in contrast to the main focus of this thesis, chironomids (Diptera, Chironomidae), which have also been used to reconstruct past eutrophication, though the response of chironomids to nutrients is indirect. Thus, the relationship of chironomid species with lakewater nutrients is complicated by various other environmental factors, including climate and oxygen concentrations, which can be difficult to predict in shallow systems.

Chironomids as proxies for eutrophication

Chironomids (non-biting midges) are ubiquitous insects with an aquatic larval phase that represents the majority of their life cycle. The chitinous head capsules of larvae are shed between instar moults and are typically well-preserved in sediments. Chironomid larvae are sensitive to limnological change making them useful paleoindicators of multiple environmental variables (Walker 2001), and sedimentary head capsule remains can be separated taxonomically based on morphology (Larocque and

Rolland 2006; Brooks et al. 2007; Andersen et al. 2013). For example, chironomids have species-specific optima for water temperature (Eggermont and Heiri 2012), with distinct latitudinal (Walker et al. 1997;

Barley et al. 2006) and altitudinal (Walker and Mathewes 1989) distributions in freshwater lakes. Thus, subfossil chironomid remains have been commonly used to reconstruct broad climatic changes in both

Europe (Velle et al. 2005; Heiri et al. 2011) and North America (Porinchu et al. 2007; Rolland et al.

2008).

Chironomid sub-fossil remains have also been used to reconstruct aspects of eutrophication, as chironomid larvae were historically categorized according to lake trophic status and oxygen content

(Thienemann 1920; 1954; Brundin 1956; Sæther 1979). For example, a characteristic difference between oligotrophic lakes and eutrophic lakes is the relative importance of Tanytarsus species compared to

Chironomus species, as the latter may dominate eutrophic lakes (Thienemann 1922; Hofmann 1988). The reconstruction of eutrophication using chironomid assemblages is predominantly based on the hypolimnetic oxygen depletion that accompanies nutrient inputs and productivity increases in stratified lakes, facilitated by the decomposition of algal blooms at the lake bottom and creating a niche for certain taxa (e.g. Chironomus) well-adapted to survive hypoxic or anoxic conditions in the profundal zone.

Species-specific survival of hypoxia and anoxia is aided by production of high concentrations of a haemoglobin-like molecule (Czeczuga 1960; Weber 1980), the ability to switch to anaerobic metabolism

(Hamburger et al. 1995), efficient osmotic and ionic regulation under anaerobia (Scholz and Zerbst-

Boroffka 1998), the ventilation of tube dwellings (Int Panis et al. 1995), and large body size (Heinis et al.

1994). Species differences caused by changes in hypolimnetic oxygen have been well-documented and often involve increases in oxy-regulator species that tolerate hypoxia (e.g. Chironomus, Procladius) and decreases in “oxy-stressors” that increase respiration in hypoxic waters up until a critical point (e.g.

Micropsectra) (Brodersen et al. 2008).

Fossil chironomid assemblages in sediment records have been used for quantitative reconstructions of hypolimnetic oxygen (Quinlan et al. 1998; Quinlan and Smol 2001; Little and Smol

2001; Luoto and Salonen 2010), chlorophyll-a concentrations (chl-a) (Brodersen and Lindegaard 1999;

Brodersen et al. 2001), and total phosphorus (TP) (Brooks et al. 2001; Zhang et al. 2006). However, the complexity of the underlying mechanisms behind chironomid responses to TP and chl-a have been demonstrated by Langdon et al. (2006), who showed that lake type, and especially lake depth, is important in determining how chironomids respond to eutrophication, as in general, chironomid-based models over-predicted TP in stratified lakes experiencing hypolimnetic oxygen depletions and under-

predicted TP in shallow eutrophic lakes with elevated bottom-water oxygen. Similarly, Clerk et al. (2000) showed that chironomid-based reconstructions of TP in a south-central Ontario lake consistently underestimated its trophic status due to high dissolved oxygen levels in the large hypolimnion.

Furthermore, Little et al. (2000) found that species shifts following eutrophication and oxygen depletion in Gravenhurst Bay (Ontario, Canada) did not track phosphorus reductions, because hypolimnetic oxygen remained depleted. Thus, the case can be made that hypolimnetic oxygen is a stronger predictor of profundal chironomid distribution than TP or primary production in stratifying eutrophic lakes.

The response of chironomids to eutrophication in shallow lakes is usually due to mechanisms other than oxygen depletion, since typically the chironomid assemblage is entirely littoral. Littoral chironomid assemblages may be more diverse because of the heterogeneity of littoral habitats, and changes that occur with eutrophication may therefore be less predictable, especially since oxygen levels are generally consistently high due to wind mixing, and hypoxia may occur only rarely (e.g. in the winter under ice) (Brodersen and Quinlan 2006). Furthermore, littoral taxa (e.g. Cricotopus) are often “oxy- conformers” that cannot maintain the uptake of oxygen when availability is low (Brodersen et al. 2008), and thus it is generally assumed that hypoxia tolerance is not an important adaptive consideration for littoral species. Factors that do affect chironomid assemblages in shallow lakes during eutrophication may include changes in food quality, sedimentary organic content, habitat structure, and trophic interactions

(Brodersen and Lindegaard 1997). For example, it has been shown that during eutrophication in a Danish lake, shifts in the sedimentary chironomid assemblage closely reflected the succession of various types of macrophytes, as well as the eventual switch to a phytoplankton-dominated state (Brodersen et al. 2001).

In general, the response of the littoral assemblage of a shallow lake to eutrophication includes decreases in the relative abundance of Ablabesmyia, Cladopelma, Paratanytarsus, Pseudochironomus, and

Psectrocladius, with concurrent increases in, Chironomus plumosus-type, Cricotopus, Glyptotendipes,

Microchironomus, and Procladius (Brodersen and Quinlan 2006).

As chironomids have strong latitudinal distributions (Walker et al. 1997; Barley et al. 2006), the response to eutrophication may vary depending on climate. Continued investigation into the response of chironomids to eutrophication in shallow sites is warranted due to the complex play of multiple interacting environmental drivers that may be more difficult to predict than oxygen dynamics in deep eutrophic lakes, and the diversity of this response across latitude is of additional interest. An example of the dissimilarity of chironomid assemblage responses under extreme environmental conditions is given by

Stewart et al. (2013), who showed that oxygen concentrations may become decoupled from nutrient changes during seabird-mediated eutrophication in shallow and well-mixed sites in the High Arctic, inhibiting chironomid species shifts that might be expected in temperate latitudes. Ponds were shallow and well-mixed, experiencing 24 hours of daylight in the growing season and thus allowing photosynthesis to continuously add oxygen to the pond (Blais et al. 2005 supplemental material). Under these eutrophic conditions, chironomid assemblages showed little notable compositional differences in comparison with oligotrophic reference sites (Michelutti et al. 2011), nor any species shifts concurrent with changes in nutrients (Stewart et al. 2013; Figure 1.2). The only major indication of eutrophication or enhanced primary production in the chironomid record was an increase in overall head capsule abundance, likely related to enhanced food availability (Stewart et al. 2013; Figure 1.2). Furthermore, the species present in these eutrophic High Arctic ponds are different from those listed as typical for shallow eutrophic lakes (Brodersen and Quinlan 2006), providing an impetus for further research, especially at higher latitudes.

In this thesis, I use shallow lakes that have undergone eutrophication across a large gradient of latitude to investigate the interactions between nutrient concentrations and oxygen levels, as well as the overall influence of climate, in order to determine the relative influences of these factors on chironomid distributions and/or species shifts over time. Such information should further our ability to reconstruct eutrophication in shallow lakes using the sediment record. Furthermore, as water quality issues rarely

occur in isolation, other anthropogenic stressors, including the overarching effect of recent climate warming and various sources of contaminant pollution (e.g. metals) must also be considered.

Thesis objectives

The overall objective of this thesis was to investigate the various roles of nutrients (particularly phosphorus), dissolved oxygen, and climate on chironomid assemblages in shallow freshwater ecosystems. Ponds and shallow lakes across a gradient of trophic states and latitude were investigated for chironomid assemblage changes over the past ~150 years at several sites where eutrophication has occurred. Each chapter in this thesis presents a diverse scenario of eutrophication, using the contrast between diatoms and chironomids to highlight the nature of chironomid responses to eutrophication.

Below the specific objectives of each chapter are outlined.

In Chapter 2, I compare the responses of diatoms and chironomids to nutrient inputs in shallow

High Arctic ponds that received raw sewage for several decades, and thus became eutrophic in an otherwise unproductive polar desert (Douglas and Smol 2000). I aimed to track the onset of eutrophication and subsequent recovery using sediment cores, as well as surface sediment samples and diatom rock scrape samples taken in various sampling seasons since 1993. Here, the main objective was to validate the High Arctic investigation of chironomid remains at Cape Vera, Nunavut (Stewart et al.

2013) using other High Arctic ponds that underwent a different mode of eutrophication, namely via human sewage inputs rather than guano inputs from a large bird colony. As the colony of seabirds at Cape

Vera was longstanding – representing the entire duration of the core – no inferences regarding the response of chironomids to the onset of nutrient inputs could be made. In the Resolute sewage ponds, the onset of sewage inputs was well-documented, and pre-sewage conditions in the cores could be assessed, thereby determining if chironomids responded to nutrient inputs, or if oxygen and nutrient levels decoupled by 24-hour daylight inhibited a species shift typically characteristic of eutrophication.

In Chapter 3, I document the response of chironomids in shallow temperate ponds that have undergone eutrophication in the past on islands in eastern Lake Ontario that host colonies of waterbirds,

which in turn, fertilize the ponds with guano and other bodily waste (Stewart et al. 2015; Hargan et al. in review). As the Cape Vera (Michelutti et al. 2011; Stewart et al. 2013) and Resolute Bay (Chapter 2) chironomid investigations were situated in the unique environment of the High Arctic, the primary aim of

Chapter 3 was to conduct a similar study in a temperate region that does not experience 24-hours of daylight, and thus, as a result, may also have different oxygen dynamics. Eutrophic bird-impacted ponds were compared to naturally eutrophic reference sites (as opposed to oligotrophic reference sites in the

High Arctic studies), and chironomid assemblages were compared with measured week-long (or longer) profiles of hourly dissolved oxygen, as well as water chemistry. This investigation, as it relates to this thesis, will focus on determining the ecology of the chironomids in the temperate shallow ponds, with an additional emphasis on the contrast with the High Arctic sites.

In Chapter 4, I assess past environmental changes in a shallow lake (Niven Lake) in the City of

Yellowknife (NT) that experienced eutrophication during its former use as a sewage lagoon. Using multiple paleolimnological proxies, the pre-impact conditions of Niven Lake were assessed and compared with the trajectory of biological and chemical change during the sewage era, as well as potential recovery afterward. In this context, the response of chironomids to eutrophication in a shallow subarctic lake was assessed alongside oxygen measurements. The subarctic latitude of this lake offers a perspective of chironomid ecology mid-way between Chapters 2 and 3, allowing some insight into the dynamics of chironomid ecology in shallow eutrophic sites across a large gradient of latitude.

References

Andersen T, Cranston PS, Epler JH (2013) The larvae of Chironomidae (Diptera) of the Holarctic region -

keys and diagnoses. Insect Syst Evol Supp 66:1–571

Appleby PG (2001) Chronostratigraphic techniques in recent sediments. In: Last WM, Smol JP (eds)

Tracking Environmental Change Using Lake Sediments. Volume 1: Basin Analysis, Coring and

Chronological Techniques. Kluwer Academic Publishers, Dordrecht, pp 172–203

Barley EM, Walker IR, Kurek J, Cwynar LC, Mathewes RW, Gajewski K, Finney BP (2006) A northwest

North American training set: distribution of freshwater midges in relation to air temperature and

lake depth. J Paleolimnol 36:295–314

Bennion H, Sayer CD, Tibby J, Carrick HJ (2010) Diatoms as indicators of environmental change in

shallow lakes. In: Smol J, Stoermer EF, editors. The Diatoms. 2nd edition. Cambridge University

Press, Cambridge, pp 152–173

Birks HJB (1995) Quantitative palaeoenvironmental reconstructions. In: Maddy D, Brews J (eds)

Statistical Modelling of Quaternary Science Data. Technical Guide 5. Quaternary Research

Association, Cambridge, pp 161–254

Blais JM, Kimpe LE, McMahon D, Keatley BE, Mallory ML, Douglas MSV, Smol JP (2005) Arctic

seabirds transport marine-derived contaminants. Science 309:445

Brodersen KP, Lindegaard C (1997) Significance of subfossile chironomid remains in classification of

shallow lakes. Hydrobiologia 342/343:125–132

Brodersen KP, Lindegaard C (1999) Classification, assessment and trophic reconstruction of Danish lakes

using chironomids. Freshw Biol 42:143–157

Brodersen KP, Ovgaard BV, Vestergaard O, Anderson NJ (2001) Chironomid stratigraphy in the shallow

and eutrophic Lake Søbygaard, Denmark: chironomid-macrophyte co-occurrence. Freshw Biol

46:253–267

Brodersen KP, Pedersen O, Walker IR, Jensen MT (2008) Respiration of midges (Diptera; Chironomidae)

in British Columbian lakes: oxy-regulation, temperature and their role as palaeo-indicators.

Freshw Biol 53:593–602

Brodersen KP, Quinlan R (2006) Midges as paleoindicators of lake productivity, eutrophication and

hypolimnetic oxygen. Quat Sci Rev 25:1995–2012

Brooks SJ, Bennion H, Birks HJB (2001) Tracing lake trophic history with a chironomid-total phosphorus

inference model. Freshw Biol 46:513–533

Brooks SJ, Langdon PG, Heiri O (2007) The Identification and Use of PalaeArctic Chironomidae Larvae

in Palaeoecology. QRA Technical Guide No. 10. Quaternary Research Association, London

Brundin L (1956) Die bodenfaunistischen Seetypen un ihre Anwenbarkeit auf die Südhalbkugel. Zugleich

eine Theorie der produktionsbiologischen Bedeutung der glazialen Erosion. Report/Institute of

Fresh-water Research Drottningholm 37:186–235

Clerk S, Hall R, Quinlan R, Smol JP (2000) Quantitative inferences of past hypolimnetic anoxia and

nutrient levels from a Canadian Precambrian Shield lake. J Paleolimnol 23: 319–336

Czeczuga B (1960) Haemoglobin content of the larvae of Tendipes fl. plumosus L. from various levels of

bed sediment. Nature 186:484

Douglas MSV, Smol JP (2000) Eutrophication and recovery in the High Arctic: Meretta Lake (Cornwallis

Island, Nunavut, Canada) revisited. Hydrobiologia 431:193–204

Edmonson WT (1974) The sedimentary record of the eutrophication of Lake Washington. Proc Nat Acad

Sci USA 71:5093–5095

Eggermont H, Heiri O (2012) The chironomid-temperature relationship: expression in nature and

palaenvironmental implications. Biol Rev 87:430–456

Finney BP, Gregory-Eaves I, Sweetman J, Douglas MSV, Smol J (2000) Impacts of climatic change and

fishing on Pacific salmon abundance over the past 300 years. Science 290:795–799

Ginn BK, Rühland KM, Young JD, Hawryshyn J, Quinlan R, Dillon PJ, Smol JP (2012) The perils of

using sedimentary phosphorus concentrations for inferring long-term changes in lake nutrient

levels: Comments on Hiriart-Baer et al., 2011. J Gt Lakes Res 38:825–829

Hall RI, Smol JP (2010) Diatoms as indicators of lake eutrophication. In: Smol JP, Stoermer EF (eds) The

Diatoms: Applications for the Environmental and Earth Sciences. Cambridge University Press,

Cambridge, pp 122–151

Hamburger K, Dall PC, Lindegaard C (1995) Energy metabolism of Chironomus anthracinus (Diptera,

Chironomidae) under laboratory and field conditions. Hydrobiol 297:187–200

Hargan KE, Stewart EM, Michelutti N, Grooms C, Kimpe L, Mallory M, Smol JP, Blais JM (in review)

Incorporating sterols and stanols as biomarkers for tracking waterbird impacts to temperate

ponds. Proc R Soc Lond B Biol Sci: manuscript ID RSPB-2017-2669

Heinis F, Sweerts J-P, Loopie E (1994) Micro-environment of chironomids larvae in the littoral and

profundal zones of Lake Maarsseveen I, The Netherlands. Fundam Appl Limnol 124:173–289

Heiri O, Brooks SJ, Birks HJB, Lotter AF (2011) a 274-lake calibration data-set and inference model for

chironomid-based summer air temperature reconstruction in Europe. Quat Sci Rev 30:3445–3456

Hofmann W (1988) The significance of chironomid analysis (Insecta: Diptera) for paleolimnological

research. Palaeogeogr Palaeoclim Palaeoecol 62:501–509

Int Panis L, Godeeris B, Verheyen RF (1995) On the relationship between oxygen microstratification in a

pond and the spatial distribution of the benthic chironomid fauna. In: Cranston P (ed)

Chironomids: From Genes to Ecosystems. CSIRO Publications, Melbourne, pp 323–328

Jacobson PC, Hansen GJA, Bethke BJ, Cross TK (2017) Disentangling the effects of a century of

eutrophication and climate warming on freshwater lake fish assemblages. PLoS ONE

12:e0182667

Jacobson PC, Stefan HG, Pereira DL (2010) Coldwater fish oxythermal habitat in Minnesota lakes:

influence of total phosphorus, July air temperature, and relative depth. Can J Fish Aquat Sci

67:2002–2013

Juggins S, Birks HJB (2012) Quantitative environmental reconstructions from biological data. In: Birks

HJB, Lotter AF, Juggins S, Smol JP (eds) Tracking Environmental Change Using Lake

Sediments. Volume 5: Data Handling and Numerical Techniques. Springer, Dordrecht, pp 201–

248

Keatley BE, Douglas MSV, Blais JM, Mallory ML, Smol JP (2009) Impacts of seabird-derived nutrients

on water quality and diatom assemblages from Cape Vera, Devon Island, Canadian High Arctic.

Hydrobiologia 61:191– 205

Korosi JB, Cheng W, Blais JM (2015) Organic pollutants in sediment core archives. In: Blais JM, Rosen

MR, Smol JP (eds) Developments in Paleoenvironmental Research volume 18: Environmental

Contaminants. Springer, Dordrecht, pp 161–185

Langdon PG, Ruiz Z, Brodersen KP, Foster IDL (2006) Assessing lake eutrophication using chironomids:

understanding the nature of community response in different lake types. Freshw Biol 51:562–577

Larocque I, Rolland N (2006) A visual guide to sub-fossil chironomids from Québec to Ellesmere Island.

Rapport de recherche R-900. Institut National de la Recherche Scientifique, Québec

Leeming R, Ball A, Ashbolt N, Nichols P (1996) Using faecal sterols from humans and animals to

distinguish faecal pollution in receiving waters. Water Res 30:2893−2900

Little J, Hall RI, Quinlan R, Smol JP (2000) Past trophic status and hypolimnetic anoxia during

eutrophication and remediation of Gravenhurst Bay, Ontario: comparison of diatoms,

chironomids, and historical records. Can J Fish Aquat Sci 57:333–341

Little JL, Smol JP (2001) A chironomid-based model for inferring late-summer hypolimnetic oxygen in

southeastern Ontario lakes. J Paleolimnol 26:259–270

Luoto TP, Salonen V-P (2010) Fossil midge larvae (Diptera: Chironomidae) as quantitative indicators of

late-winter hypolimnetic oxygen in southern Finland: a calibration model, case studies and

potentialities. Boreal Environ Res 15:1–18

Michelutti N, Mallory ML, Blais JM, Douglas MSV, Smol JP (2011) Chironomid assemblages from

seabird-affected high Arctic ponds. Polar Biol 34:799–812

Michelutti N, Smol JP (2016) Visible spectroscopy reliably tracks trends in paleo-production. J

Paleolimnol 56:253–265

Porinchu DF, Moser KA, Munroe JS (2007) Development of a midge-based summer surface water

temperature inference model for the Great Basin of the western United States. Arct Antarct Alp

Res 39:566–577

Quinlan R, Smol JP, Hall RI (1998) Quantitative inferences of past hypolimnetic anoxia in south-central

Ontario lakes using fossil midges (Diptera: Chironomidae). Can J Fish Aquat Sci 55:587–596

Quinlan R, Smol JP (2001) Chironomid-based inference models for estimating end-of-summer

hypolimnetic oxygen from south-central Ontario lakes. Freshw Biol 46:1529–1551

Rolland N, Larocque I, Francus P, Pienitz R, Laperriére L (2008) Holocene climate inferred from

Biological (Diptera: Chironomidae) analyses in a Southampton (Nunavut, Canada) lake.

Holocene 18:229–241

Sæther OA (1979) Chironomid communities as water quality indicators. Holarctic Ecol 2:65–74

Scheffer M, Hosper SH, Meijer M-L, Moss B, Jeppesen E (1993) Alternative equilibria in shallow lakes.

Trends Ecol Evolut 8:275–278

Scheffer M, Jeppesen E (2007) Regime shifts in shallow lakes. Ecosystems 10:1–3

Schindler DW (1977) Evolution of phosphorus limitation in lakes. Science 195:260–262

Schindler DW, Carpenter SR, Chapra SC, Hecky RE, Orihel DM (2016) Reducing phosphorus to curb

lake eutrophication is a success. Environ Sci Technol 50:8923–8929

Scholz F, Zerbst-Boroffka I (1998) Environmental hypoxia affects osmotic and ionic regulation in

freshwater midge-larvae. J Insect Physiol 44:427–436

Seehausen O, van Alphen JJM, Witte F (1997) Cichlid fish diversity threatened by eutrophication that

curbs sexual selection. Science 277:1808–181

Smith VH (2003) Eutrophication of freshwater and coastal marine ecosystems: a global problem. Environ

Sci Pollut Res 10:126–139

Smith VH, Joye SB, Howarth RW (2006) Eutrophication of freshwater and marine ecosystems. Limnol

Oceanogr 51:351–355

Smith VH, Schindler DW (2009) Eutrophication science: Where do we go from here? Trends Ecol Evol

24:201–207

Smol JP (2008) Pollution of Lakes and Rivers: A Paleoenvironmental Perspective. Blackwell Publishing,

Malden

Smol JP (2010) The power of the past: using sediments to track the effects of multiple stressors on lake

ecosystems. Freshw Biol 55:43–59

Stewart EM, Michelutti N, Blais JM, Mallory ML, Douglas MSV, Smol JP (2013) Contrasting the effects

of climate, nutrient, and oxygen dynamics on subfossil chironomid assemblages: a

paleolimnological experiment from eutrophic High Arctic ponds. J Paleolimnol 49:205–219

Stewart EM, Michelutti N, Shenstone-Harris S, Grooms C, Weseloh C, Kimpe LE, Blais JM, Smol JP

(2015) Tracking the history and ecological changes of rising Double-crested Cormorant

populations using pond sediments from islands in eastern Lake Ontario. PLoS ONE 10:e0134167

Thienemann A (1920) Untersuchungen über die Bziehungen zwischen dem Sauerstoffgehaly des Wassers

und der zusammensetzung der Fauna in norddeutschen Seen. Arch Hydrobiol 12:1–65

Thienemann A (1922) Die beiden Chironomustarten der Tiefenfauna der norddeutschen Seen. Ein

hydrobiologisches Problem. Arch Hydrobiol 13:609–646

Thienemann A (1954) Chironomus. Leben, Verbreitung und wirtschaftliche Bedeutung der

Chironomiden. Binnengewӓsser 20:1–834

Vane CH, Kim AW, McGowan S, Leng MJ, Heaton TH, Kendrick CP, Coombs P, Yang H, Swann GE

(2010) Sedimentary records of sewage pollution using faecal markers in contrasting periurban

shallow lakes. Sci Tot Environ 409:345−56

Velle G, Brooks SJ, Birks HJB, Willassen E (2005) Chironomids as a tool for inferring Holocene climate:

an assessment based on six sites in southern Scandinavia. Quat Sci Rev 24:1429–1462

Walker IR (2001) Midges: Chironomidae and related Diptera. In: Smol, JP, Birks HJB, Last WM (eds)

Tracking Environmental Change Using Lake Sediments volume 4: Zoological Indicators. Kluwer

Academic Publishers, Dordrecht, pp 43–66

Walker IR, Levesque AJ, Cwynar LC, Lotter AF (1997) An expanded surface-water palaeotemperature

inference model for use with fossil midges from eastern Canada. J Paleolimnol 18:165–178

Walker IR, Mathewes RW (1989) Chironomidae (Diptera) remains in sudicial lake sediments from the

Canadian Cordillera: analysis of the fauna across an altitudinal gradient. J Paleolimnol 2:61–80

Weber RE (1980) Functions of invertebrate haemoglobins with special reference to adaptations to

environmental hypoxia. Amer Zool 20:79–101

Wolfe AP, Vinebrooke RD, Michelutti N, Rivard B, Das B (2006) Experimental calibration of lake-

sediment spectral reflectance to chlorophyll a concentrations: methodology and paleolimnological

validation. J Paleolimnol 36:91–100

Zhang E, Bedford A, Jones R, Shen J, Wang S, Tang H (2006) A subfossil chironomid-total phosphorus

inference model for lakes in the middle and lower reaches of the Yangtze River. Chinese Sci Bull

51:2125–2132

Table 1.1 Possible negative consequences of freshwater eutrophication (Smith and Schindler 2009; Smith 2003).  Increased primary production (phytoplankton, periphyton, or macrophytes)  Increased consumer biomass  Shifts to phytoplankton and bloom-forming algae  Potential cynanobacterial blooms and the production of toxins  Decreased hypolimnetic dissolved oxygen availability  Increased turbidity and decreased light penetration  Fish kills  Reduced species diversity or threats to endangered species  Taste and odour issues  Increased difficulty in water treatment processes  Aesthetic issues and resultant declines in economic or recreational value/use

Figure 1.1 The dynamics of eutrophication in A) deep stratifying lakes, where decomposition of algal blooms in the hypolimnion can cause hypoxia leading to fish kills, and B) shallow lakes with alternate stable states, where macrophytes provide resistance to algal blooms and high turbidity when undergoing eutrophication, up until a point (as described by Scheffer et al. 1993).

Figure 1.2 Chironomid assemblages from bird-impacted ponds (A-C) and a non-impacted pond (D) at Cape Vera, on Devon Island (NU) showed that characteristic eutrophication-indicator species, Chironomus plumosus-type, did not change with increased nutrients as indicated by δ15N (see Keatley et al. 2009), but increases in total chironomid abundance occurred (HC per g DM). Modified from Stewart et al. (2013). 19

Chapter 2

Assessing the efficacy of chironomid and diatom assemblages in tracking

eutrophication in High Arctic sewage ponds

Published as and formatted for:

Stewart, E.M., R. McIver, N. Michelutti, M.S.V. Douglas & J.P. Smol, 2014. Assessing the efficacy of chironomid and diatom assemblages in tracking eutrophication in High Arctic sewage ponds.

Hydrobiologia 721: 251–268.

Abstract

Eutrophication is the most common water quality issue affecting freshwaters worldwide.

Paleolimnological approaches have been used in temperate regions to track eutrophication over time, placing changes in historical context. Diatoms (Bacillariophyta) have a direct physiological response to changes in nutrients and are effective indicators of lake trophic status. Chironomids (Diptera) have also been used to track nutrient conditions; however, given that nutrients and oxygen are often tightly linked, it is difficult to disentangle which variable is driving shifts in assemblages. Here, we analyze chironomid and diatom remains in sediments from sewage-impacted ponds in the High Arctic. These ponds have the unusual characteristics of elevated nutrient and oxygen concentrations, unlike those of typical eutrophic lakes where deepwater oxygen is often depleted. Our data show that, while diatom assemblages responded to changing nutrients, no concomitant changes in chironomid assemblage composition were recorded. Furthermore, the dominance of oligotrophic, cold stenothermic chironomid taxa and lack of so- called “eutrophic” species in the eutrophic sewage ponds suggests that oxygen, not nutrients, structures chironomid assemblages at these sites. These data support findings from eutrophic High Arctic ponds affected by seabird-derived nutrients and provide evidence that chironomid assemblage changes are driven partly by the strong influence of climate.

Keywords: paleolimnology, chironomids, diatoms, eutrophication, High Arctic

Introduction

Sedimentary chironomid (Diptera) remains are often used in paleolimnological investigations of eutrophication because they respond to numerous production-related variables. As a result, chironomid- based inference models have been developed to infer concentrations of chlorophyll a (chl-a) (Brodersen &

Lindegaard, 1999), total phosphorous (TP) (Brooks et al., 2001), and hypolimnetic oxygen (Quinlan et al.,

1998). However, in temperate dimictic lakes, nutrients and deepwater oxygen concentrations are often tightly linked (because of interactions between primary production and decomposition) and so, in most deep eutrophic lakes, hypolimnetic oxygen is low. Monitoring and experimental evidence have documented the many physiological, morphological, and behavioural adaptations that allow some chironomid species to tolerate low oxygen conditions (Brodersen & Quinlan, 2006). However, there have been difficulties disentangling the respective roles of nutrients and oxygen (under eutrophic conditions especially) in structuring chironomid assemblages because chironomids also respond to the increased food availability and changing habitat structure that may occur with increased nutrient levels.

One way in which to ascertain whether chironomids are responding primarily to nutrients or oxygen is by studying sites where oxygen and nutrient concentrations are not closely associated with one another. Although rare, such limnological conditions were documented in a series of High Arctic ponds near Cape Vera on Devon Island, Nunavut (Blais et al., 2005). The Cape Vera ponds are eutrophic due to their proximity to a large colony of seabirds that releases nutrient-rich wastes to the surrounding catchment. However, in contrast to most eutrophic temperate lakes, the oxygen levels in the ponds remain high due to their shallow nature (Zmax < 2m) and 24 hours of daylight during the growing season (i.e. no periods of respiration without concurrent photosynthesis). Because of the varying distances of each pond to the nesting colony, there exists a gradient of high-to-low nutrient concentrations extending away from the colony (Keatley et al., 2009). Although the impacted ponds of Cape Vera are highly eutrophic

(Keatley et al., 2009), the chironomid species recorded in both surface sediments (Michelutti et al., 2011)

and sediment cores (Stewart et al., 2013) of these ponds contain mainly cold stenothermic taxa that are typical of oligotrophic waters. Furthermore, sediments from the eutrophic ponds recorded remarkably similar species to those of a nearby oligotrophic control site outside the area affected by the seabird colony (Stewart et al., 2013). The similarity of chironomid community composition along a large nutrient gradient suggested that dissolved oxygen, which was elevated in all study sites, played a more important role than nutrients in governing chironomid assemblages.

The Cape Vera study (Stewart et al., 2013) challenged some widely held beliefs about the applicability of chironomids in eutrophication studies, namely their suitability for quantitative inference models of TP. However, this was the only study completed to date that was able to explicitly evaluate the relative influences of nutrient levels versus oxygen concentrations on chironomid assemblages, as typically lakewater TP and deepwater oxygen levels are negatively correlated. The question remained whether another “natural laboratory”, such as the Cape Vera ponds, could be found to further investigate these important relationships.

Several sewage-affected ponds near Resolute Bay on Cornwallis Island (Nunavut, Canada; Figure

2.1) provided the opportunity to evaluate and expand on the conclusions reached in the Cape Vera studies as to whether chironomids are responding primarily to nutrients or oxygen with eutrophication. In contrast to Cape Vera, where nutrient-rich seabird wastes have eutrophied the ponds since their inception several centuries ago (Michelutti et al., 2009), a series of shallow ponds near Resolute Bay have only recently been eutrophied as a result of human sewage inputs (Figure 2.1; Schindler et al., 1974).

Beginning in 1949, sewage from a newly-constructed Department of Transport Base (i.e. airport facilities) was released onto the landscape via two small watercourses (one of which crossed several naturally- existing ponds referred to as the “sewage ponds”) before ultimately ending up in a terminal lake (Meretta

Lake; Figure 2.1; Douglas & Smol, 2000). In 1979, effluent from one of the sewage outlets (i.e. utilidor) was dismantled because of a steady decline in the number of residents, and all inputs from the other utilidor to Meretta Lake stopped completely by 1998 (Figure 2.1; Douglas & Smol, 2000). The Resolute

Bay sewage ponds contained elevated oxygen levels (Supplemental Figure 2.2), for similar reasons to the 22

seabird-affected Cape Vera ponds, namely that they are shallow and thus well-mixed, and have no periods of respiration without concurrent photosynthesis during the 24-hour daylight in the summer months. Like the shallow Cape Vera ponds, the Resolute Bay sewage ponds freeze solid through to the sediments during the winter.

One of our main study questions is “How have the chironomid assemblages in the sewage ponds responded to eutrophication and recovery, all the while remaining in an oxic environment?” We answer this question by examining subfossil chironomid remains in sediment cores and in surface sediment samples recovered during a long-term monitoring program that we began in the 1990s. By tracking the biological response to eutrophication and recovery within each Resolute sewage pond, we reduce any variability that may arise from inter-pond comparisons, such as was done at Cape Vera (Stewart et al.,

2013), where we used a natural nutrient gradient (i.e. distance of each pond from nesting colony) to compare the changes in chironomid assemblage associated with varying degrees of eutrophication.

In addition to chironomids, we also examine subfossil diatom assemblages from the Resolute Bay sewage ponds. Because diatoms have a direct physiological dependence on nutrients (i.e. they uptake nutrients directly from the water column and respond to habitat changes), they are often used for tracking eutrophication (Hall & Smol, 2010). To date, only a few paleolimnological studies have used diatoms to track cultural eutrophication in the Arctic, including one from Meretta Lake, the terminal lake for the

Department of Transport Base sewage inputs at Resolute Bay (Douglas & Smol, 2000). Previous studies have shown that diatoms in Arctic lakes and ponds often display a relatively subtle taxonomic shift in response to nutrient enrichment (Douglas & Smol, 2000), and even a delayed response (Michelutti et al.,

2007a), compared to changes recorded in temperate regions. This muted response is believed to be related to the overriding influence of the cool climate on aquatic biota in Arctic lakes and ponds (Douglas &

Smol, 2000; Michelutti et al., 2007a; Smol & Douglas, 2007a). This relationship illustrates an important caveat of using any bioindicator: diatoms, like most aquatic biota, will respond to multiple, interacting environmental drivers, such as nutrients as well as climate, and thus the relative influences of each must be considered. Nonetheless, although the diatom response to eutrophication in the Arctic is not as marked 23

as in temperate regions, the diatoms do change with nutrient additions, and there are documented differences in diatom assemblages between eutrophic and oligotrophic sites (Keatley et al., 2009) from the same region as our previous chironomid work (Michelutti et al., 2011; Stewart et al., 2013).

In this current study, the inclusion of both diatom and chironomid indicators allows us to compare the responses to eutrophication between primary producers (diatoms) and primary consumers

(chironomids). Our limnological monitoring program of the Resolute Bay sewage ponds spans from 1993 to 2011, making it one of the longest of its kind in the High Arctic. The sewage ponds, as well as nearby control ponds unaffected by sewage inputs, have been sampled for water chemistry, surface sediments, and rock scrapes (epilithon). In addition to the annual collections, sediment cores were recovered from the sewage ponds and nearby control sites, which provide a long-term record of chironomid and diatom changes over time. Our main objectives were to: (1) compare the responses of chironomid and diatom assemblages in the Resolute Bay sewage ponds to eutrophication and subsequent recovery; (2) compare the chironomid and diatom assemblages in sewage ponds to those in nearby control ponds that have never received human sewage inputs; and (3) assess the potential and efficacy of using chironomids and diatoms for understanding eutrophication and reconstructing production-related variables, while exploring the confounding effects of recent climate warming (Smol, 2010).

Materials and methods

Site description Resolute Bay (74º41’11” N, 94º54’33” W), Nunavut, is on the southwestern coast of Cornwallis

Island in the Canadian Arctic Archipelago (Figure 2.1). Resolute Bay and the surrounding area are classified as having a polar desert climate with a current mean annual temperature of -16.4ºC and mean annual precipitation of 150 mm (Environment Canada, 2012). A weather station was built near Resolute

Bay in 1948, and an air base (Department of Transportation, or “North Base”) was constructed in 1949.

North Base held a working population of ~150 people between 1949 and 1971, and the sewage from the base was disposed of directly onto the landscape through a series of utilidors or above ground pipes

(photos in Douglas & Smol, 2000). Sewage flowed from each utilidor output along a north-to-south watercourse that spanned ~2 km until reaching the terminal site, Meretta Lake (Figure 2.1). Our study sites include the four “sewage” ponds along one of the disposal pathways north of Meretta Lake.

The utilidor that released sewage through the pathway containing the four sewage ponds was dismantled in 1979, and all sewage inputs along this watercourse ceased at that time. A second utilidor remained in operation at another location until 1998; however, none of the sewage along this pathway would have entered into the four sewage ponds used in this study. The sewage ponds were informally named R-10, R-11, R-12, and R-13; with sewage outputs traveling from R-13, through R-12, then R-11, and finally R-10, before reaching Meretta Lake (Figure 2.1, 2.2). In addition to the sewage ponds, two nearby control sites, informally named R-1 and R-2, that have never received sewage inputs were included in the study (Figure 2.1, 2.2).

The pH of the sewage ponds over our 18-year sampling period remained stable at pH ~9, whereas the control ponds maintained a pH of 8.5 over the same time period. Temperature in the Resolute Bay area has steadily increased from 1948, when meteorological measurements were first taken, until 2010

(Environment Canada, 2012). This 60-year period shows a total increase in mean annual temperature from -17˚C to -16˚C (Figure 2.3). All ponds in this study are shallow and therefore reach relatively warm temperatures (e.g. 5 – 10°C) compared to deep ice-covered Arctic lakes during the short growing season

(Figure 2.4, 2.5).

Water chemistry In 1993, 2002, 2006, 2009, and 2011, water samples for chemical analyses of the sewage ponds

(R-10, R-11, R-12, and R-13) were taken following the identical procedures used by our lab in other High

Arctic limnological work over the previous three decades (e.g. Antoniades et al., 2003). In 1992, then each year from 1994 to 2009, and then again for 2011, water samples were collected for the control ponds

(R-1 and R-2). Water was collected from ~10 cm below the surface, and was filtered on-site for chlorophyll a (chl-a) and total filtered nitrogen (TN-f) according to standard methods provided by

Environment Canada (1979). Water samples for total unfiltered phosphorus (TP-u), as well as the chl-a and TN-f filters, were analyzed at the National Laboratory for Environmental Testing (NLET) in

Burlington, Ontario. Specific conductance and pH measurements were obtained on-site using field meters.

Epilithic diatom sampling Rock scrapes for diatom analyses were collected during the same approximate time interval each year (mid-July), which allows for more meaningful comparisons over individual years. The sewage ponds were sampled in 1993, 2002, 2006, 2009, and 2011. These samples represent the only diatom data from

R-10 and R-11, for which sediment cores could not be collected due to the rocky bottom substrate. In

1993, R-13 was considered too hazardous to sample due to the high concentration of sewage still present in the pond, and therefore the first sampling of this site did not occur until 2006. The rock scrape samples were a composite of four to five rocks collected from different locations in the ponds, all of which were scraped with a small brush, rinsed into a plastic scintillation vial, and preserved with Lugol’s solution.

Surface sediment and sediment core sampling Using identical sampling methods each year, the uppermost ~1 cm of surface sediments of the sewage ponds was collected by hand in 1993, 2002, 2006, 2009, and 2011. In addition, sediment cores were retrieved using a 7.6-cm diameter Lexan core tube. The sediment cores were sectioned on-site in

0.25-cm intervals (R-12 and R-13) or 1-cm intervals (R-1 and R-2) using a Glew (1988) extruder. The sediment cores from R-12 and R-13 were taken in July 2011 and those from R-1 and R-2 in July of 1992.

Both chironomids and diatoms were analyzed from the same sediment core intervals. Sediment cores were dated using excess 210Pb activities and developed into core chronologies using the constant-rate-of- supply (CRS) model (Appleby & Oldfield, 1978). The 210Pb chronology was verified using 137Cs, which indicates the circa 1963 peak of aboveground nuclear weapons testing. The 210Pb activities are presented in Supplemental Figure 2.1. The control pond, R-1, had insufficient 210Pb activities, and so accurate dates could not be generated – a problem that is common in High Arctic environments (Wolfe et al., 2004).

Diatom processing and assessment. Diatoms slides were prepared for the sediment cores from ponds R-1, R-2, R-12, and R-13 and from the epilithon samples from ponds R-10, R-11, R-12, and R-13 using the procedures of Battarbee et al. (2001). Siliceous diatom valves were identified and enumerated using a Leica DMR HC light microscope. A minimum of 400 diatom valves were identified to the species level for each sample, primarily following the taxonomy of Antoniades et al. (2008). Stratigraphies were drawn using C2 Data

Analysis Program version 1.7.2 and edited with CorelDRAW Graphics Suite 12.

Chironomid head capsule collection and assessment The collection of chironomid head capsules followed standard paleolimnological procedures outlined in Walker (2001). Identification to the lowest taxonomic level possible was achieved using a

Leica DMR HC light microscope set to brightfield illumination at 100X – 400X magnification and primarily following the taxonomic guide of Brooks et al. (2007). A minimum of 50 whole head capsules was identified for each interval (when possible), as it is the statistically relevant minimum count required for making inferences about chironomid assemblages (Quinlan & Smol, 2001a). Minimum head capsule counts were not achieved for all samples (discussed below), but are reported nonetheless as they still provide important ecological information about the assemblage.

Chironomid head capsules were collected from the surface sediments of R-10, R-11, R-12, and R-

13 (between the years of 1993 and 2011; data provided as supplementary material in Supplemental Table

2.1). Stratigraphies for the chironomid assemblages from the sediment cores of R-12 and R-13 (collected in July 2011), as well as R-1 and R-2 (collected in July 1992) were created using C2 Data Analysis

Program version 1.7.2 and edited with CorelDRAW Graphics Suite 12.

Statistical analysis Diatoms from the sediment cores were assessed by indirect ordination using a detrended correspondence analysis (DCA) in order to compare the assemblages of the sewage ponds (R-12 and R-

13) with those of the control ponds (R-1 and R-2). A similar analysis was performed for chironomids

from R-12, R-13, and R-2 using a Principal Component Analysis (PCA). The chironomid samples from

R-1 could not be included in this analysis because counts were too low to yield reliable relative abundance estimates (Quinlan & Smol, 2001a). Additionally, an individual PCA was run for each sediment core (R-12, R-13, R-1, and R-2) for both diatoms and chironomids revealing the main direction of variation in fossil assemblages over time (plotted as “PCA axis 1” on the right-hand side of each stratigraphy). All species data were log (x+1) transformed, and all analyses were completed using

CANOCO 5.

Results

Water chemistry The TP-u values in the control ponds have remained low over the past two decades with values ranging from below detection limits (<0.2 µg/L) to 6.5 µg/L in R-1 and from 0.4 to 24 µg/L in R-2

(Figure 2.4). This is in marked contrast to the sewage ponds that, following the time of sewage inputs, recorded values ranging from 35 - 883 µg/L (Figure 2.5). TP-u values in the sewage ponds decreased dramatically early in our monitoring program. For example, over a ~10-year period between 1993 and

2002, TP-u values dropped from 883.5 to 4.1 µg/L in R-12, from 435 to 12.4 µg/L in R-11, and from 35 to 4.7 µg/L in R-10. R-13 was not sampled (due to health concerns from the high concentration of sewage in the pond) until 2006, and so no record of a dramatic decrease in measured nutrients can be documented, although recovery could be surmised from simple visible changes in the pond.

Values of TN were similarly greatly elevated in the sewage ponds compared to the control ponds, as is clearly demonstrated by the ranges of measured values (Figure 2.4, 2.5). TN values in the sewage ponds ranged from 0.268-1.18 mg/L (Figure 2.5), whereas the range of TN values in the control ponds spanned from 0.159-0.453 mg/L over the ~20-year period of water chemistry sampling (Figure 2.4). The

TN values of the control ponds show an increase of approximately 0.1 mg/L over the past few years of sampling, and approximately a 0.6 mg/L increase in the sewage ponds.

The values for chl-a in the sewage ponds were elevated compared to the values in the control ponds. For example, in 1993 the chl-a in R-11 was 8.3 µg/L and 39.9 µg/L in R-12 (Figure 2.5), whereas in the two control ponds it was 1 µg/L in R-1 and 2 µg/L in R-2 in 1992 (Figure 2.4). The chl-a values in

R-1 and R-2 remained relatively stable until 2011, although the values in R-11 and R-12 decreased to 0.3

µg/L and 2.1 µg/L, respectively, over the same time.

The specific conductance for the two control ponds were similar in value, ranging from 78 µS/cm to 261 µS/cm in R-1 and 118 µS/cm to 349 µS/cm in R-2 (Figure 2.4). In both control ponds, specific conductance increased between 2008 and 2011. R-10 is the largest sewage pond that is furthest from the sewage output, and showed a very similar pattern in specific conductance to the control ponds (Figure

2.5). However, the specific conductance in the sewage ponds R-11 and R-12 was higher than in the control ponds and showed a pattern of decline from 1993 to 2006 (Figure 2.5). In R-11 the specific conductance was 405 µS/cm in 1993 and declined to 215 µS/cm by 2006, staying relatively stable near

200 µS/cm. R-12 had a specific conductance of 680 µS/cm in 1993 which decreased to 249 µS/cm by

2006. R-13 showed no discernible pattern in specific conductance, likely because it was not sampled until

2006, though its values from 2006 and onwards were similar to those of R-11 and R-12 (Figure 2.5). In all sewage ponds, there is a slight increase in specific conductance from 2006 to 2011 (Figure 2.5).

Epilithic diatom rock scrapes The epilithic assemblages in the four sewage ponds were dominated by eutrophic indicators including high relative abundances of Nitzschia perminuta, Nitzschia alpina, and Fistulifera saprophila

(Figure 2.6). The dominance of these three taxa has gradually declined since 1993 in R-10, R-11, and R-

12, and Cymbella cleve-eulerai and Achnanthidium minutissimum have increased in abundance. A nutrient gradient is apparent among the four sewage ponds, with decreasing dominance of the nutrient- tolerant taxa (e.g., N. perminuta, N. alpina, F. saprophila) in the ponds further from the utilidor outfall

(e.g., R-10; Figure 2.1, 2.6).

Sediment core diatom assemblages 29

The diatom assemblages in the sewage pond sediment cores (R-12 and R-13, Figure 2.7A, B) were characterized by high relative abundances of Nitzschia perminuta and Nitzschia alpina. In R-12 at 4 cm depth, two taxa, Cyclotella striata and Cyclotella meneghiniana, appear alongside Stephanodiscus minutulus, all of which subsequently disappear circa 1970 (Figure 2.7A). This finding is reflected in the

PCA axis 1 scores, which show a sharp increase and decrease during the same time (Figure 2.7A). In contrast, the diatom assemblage in R-13 remained relatively stable over time, though a change in the PCA axis 1 scores likely reflect subtle decreases in Nitzschia taxa and the appearance of Cymbella cleve- eulerai by the late 1990s (Figure 2.7B). The two control ponds (R-1 and R-2) recorded markedly different assemblages compared to the sewage ponds, with dominant species including Achnanthidium minutissimum and several Cymbella sensu lato taxa (Figure 2.7C, D). This is further supported by the fact that sewage pond sediment intervals group separately from the control pond intervals in the DCA biplot

(Figure 2.9B, C).

Surface sediment chironomid assemblages Low chironomid abundances were recorded in the surface sediment collections of all ponds, and therefore absolute abundances are reported, as relative abundances may be misleading with a low sample size (Quinlan & Smol, 2001a). However, species presence-absence data are provided for each pond for every year that sediments were collected (Supplemental Table 2.1). As no sediment cores were taken from R-10 and R-11, these surface sediment data are the only samples of chironomid assemblages in these ponds. Examination of the surface samples showed that undifferentiated Tanytarsini species were most abundant, followed by Tanytarsus gracilentus-type. Other common chironomids included

Corynoneura arctica-type, Hydrobaenus/Oliveridia, Metriocnemus hygropetricus-type, Psectrocladius group, Chironomus plumosus-type, and Procladius. Chironomid head capsules were notably more abundant in the two ponds closest to the sewage outfall, R-12 and R-13, compared to the two ponds furthest from the sewage outfall, R-10 and R-11.

Sediment core chironomid assemblages 30

Both sewage ponds (R-12, R-13) have similar chironomid assemblages, with the most common taxa being undifferentiated Tanytarsini species, followed by T. gracilentus-type, as well as C. arctica- type, Hydrobaenus/Oliveridia, M. hygropetricus-type, and Psectrocladius group (Figure 2.8A, B). Trace abundances of C. plumosus-type, and Procladius were present in both R-12 and R-13. No major changes in chironomid assemblage are apparent over the time of sewage inputs (1949-1979), although by the late-

1980s C. arctica-type appears and continues to increase in relative abundance to the present, as reflected by increases in the PCA axis 1 scores from negative to positive values (Figure 2.8A, B). Head capsule abundance increases from the bottom of the core and remains constant at ~200-300 head capsules per gram dry sediment by the 1970s for both R-12 and R-13, with slight decreases beginning in the early

2000s.

The chironomid assemblages recorded in the oligotrophic control ponds (R-1, R-2) are similar in terms of the species present to those in the eutrophic sewage ponds (R-12 and R-13, Figure 2.8).

Undifferentiated Tanytarsini species, C. arctica-type, Hydrobaenus/Oliveridia, M. hygropetricus-type, and Psectrocladius group dominate, with low abundances of C. plumosus-type and Procladius (Figure

2.8C, D). The most notable difference between the control and sewage ponds is not in the taxa that make up the assemblages, rather that the relative abundance of C. plumosus-type is slightly greater in the sewage ponds, which typically had 30-40 times greater concentrations of head capsules per gram dry weight compared to the control ponds (Figure 2.8). In R-2, a notable increase in head capsules per gram dry sediment (HC/g DM) occurs near the surface of the core (Figure 2.8D). According to a PCA, the sewage ponds sediment intervals separate from the control pond intervals based primarily on differences in the relative abundances of C. plumosus-type, though this must be interpreted with caution because counts from the control ponds were low (Figure 2.9A).

Discussion

Limnological responses to changing sewage inputs

Our limnological monitoring program of the sewage ponds began in 1992, which is 13 years after the cessation of direct sewage inputs to the ponds (a history of the sewage inputs is discussed in Douglas

& Smol (2000)). Our data show changing production-related variables in the sewage pond waters over time (Figure 2.5). For example, in R-11, R-12 and R-13, TP-u, chl-a, and specific conductance decreased drastically from our 1993 sampling year to 2006, the next year data were collected (Figure 2.5). This trend is not apparent in R-10, which is the farthest from the sewage outlet and has a much larger catchment, which may be releasing nutrients from previous years’ accumulation (Figure 2.5). Because the water chemistry measurements begin 13 years after sewage inputs stopped, we do not record eutrophication maxima in the ponds, nor representative temporal trends of chemical recovery (including the rate or trajectory). However, the studies done of the terminal sewage basin, Meretta Lake, have more complete temporal records of recovery from eutrophication showing a slow continuous recovery from

1998 to the present (Antoniades et al., 2011; Douglas & Smol, 2000; Michelutti et al., 2002).

Our water chemistry data also show strong evidence of a chemical gradient among sewage ponds, as higher values of production-related variables are recorded in ponds closest to the sewage output. This trend would likely be apparent in R-13, but we do not have data for R-13 before 2006 (as in fact the pond appeared to be so contaminated by sewage we chose not to sample it before 2006 for health concerns). As expected, there is no evidence of declining nutrients and primary production in the control ponds R-1 and

R-2, and their concentrations are typical of other unaffected sites in the Arctic (Figure 2.4; Douglas &

Smol, 1994) and the rest of Cornwallis Island (Michelutti et al., 2007b). The only change in the water chemistry of the control ponds is the slight increase in specific conductance and TN that both ponds experience by 2008, which is most likely linked to recent warming, as Resolute Bay has seen increases in mean annual temperature over the past 60 years (Figure 2.3). Warming temperatures can be linked to greater evaporation during the ice-off period, and thereby concentrate ions in the water column (Smol et al., 2005; Smol & Douglas, 2007b). This trend is also apparent in the sewage ponds, as specific conductance shows increases from 2006 to 2011, though it should be noted that these changes are subtler than the early decreases linked to declining sewage inputs (Figure 2.5). Furthermore, TN has likewise 32

increased recently in the sewage ponds, however, unlike most other water chemistry variables measured,

TN was not recorded in our data set until 2002, and thus our data do not show the elevated TN levels we would expect in the early 1990s.

Diatom response to eutrophication and recovery from the sediment cores

The diatom assemblages in the sediment cores from the two sewage ponds (R-13 and R-12) are dominated by Nitzschia perminuta and Nitzschia alpina. Although these two Nitzschia taxa are common to Arctic freshwaters (Antoniades et al., 2008), in high latitude regions they typically only occur in large abundances in nutrient-rich, eutrophic lakes and ponds (e.g., Michelutti et al., 2007a; Keatley et al.,

2011). The markedly different diatom assemblages between the sewage ponds and the nearby control ponds (Figure 2.7, 2.9B) suggest that sewage inputs have altered the diatom assemblages in the affected ponds. This is in contrast to the fossil chironomid data that show no changes in species present in these same impacted and control ponds (Figure 2.8).

Though the response of the sewage pond diatoms to eutrophication is subtle, R-12 shows a likely response to sewage inputs around the 4-cm depth (before 1952 according to our 210Pb dates) as several taxa indicative of high conductivity (Cyclotella striata and Cyclotella meneghiniana) and high TP

(Stephanodiscus minutulus) appear, and subsequently disappear circa 1970. This change is reflected in a

PCA of the diatom assemblages in the R-12 core, as the sample score increased and decreased over the same time frame, indicating a notable assemblage change had occurred (Figure 2.7A). The disappearance of these taxa approximately corresponds to the cessation of sewage inputs to the ponds in 1979 (Douglas

& Smol, 2000), though not exactly due to the difficulty in obtaining precise dates. Our 210Pb dates should be treated as rough estimates of the timing of changes in our sediment cores, as 210Pb activity in most

Arctic sediments is often too low to give precise dates (Wolfe et al., 2004). The appearance and disappearance of diatom taxa indicative of high TP and high conductivity was not apparent in R-13 likely due to low resolution or dating issues, though both R-12 and R-13 record elevated abundances of

Nitzschia taxa, which, as noted above, are typically only recorded at such high abundances in eutrophic

systems in the High Arctic (e.g., Michelutti et al., 2007a; Keatley et al., 2011). The change in PCA axis 1 sample scores in R-13 by the late-1990s may reflect subtle decreases in the relative abundances of

Nitzschia taxa.

Epilithic diatom response to continuing recovery

The epilithic rock samples provide an annual resolution of diatom assemblages (because winter ice effectively scrapes away the previous year’s growth) for all of the sewage ponds and illustrate how closely diatoms track even subtle changes along the nutrient gradient among the sewage ponds.

Furthermore, the epilithic samples represent the only record of diatom assemblages available for Ponds R-

10 and R-11. Pond R-13 is closest to the sewage outfall, and the epilithon are dominated by Nitzschia taxa, which at these high abundances typically indicate elevated nutrient levels (Figure 2.6). In recent years, however, Nitzschia taxa have gradually declined and taxa characteristic of those found in the control sites, including Cymbella cleve-eulerai and Achnanthidium minutissimum, have increased (Figure

2.6D). This shift in diatom assemblages is consistent with improving water quality variables, such as TP- u (Figure 2.5), and therefore continuing biological recovery from eutrophication is evident in the epilithic diatom samples.

Ponds R-12 and R-11 are the next two closest ponds to the sewage outfall. In 1993, the epilithon in both sites was dominated by Fistulifera saprophila. This taxon is known to reach high abundances in eutrophic and polluted waters (Spaulding & Edlund, 2009). Indeed, F. saprophila reached its highest abundances during the period when TP concentrations were also highest (e.g., > 400 µg/L). As TP concentrations decreased below hyper-eutrophic levels (Figure 2.5), Nitzschia perminta and N. alpina increased in abundance at the expense of F. saprophila (Figure 2.6B, C).

Pond R-10 is the site furthest from the sewage outfall. In 1993, the epilithon in Pond R-10 record an assemblage dominated by Nitzschia taxa (Figure 2.6A). The absence of F. saprophila suggests that, not surprisingly due to its location, R-10 was less affected by nutrient inputs compared to ponds R-12 and

R-11. In the early 1990s, the epilithon is dominated by Cymbella, Fragilaria and Nitzschia taxa. This is in

contrast to the more eutrophic sites where one or two diatom taxa dominated the assemblages. Indeed, the higher diatom diversity recorded in the post-sewage epilithon of Pond R-10 is similar to the control pond sediment cores (Figure 2.6, 2.7).

Epilithic diatoms from the sewage ponds record assemblage changes that are consistent with declining nutrient concentrations and gradually improving water quality variables (Figure 2.5, 2.6).

Moreover, comparing epilithic diatom assemblages among the sewage ponds shows that assemblage composition reflects the nutrient gradient dictated by distance from the sewage outfall. This is in contrast to the chironomid assemblages that recorded largely similar species composition between sewage- affected and control ponds, as discussed in detail below.

Chironomid response to eutrophication and recovery The chironomid taxa that dominated the sewage ponds included Tanytarsus gracilentus-type and undifferentiated Tanytarsini species, the former being common in Arctic sediments (Brodersen et al.,

2004), and the latter being a large group made up of many ecological types (Brooks et al., 2007). We therefore refrain from over-interpreting the Tanytarsini group, as the group is too ecologically broad at this taxonomic resolution. Furthermore, there are cold-stenothermic taxa, which are commonly reported in oligotrophic waters, in both the sewage ponds as well as the control ponds (e.g. C. arctica-type,

Hydrobaenus/Oliveridia, M. hygropetricus-type, and Psectrocladius group). In fact, the chironomid species present in the sewage ponds (R-12, R-13) were broadly similar to those recorded in the oligotrophic control ponds, as was the case for the seabird-eutrophied ponds and reference sites at Cape

Vera (Stewart et al., 2013). Nonetheless, taxa typically considered eutrophic (e.g., Chironomus plumosus- type and Procladius) occurred at greater relative abundances in the sewage ponds, which also had much larger numbers of head capsules retrieved from the sediments, likely reflecting greater chironomid production as a result of greater food availability (Figure 2.8, 2.9). The similarities in species present between the sewage and control ponds demonstrate that nutrient concentrations are not directly affecting chironomid species assemblages. This is further corroborated by our observations that chironomid

assemblages in the sewage ponds showed no directional compositional changes during the periods when the trophic status of the ponds changed the most, in contrast to the diatom assemblages. Conversely, the

Cape Vera study ponds have nutrient concentrations that are consistently elevated over time because of the prolonged presence of the seabird colony, and thus both chironomids and diatom assemblages show no changes throughout the sediment record (Keatley et al., 2011; Stewart et al., 2013).

Chironomids feed on algal matter and detritus, and are thus undoubtedly indirectly affected by nutrients through their food source. Chironomids may respond to both increases in food quantity and quality. For example, chl-a concentrations in the sewage ponds are generally much higher than in the control sites (Figure 2.4, 2.5), meaning food availability is greater, which then gives rise to the greater abundance of chironomids in the sewage ponds compared to the control ponds. Changes in the type of food available to the chironomids also may have occurred, as indicated by changes in diatom assemblages over time (in both epilithic and sediment core samples) in most of the sewage ponds (Figure 2.6, 2.7).

Interestingly, the chironomid species assemblages in the sewage ponds show no directional assemblage shifts over periods of eutrophication, despite changes to both the availability and type of food source. This indicates that the primary control on species assemblage shifts for chironomids in these ponds is not nutrient concentrations.

In other studies, changes in lake trophic status have been linked to shifts in chironomid assemblages. In temperate regions, eutrophic lakes often record an abundance of Chironomus species, whereas oligotrophic lakes are typically characterised by Tanytarsus species (Brundin, 1949, 1956;

Thienemann, 1920, 1954). Thus, chironomids have been used quantitatively to infer changes in production-related variables including TP (Brooks et al., 2001) and chl-a (Brodersen & Lindegaard,

1999). Chironomids respond indirectly to nutrient increases through increased food availability, changes in habitat structure, and hypolimnetic oxygen depletion, and so species shifts will occur with changing nutrient concentrations (Brodersen & Quinlan, 2006). However, our data indicate that chironomid assemblages are not responding directly to changes in TP or other nutrients, but rather appear to be more strongly governed by elevated oxygen concentrations caused by continuous production of oxygen by 36

algae with access to 24 hours of sunlight during the Arctic summer and wind mixing of the shallow water column (Zmax < 1m). The only apparent response of chironomids to nutrients in our sewage ponds are the higher abundances of individuals that occur with greater food availability, as was also the case in the seabird-impacted sites of Cape Vera (Stewart et al., 2013).

Physiological and behavioural evidence shows that chironomid species respond to changes in oxygen concentrations through various mechanisms including the concentration of haemoglobin in the haemolymph (Czeczuga, 1960; Weber, 1980), the ability to switch to anaerobic metabolism (Hamburger et al., 1995), efficient osmotic and ionic regulation under anaerobia (Scholz & Zerbst-Boroffka, 1998), the ventilation of tube dwellings (Int Panis et al., 1995), and large body size (Heinis et al., 1994).

Accordingly, chironomids have been effectively used to infer past changes in hypolimnetic oxygen under eutrophication because of their strong relationships with oxygen depletion (Quinlan et al., 1998; Quinlan

& Smol, 2001b). Our data on these eutrophic but well-oxygenated ponds indicate that dominant chironomid taxa do not change with eutrophication if oxygen concentrations remain elevated.

Temperature is another factor that controls chironomid assemblage composition, as it governs the metabolic rate of chironomids (Eggermont & Heiri, 2012). In deep, cold, and eutrophic Meretta Lake (the terminal lake for North Base sewage disposal in Resolute), the cold climate prevented the proliferation of chironomids until the late 1970s, when anthropogenic climate warming reduced ice cover enough to allow the persistence of chironomids (Antoniades et al., 2011). This trend is not apparent in our small sewage ponds, nor is it apparent in our control ponds, as these ponds are so shallow that they thaw and warm relatively early in the growing season, and certainly much earlier than Meretta Lake (Douglas & Smol,

2000). Though the cold climate does not prevent chironomid production in the sewage ponds like it did in

Meretta Lake, the effects of temperature on chironomids are evident in both the sewage and control ponds. By the late 1980s, in both R-12 and R-13, C. arctica-type appears and increases in abundance until the present, as do several other taxa, and is reflected in the changes of the PCA sample scores from negative to positive values (Figure 2.8A, B). This is similar to the changes documented in the oligotrophic ponds of Cape Herschel, Ellesmere Island, where increases in Corynoneura taxa, as well as increases in 37

chironomid diversity, were attributed to the effects of recent climate warming (Quinlan et al., 2005). A marked directional shift in chironomid assemblage in our previous study of seabird-impacted ponds of

Cape Vera can also likely be attributed to recent climate warming because of similar chironomid changes in sites ~200 km apart, as well as the nature of the species involved (Stewart et al., 2013). Furthermore,

R-2, one of the oligotrophic control ponds, has recorded increasing abundances of head capsules concentrations in recent years, suggesting that greater ice-off periods have extended the growing season, allowing for greater numbers of chironomids. The evidence of recent climate warming in the Resolute

Bay ponds adds to the growing body of scientific literature documenting the effects of anthropogenic climate change in the circumpolar region (Smol et al., 2005; Smol & Douglas, 2007a, b).

Our Resolute Bay pond findings are consistent with our previous studies of chironomid assemblages in seabird-affected ponds at Cape Vera on Devon Island in the Canadian Arctic (Michelutti et al., 2011; Stewart et al., 2013). By using the shallow eutrophic ponds of Cape Vera, we showed, using a surface sediment approach, that so-called “oligotrophic” assemblages dominated eutrophic ponds, likely because of high oxygen levels. This was confirmed using detailed sediment core analysis, showing that chironomid assemblages were always characterized by so-called “oligotrophic” taxa in ponds that have clearly always been eutrophic, as the seabird colony has existed for the duration of the sediment record

(Stewart et al., 2013). Moving from naturally eutrophic seabird-impacted ponds to the culturally eutrophied ponds of Resolute, we show that the Cape Vera study was not an isolated phenomenon. The dominance of the so-called “oligotrophic” chironomid assemblages in the well-oxygenated sewage ponds demonstrates that nutrient levels do not directly affect chironomid assemblages.

Evidence of recent climate warming The influence of recent climate warming on the physical and biological limnology of the study ponds also seems apparent. Mean annual temperature records gathered by Environment Canada (2012) show an increase in the local temperature of Resolute Bay for approximately the past 20 years (Figure

2.3), which is consistent with recent warming in other regions of the Arctic (Smol et al., 2005). In our

study, the water chemistry and the chironomid data suggest that this recent temperature increase has had an effect on the chemical and biological properties of these ponds.

In both the control ponds (R-1 and R-2) and sewage ponds (R-10, R-11, R-12, and R-13), specific conductance and TN have increased in our monitoring data, which is likely due to recent warming that has caused increased evaporation rates, leading to the concentration of ions and nutrients in the water column. The phenomenon of increased solute concentrations due to evaporation, and even unprecedented drying of ponds, has been recently documented at Cape Herschel, Ellesmere Island (Smol & Douglas,

2007b). Contrasting our findings with those from nearby Char Lake, which is large and deep, illustrates that the small size of our study ponds makes them especially susceptible to increased evaporation due to warming, as larger lakes tend to have delayed responses to warming given their larger thermal inertia

(Michelutti et al., 2003). Finally, the increase in specific conductance and TN in our ponds is opposite to what was expected of the water chemistry in the sewage ponds, as we would have hypothesized continuing recovery from eutrophication. Our water chemistry monitoring shows the confounding effects of recent warming and furthermore illustrates the need to consider multiple stressors simultaneously.

The chironomids of the sewage and control ponds were responding to recent temperature increases in that head capsules per gram dry sediment (an indication of overall chironomid production) has tripled in the topmost centimeter of R-2 (Figure 2.8). This phenomenon has been observed in the control site of Cape Vera as well, and was attributed to recent warming allowing for a longer growing season, and therefore greater chironomid production (Stewart et al., 2013). Furthermore, the increase of

Corynoneura in the recent sediments of both sewage ponds is similar to the increases observed at Cape

Herschel, which was also attributed to recent climate warming (Quinlan et al., 2005). The appearance of

Corynoneura in the sediment cores of the sewage ponds coincides with the increases in specific conductance and TN that also suggest the effects of recent climate warming.

The recent subtle changes in both water chemistry and chironomid assemblage observed in the

Resolute ponds cannot be attributed to human-mediated eutrophication, as the changes occur in both the control ponds as well as the sewage ponds, and taxonomic changes have no linkage to previously claimed 39

nutrient preferences. We therefore conclude that the study ponds at Resolute Bay are responding to recent warming in a similar fashion to what has been documented elsewhere in the Arctic (Smol et al., 2005;

Smol & Douglas, 2007a, b).

Conclusions

Our Resolute pond dataset is a rare example of a long-term limnological monitoring program from the High Arctic that allows us to track and assess chemical and biological recovery from eutrophication. We used modern samples (water chemistry, surface sediments, and rock scrapes) from several years, as well as sediment cores, to track the changes in diatom and chironomid assemblages through time. The water chemistry data track chemical recovery from declining sewage inputs, and also record a nutrient gradient reflecting distance from the sewage outfall. The epilithic diatom assemblages closely tracked nutrient subsidies from sewage inputs and subsequent recovery over time. Furthermore, differences in diatom assemblages are evident between control and sewage-affected sites, as well as along a nutrient gradient in the sewage ponds, which are consistent with the known nutrient requirements of many of these taxa. In contrast, chironomid assemblages from sediment cores collected from the sewage ponds did not reflect the eutrophic nature of the ponds, nor did the assemblages change with decreasing nutrient concentrations. Rather, chironomid assemblages from sewage ponds resembled those of nearby control ponds, being dominated by chironomid taxa typically referred to as “oligotrophic.” Finally, chironomid assemblages showed evidence of the effects of recent climate warming in both the control ponds and sewage ponds, as has been documented elsewhere in circumpolar regions.

The lack of major differences in the chironomid assemblage composition between sewage ponds and control ponds indicates that nutrients are not directly affecting chironomid species, as they do diatoms. Whereas diatom assemblage composition has been shown to be an effective tool for quantitatively inferring total phosphorus (reviewed in Hall & Smol, 2010), chironomids have no direct physiological connection to nutrient concentrations. Instead, chironomids have complex responses to eutrophication as reflected through interacting factors such as oxygen concentration, habitat structure, and

food availability. Our findings show that chironomid survival and reproduction are more directly controlled by oxygen concentrations and temperature, which in turn may have strong relationships with nutrients.

Acknowledgements

This work was made possible because of the logistical and financial support provided by Natural

Science and Engineering Research Council (NSERC), Indian and Northern Affairs Canada (NSTP),

Natural Resources Canada, the Polar Continental Shelf Program (PCSP), and the W. Garfield Weston

Foundation. We would like to thank Xiaowa Wang and the National Laboratory for Environmental

Testing (NLET, Burlington, ON) for water chemistry analyses. We would also like to thank our colleagues at the Paleoecological Environmental Assessment and Research Laboratory (Queen’s

University, Kingston, Canada) for advice and support. Finally, we would like to thank two anonymous reviewers whose comments strengthened our manuscript.

References

Antoniades, D., M.S.V. Douglas & J.P. Smol, 2003. Comparative physical and chemical limnology of

two Canadian High Arctic regions: Alert (Ellesmere Island, NU) and Mould Bay (Prince Patrick

Island, NWT). Archiv für Hydrobiologie 158: 485-516.

Antoniades, D., P.B. Hamilton, M.S.V. Douglas & J.P. Smol, 2008. Diatoms of North America: the

freshwater floras of Prince Patrick, Ellef Ringnes, and Northern Ellesmere Islands from the

Canadian Arctic Archipelago. Iconographia Diatomologica, 17. A.R.G. Gantner/Verlag, Ruggell,

Liechtenstein.

Antoniades, D., N. Michelutti, R. Quinlan, J.M. Blais, S. Bonilla, M.S.V. Douglas, R. Pienitz, J.P. Smol

& W.F. Vincent, 2011. Cultural eutrophication, anoxia, and ecosystem recovery in Meretta Lake,

High Arctic Canada. Limnology and Oceanography 56: 639-65.

Appleby, P.G. & F. Oldfield, 1978. The calculation of 210Pb dates assuming a constant rate of supply of

unsupported 210Pb to the sediment. Catena 5: 1-8.

Battarbee, R.W., L. Carvalho, V.G. Jones, R.J. Flower, et al., 2001. Diatoms. In W.M. Last & J.P. Smol

(eds), Tracking environmental change using lake sediments. Volume 3: Terrestrial, Algal, and

Siliceous Indicators. Kluwer Academic Publishers, Dordrecht: 155-202.

Blais, J.M., L.E. Kimpe, D. McMahon, B.E. Keatley, M.L. Mallory, M.S.V. Douglas MSV & J.P. Smol,

2005. Arctic seabirds transport marine-derived contaminants. Science 309: 445.

Brodersen, K.P. & C. Lindegaard 1999. Classification, assessment and trophic reconstruction of Danish

lakes using chironomids. Freshwater Biology 42: 143-157.

Brodersen, K.P. & R. Quinlan, 2006. Midges as palaeoindicators of lake productivity. Quaternary Science

Reviews 25: 1995-2012.

Brodersen, K.P., O. Pedersen, C. Lindegaard & K. Hambuger, 2004. Chironomids (Diptera) and oxy-

regulatory capacity: An experimental approach to paleolimnological interpretation. Limnology

and Oceanography 49: 1549-1559.

Brooks, S.J., H. Bennion & H.J.B. Birks, 2001. Tracing lake trophic history with a chironomid total

phosphorus inference model. Freshwater Biology 46: 513-533.

Brooks, S.J., P.G. Langdon & O. Heiri, 2007. The identification and use of PalaeArctic Chironomidae

larvae in palaeoecology, QRA Technical Guide No. 10. Quaternary Research Association,

London.

Brundin, L., 1949. Chironomiden und andere Bodentiere der südschwedischen urgebirgsseen.

Report/Institute of Fresh-water Research Drottningholm 30: 1-914.

Brundin, L.,1956. Die bodenfaunistischen Seetypen un ihre Anwenbarkeit auf die Südhalbkugel. Zugleich

eine Theorie der produktionsbiologischen Bedeutung der glazialen Erosion. Report/Institute of

Freshwater Research Drottningholm 37: 186-235.

Czeczuga, B., 1960. Haemoglobin content of the larvae of Tendipes f l. plumosus L. from various levels

of bed sediment. Nature 186: 484.

Douglas, M.S.V. & J.P. Smol, 1994. Limnology of high arctic ponds (Cape Herschel, Ellesmere Island,

N.W.T.). Archiv für Hydrobiologie 131: 401-434. 42

Douglas, M.S.V. & J.P. Smol, 2000. Eutrophication and recovery in the High Arctic: Meretta Lake

(Cornwallis Island, Nunavut, Canada) revisited. Hydrobiologia 431: 193-204.

Eggermont, H. & O. Heiri, 2012. The chironomid-temperature relationship: expression in nature and

palaeoenvironmental implications. Biological Reviews 87: 430-456.

Environment Canada, 1979. Analytical Methods Manual. Water Quality Branch, Inland Water Directorate

Ottawa, Ontario. K1A 0E7.

Environment Canada, 2012. Canadian Climate Normals 1971-2000 [Internet]. [Accessed 2012 May 16].

Available from

http://www.climate.weatheroffice.gc.ca/climate_normals/results_e.html?stnID=1776&autofwd=1

Glew, J.R., 1988. A portable extruding device for close interval sectioning of unconsolidated core

samples. Journal of Paleolimnology 1: 235-239.

Hall, R.I. & J.P. Smol, 2010. Diatoms as indicators of lake eutrophication. In J.P. Smol & E.F. Stoermer

(eds), The Diatoms: Applications for the Environmental and Earth Sciences. Cambridge

University Press, Cambridge: 122-151.

Hamburger, K., P.C. Dall & C. Lindegaard, 1995. Energy metabolism of Chironomus anthracinus

(Diptera, Chironomidae) under laboratory and field conditions. Hydrobiologia 297:187-200.

Heinis, F., J-P. Sweerts & E. Loopie, 1994. Micro-environment of chironomids larvae in the littoral and

profundal zones of Lake Maarsseveen I, The Netherlands. Fundamental Applied Limnology 124:

173-289.

Int Panis, L., B. Godeeris & R.F. Verheyen, 1995. On the relationship between oxygen microstratification

in a pond and the spatial distribution of the benthic chironomid fauna. In P. Cranston (ed)

Chironomids: From Genes to Ecosystems. CSIRO Publications, Melbourne: 323-328.

Keatley, B. E., Douglas, M. S. V., Blais, J. M., Mallory, M. L. & J.P. Smol, 2009: Impacts of seabird-

derived nutrients on water quality and diatom assemblages from Cape Vera, Devon Island,

Canadian High Arctic. Hydrobiologia 61: 191– 205.

Keatley, B.E., J.M. Blais, M.S.V. Douglas, I. Gregory-Eaves, M.L. Mallory, N. Michelutti & J.P. Smol,

2011. Historical seabird population dynamics and their effects on Arctic pond ecosystems: a

multi-proxy paleolimnological study from Cape Vera, Devon Island, Arctic Canada. Fundamental

Applied Limnology 179: 51-66.

Michelutti, N., M.S.V. Douglas & J.P. Smol, 2002. Tracking recent recovery from eutrophication in a

High-Arctic lake (Meretta Lake, Cornwallis Island, Nunavut, Canada) using fossil diatom

assemblages. Journal of Paleolimnology 28: 377-381.

Michelutti, N., M.S.V. Douglas & J.P. Smol, 2003. Diatom response to recent climatic change in a high

arctic lake (Char Lake, Cornwallis Island, Nunavut). Global and Planetary Change 38: 257-271.

Michelutti, N., M.H. Hermanson, J.P. Smol, P.J. Dillon & M.S.V. Douglas, 2007a. Delayed response of

diatom assemblages to sewage inputs in an Arctic lake. Aquatic Science 69: 523-533.

Michelutti, N., M.S.V. Douglas & J.P. Smol, 2007b. Evaluating diatom community composition in the

absence of marked limnological gradients in the high Arctic: a surface sediment calibration set

from Cornwallis Island (Nunavut, Canada). Polar Biology 30: 1459-1473.

Michelutti, N., B.E. Keatley, S. Brimble, J.M. Blais, H. Liu, M.S.V. Douglas, M.L. Mallory, R.W.

Macdonald & J.P. Smol, 2009. Seabird-driven shifts in Arctic pond ecosystems. Proceedings of

the Royal Society B Biological Sciences 276: 591-596.

Michelutti, N., M.L. Mallory, J.M. Blais, M.S.V. Douglas & J.P. Smol, 2011. Chironomid assemblages

from seabird-affected high Arctic ponds. Polar Biology 34: 799-812.

Quinlan, R., J.P. Smol & R.I. Hall, 1998. Quantitative inferences of past hypolimnetic anoxia in south-

central Ontario lakes using fossil midges (Diptera: Chironomidae). Canadian Journal of Fisheries

and Aquatic Sciences 55: 587-596.

Quinlan, R. & J.P. Smol, 2001a. Setting minimum head capsule abundance and taxa deletion criteria in

chironomid-based inference models. Journal of Paleolimnology 26: 327-342.

Quinlan, R. & J.P. Smol, 2001b. Chironomid-based inference models for estimating end-of-summer

hypolimnetic oxygen from south-central Ontario lakes. Freshwater Biology 46: 1529-1551. 44

Quinlan, R., M.S.V. Douglas & J.P. Smol, 2005. Food web changes in arctic ecosystems related to

climate warming. Global Change Biology 11: 1381-1386.

Schindler, D.W., J. Kalff, H.E. Welch, G.J. Brunskill, H. Kling & N. Kritsch, 1974. Eutrophication in the

High Arctic — Meretta Lake, Cornwallis Island (75° N Lat.). Journal of Fisheries Research

Board of Canada 31: 647-662.

Scholz, F. & I. Zerbst-Boroffka, 1998. Environmental hypoxia affects osmotic and ionic regulation in

freshwater midge-larvae. Journal of Insect Physiology 44: 427-436.

Smol, J.P., 2010. The power of the past: Using sediments to track the effects of multiple stressors on lake

ecosystems. Freshwater Biology 55 (Suppl. 1): 43-59.

Smol, J.P., A.P. Wolfe, H.J.B. Birks, M.S.V. Douglas, V.J. Jones, A. Korhola, R. Pienitz, K. Rühland, S.

Sorvari, D. Antoniades, S.J. Brooks, M. Fallu, M. Hughes, B.E. Keatley, T.E. Laing, N.

Michelutti, L. Nazarova, N. Nyman, A.M. Paterson, B. Perren, R. Quinlan, M. Rautio, E.

Saulnier-Talbot, S. Siitonen, N. Solovieva & J. Weckstrom, 2005. Climate driven regime shifts in

the biological communities of arctic lakes. Proceedings of the National Academy of Sciences

USA 102: 4397-4402.

Smol, J.P. & M.S.V. Douglas, 2007a. From controversy to consensus: making the case for recent climatic

change in the Arctic using lake sediments. Frontiers in Ecology and the Environment 5: 466-474.

Smol, J.P. & M.S.V. Douglas, 2007b. Crossing the final ecological threshold in high Arctic ponds.

Proceedings of the National Academy of Sciences 104: 12395-12397.

Spaulding, S. & M. Edlund, 2009. Fistulifera. In: Diatoms of the United States. Retrieved October 12,

2012, from http://westerndiatoms.colorado.edu/taxa/genus/Fistulifera

Stewart, E.M., N. Michelutti, J.M. Blais, M.L. Mallory, M.S.V. Douglas & J.P. Smol, 2013. Contrasting

the effects of climate, nutrient, and oxygen dynamics on subfossil chironomid assemblages: a

paleolimnological experiment from eutrophic High Arctic ponds. Journal of Paleolimnology 49:

205-219.

Thienemann, A., 1920. Untersuchungen über die Bziehungen zwischen dem Sauerstoffgehaly des

Wassers und der zusammensetzung der Fauna in norddeutschen Seen. Archiv für Hydrobiologie

12: 1-65.

Thienemann, A., 1954. Chironomus. Leben, Verbreitung und wirtschaftliche Bedeutung der

Chironomiden. Binnengewӓsser 20: 1-834.

Walker, I.R., 2001. Midges: Chironomidae and related Diptera. In: J.P. Smol, H.J.B. Birks & W.M. Last

(eds), Tracking environmental change using lake sediments vol. 4, zoological indicators. Kluwer

Academic Publishers, Dordrecht: 43-66.

Weber, R.E., 1980. Functions of invertebrate haemoglobins with special reference to adaptations to

environmental hypoxia. American Zoologist 20: 79-101.

Wolfe, A.P., G.F. Miller, C.A. Olsen, S.L. Forman, P.T. Doran & S.U. Holmgren, 2004. Geochronology

of high latitude lake sediments. In: R. Pienitz, M.S.V. Douglas & J.P. Smol (eds), Developments

in Paleoenvironmental Research vol. 8, Long-term environmental change in the Arctic and

Antarctic lakes. Springer, Dordrecht: 19–52.

Figure 2.1 Map showing the location of Resolute Bay on Cornwallis Island (marked with a star), with an inset of Meretta Lake, sewage ponds (R-10, R-11, R-12, R-13), and control ponds (R-1, R-2). The solid line shows the water course that raw sewage travelled through the sewage ponds to Meretta Lake from 1949 to 1979. The dashed line shows the second water course that carried sewage directly to Meretta Lake from 1949-1998. Modified from Google maps.

Figure 2.2 Photographs of the sewage ponds near the Ministry of Transport base (i.e. airport facilitates) in the hamlet of Resolute on Cornwallis Island, Nunavut. Closest to the base is R-13 (middle right), and moving south along the watercourse towards Meretta Lake the ponds are: R-12 (middle left), R-11 (top right), and R-10 (top left). The reference sites are R-1 (bottom left) and R-2 (bottom right). Photographs are courtesy of Christopher Grooms.

-10 y = 0.0284x - 72.383

(ᵒC) R² = 0.1866 -12

-14

-16

-18 MeanTemperature -20

Year AD

Figure 2.3 Mean annual temperature for Resolute Bay, Nunavut, from 1948 to 2010. Data taken from Environment Canada (2012).

Figure 2.4 Limnological variables over time for the control ponds, R-1 in grey and R-2 in black. (A) Chlorophyll-a (Chl-a) in µg/L (note: no value was available for R-2 in 1999), (B) Specific conductance in µS/cm, (C) Total filtered nitrogen (TN) in mg/L, and (D) Total unfiltered phosphorus (TP-u) in µg/L.

Figure 2.5 Limnological variables over time for the sewage ponds, R-10 (solid black line), R-11 (dashed black line), R-12 (solid grey line), and R- 13 (dashed grey line). (A) Chlorophyll-a (Chl-a) in µg/L, (B) Specific conductance in µS/cm, (C) Total filtered nitrogen (TN) in mg/L, and (D) Total unfiltered phosphorus (TP-u) in µg/L.

Figure 2.6 The epilithic diatom assemblages for all sewage ponds (A) R-10, (B) R-11, (C) R-12, and (D) R-13 for several sampling years between 1993 and 2011.

Figure 2.7 The diatom species assemblage changes of the dominant taxa from the sediment cores for sewage ponds R-12 (A) and R-13 (B), as well as for the control ponds R-1 (C) and R-2 (D), presented as percent relative abundance (%). 210Pb dates in years AD are given to the left of the R-12, R-13, and R-2 stratigraphies. Sediments from R-1 had insufficient 210Pb activities to generate dates (see Supplemental Figure 2.1). Principal Component Analysis sample scores from axis 1 are plotted on the right-hand side of each stratigraphy (PCA axis 1).

Figure 2.8 The chironomid assemblage changes of the dominant taxa from the sewage pond cores R-12 (A) and R-13 (B), as well as for the control pond R-2 (D). Data for the other control pond R-1 (C) is presented as number of head capsules per gram of dry sediment (HC per g DM) due to the paucity of remains. For all ponds, the number of head capsules per gram of dry sediment is shown on the rightmost side of each stratigraphy (HC per g DM). 210Pb dates in years AD are given to the left of the R-12, R-13, and R-2 stratigraphies. Sediments from R-1 had insufficient 210Pb activities to generate dates (see Supplemental Figure 2.1). Principal Component Analysis sample scores from axis 1 are plotted on the right-hand side of each stratigraphy (PCA axis 1).

Figure 2.9 (A) Principal Component Analysis biplot showing sediment core intervals (points) and chironomid species (arrows) for ponds R-2 (circle), R-12 (star), and R-13 (square). R-1 was not included in the analysis due to low counts of head capsules. Detrended Correspondence Analysis plot showing (B) diatom species distribution and (C) sediment core intervals from ponds R-1 (triangle), R-2, R-12, and R- 13 (represented using the same symbols as above). Species abbreviations are given in Supplemental Table 2.2.

Supplemental Figure 2.1 210Pb, 214Bi, and 137Cs radioactivities in decays per minute per gram (dpm/g) for (A) R-12, (B) R-13, (C) R-1, and (D) R-2 sediment cores.

Supplemental Table 2.1 Chironomid species presence (X) or absence in the surface sediments of the years listed at the top of the table for the sewage ponds near Resolute Bay, Cornwallis Island, Nunavut. R-10 1993 2002 2008 2011 Corynoneura arctica-type X X X Hydrobaenus/Oliveridia X Metriocnemus hygropetricus-type X X Psectrocladius group X Tanytarsus gracilentus-type X X Tanytarsini spp. X X X Chironomus plumosus-type X X Procladius X Total number of species 5 4 2 4 Total # head capsules counted 7 8 2 6 R-11 1993 2006 2008 2009 2011 Corynoneura arctica-type X X X Hydrobaenus/Oliveridia X Metriocnemus hygropetricus-type X X X X Psectrocladius group X X X Tanytarsus gracilentus-type X X X X X Tanytarsini spp. X X X X X Chironomus plumosus-type X X X Procladius X Total number of species 2 7 6 3 7 Total # head capsules counted 5 32.5 70 6 126.5 R-12 1993 2002 2006 2008 2009 2011 Corynoneura arctica-type X X X X X Hydrobaenus/Oliveridia X X Metriocnemus hygropetricus-type X X X X X X Psectrocladius group X X X X Tanytarsus gracilentus-type X X X X X Tanytarsini spp. X X X X X X Chironomus plumosus-type X X X X Procladius X X Total number of species 7 5 5 7 7 3 Total # head capsules counted 79 17.5 31 100.5 123.5 7 R-13 2006 2008 2009 2011 Corynoneura arctica-type X X X X Limnophyes/Paralimnophyes X Metriocnemus hygropetricus-type X X X X Psectrocladius group X X Tanytarsus gracilentus-type X X X X Tanytarsini spp. X X X X Chironomus plumosus-type X X X X Procladius X X X Total number of species 6 6 8 6 Total # head capsules counted 28 154.5 103.5 29

Supplemental Table 2.2 Species abbreviations used in PCA and DCA of Figure 2.9 listed alphabetically by proxy. Chironomids Diatoms Abbreviation Species Name Abbreviation Species Name ChirPlum Chironomus Plumosus-type AchnMint Achnanthidium minutissimum CornArct Corynoneura arctica-type AmphCopl Amphora copulate Eukieffe Eukiefferiella/Tvetenia CalnSilc Caloneis silicula Hydrobae Hydrobaenus/Oliveridia CyclCfSt Cyclotella cf striata (pinched) Limnophy Limnophyes/Paralimnophyes CyclMeng Cyclotella menighiniana MetrHygr Metriocnemus hygropetricus-type CyclStri Cyclotella striata Procladi Procladius CymAngSp Cymbelopleura angustata var spitsbergensis PsecGrou Psectrocladius group CymbBotl Cymbella botellus TantGrac Tanytarsus gracilentus-type CymbCfMc Cymbella cf microcephala TanytSpp Tanytarsini spp CymbClev Cymbella cleve-eulerai CymbDelc Cymbella delicaticula CymbDesg Cymbella designate CymbMicr Cymbella microcephala CymIncSp Cymbelopleura incerta var spitsbergensis DiatMonl Diatoma moniliformis EncnCest Encynopsis cesatii EucoFlex Eucocconeis flexella EucoLept Eucocconeis leptostriata FragTenr Fragilaria tenera NavcChia Navicula chiarae NavcMins Navicula minuscule NavcPhyl Navicula phyllepta NavcVulp Navicula vulpine NitzAlpn Nitzschia alpine NitzPerm Nitzschia perminuta Sp1(cf Sp 1 (cf Nedium affine) StepMint Stephanodiscus minutulus

Supplemental Figure 2.2 Dissolved oxygen concentration (mg/L) and saturation (%) for sewage ponds, R-12 and R-13. Measurements were made hourly in July 2014 using a HOBO DO logger (model U26- 001).

Chapter 3

Bird-mediated eutrophication and enrichment of heavy metals in shallow

ponds on islands in Lake Ontario

Abstract

Ring-billed Gulls (Larus delawarensis) and Double-crested Cormorants (Phalacrocorax auritus) in the North American Great Lakes have experienced rapid population increases during the early- and mid-20th century (respectively), such that densely populated nesting sites are a common sight throughout the region. Sediment cores from ponds located near nesting sites have been successfully used to track the presence/absence of these large colonies on islands in eastern Lake Ontario, as well as the marked fertilizing effect that guano has on aquatic primary production. However, less is known about bird- mediated biotransport of contaminants to these nesting sites and the effect it has on mid-trophic levels, such as benthic invertebrates. Using a comparative limnological approach, we found that modern water chemistry from bird-impacted ponds had concentrations of Al, Cd, Cu, Fe, Pb, and Se that exceeded the

Canadian guidelines for the protection of aquatic life and were elevated compared to the non-impacted reference ponds. Ponds surrounded by Ring-billed Gulls were enriched in a greater number of trace elements compared to those impacted by mainly cormorants, possibly reflecting the diverse feeding strategy of gulls compared to the strictly piscivorous diet of cormorants. Subfossil chironomid assemblages may have also reflected metal pollution, as species in the bird-impacted sites included many taxa tolerant of heavy metal contamination, and chironomids were largely absent from our most contaminated site. Previous work on diatoms from the bird-impacted ponds likewise recorded assemblages dominated by species tolerant of metal pollution. These findings have implications for understanding and managing contaminant cycling in the Great Lakes, as waterbird biovectors that nest in dense colonies may constitute the primary pathway for the movement of metals to new and unexpected locations, and possibly, in concentrations that may affect aquatic biota.

Keywords: Lake Ontario, ponds, paleolimnology, heavy metals, chironomids, Double-crested

Cormorants, Ring-billed Gulls

Introduction

Paleolimnology is being increasingly employed to track the history and impacts of large bird populations, especially where monitoring records are absent (Blais et al. 2005; Michelutti et al. 2010;

Luoto and Brooks 2014). This type of work has determined the century-scale history of remote seabird colonies (Michelutti et al. 2009), demonstrating the large fertilizing effect of guano on nearby freshwaters

(Keatley et al. 2009; Stewart et al. 2013; 2015) and the movement of contaminants from marine food webs to terrestrial nesting sites (Blais et al. 2005). Contaminants that bioaccumulate and biomagnify in upper trophic level bird species are concentrated in the nesting area via bird waste products, such as guano, feathers, and carcasses. Paleolimnological methods have also established that different species of birds deliver various mixtures of contaminants based on their trophic feeding position (Michelutti et al.

2010), illustrating the importance of food source to contaminant loads.

Bird colonies in Lake Ontario have recently generated paleolimnological interest (Stewart et al.

2015; Hargan et al. in review; Appendix B), as several species have rapidly increased in population over the past several decades, leading to the prevalence of large colonies throughout the Great Lakes Region.

Ring-billed Gulls (Larus delawarensis) and Double-crested Cormorants (Phalacrocorax auritus, hereafter

“cormorants”) are two species that have experienced amongst the largest population increases. Ring- billed Gull populations increased as a result of urban expansion (using garbage dumps for food) after

WWII (Ludwig 1974), making them now the most numerous waterbird species in the Great Lakes (Morris et al. 2011). Double-crested Cormorants became a symbol of Great Lakes water quality awareness among the scientific community in the 1960s and 70s with the discovery of contaminant-related deformities in adult birds and their young (Ludwig 2013), which resulted in a major population crash (Weseloh et al.

2002). After the ban of DDT (dichlorodiphenyltrichloroethane) in 1973, cormorant populations in the

Great Lakes rapidly rebounded, and then continued to rise in part due to increases in invasive fish food

sources (Weseloh et al. 1995; Somers et al. 2003). This led to public and scientific controversy about large colonies that resulted in the dramatic alteration of nesting sites (Boutin et al. 2011) and that cause potential threats to other co-habiting species such as the Black-crowned Night Heron (Nycticorax nycticorax) (OMNR 2011; Rush et al. 2015).

Paleolimnological techniques have recently been used to track the impacts of Ring-billed Gulls and cormorants on nesting islands in the Great Lakes Region, showing that, in many locations in Lake

Ontario, large colonies of these species are unprecedented in the past ~150 years (Stewart et al. 2015;

Hargan et al. in review; Appendix B). For example, some islands that currently host large colonies of cormorants were never-before impacted by large waterbird colonies, according to the multi-century sediment record, whereas some islands have likely long-sustained large populations of waterbirds

(Stewart et al. 2015; Hargan et al. in review; Appendix B). Geochemical indicators (stable nitrogen isotopes, chlorophyll a concentrations) and biological proxies (diatoms) in the pond sediments near nesting areas demonstrated striking changes associated with eutrophication due to large subsidies of nutrient-rich guano (Stewart et al. 2015; Appendix B). Furthermore, the recent application of sedimentary sterol-based biomarkers directly linked the presence of birds with changes in sediment records, providing unequivocal evidence of the bird impacts (Hargan et al. in review).

Here, we investigate the impacts of large bird colonies on water chemistry, and particularly heavy metal and trace element concentrations, as well as sedimentary benthic invertebrate (chironomid, or non- biting midge) assemblages, across sites with various degrees of impact from multiple bird species, including Ring-billed Gulls and cormorants. The objectives of this study are to: 1) determine possible bird-mediated pollution of these ponds by comparing water and sediment chemistry between bird- impacted and non-impacted sites, as well as with the freshwater quality guidelines for the protection of aquatic life set by the Canadian Council of Ministers of the Environment (CCME 1999); and 2) assess sediment profiles for changes in subfossil chironomid assemblages in response to eutrophication and possible metal contamination, using water chemistry and hourly dissolved oxygen profiles to disentangle the effects of multiple environmental parameters on species distributions. 62

We determined the onset of eutrophication at bird-impacted sites by using down-core profiles of sterol biomarkers, stable nitrogen isotopes (δ15N), and sedimentary concentrations of chlorophyll-a previously published for these sites (Stewart et al. 2015; Hargan et al. in review; Appendix B). Under scenarios of eutrophication, chironomids may respond to changes in habitat (Brodersen et al. 2001), food abundance (Stewart et al. 2013; 2014), or oxygen conditions (Little and Smol 2001; Luoto and Salonen

2010). Some genera of chironomids are also known for being particularly tolerant to pollution (e.g.

Chironomus), especially metals from, for example, mining operations (Ilyashuk et al. 2003). Thus, the combination of modern water chemistry analysis compared with water quality guidelines and downcore chironomid analysis, as well as with previous studies of algae at these sites (Stewart et al. 2015; Appendix

B), may provide further insight into the ecological impacts of large waterbird colonies on these freshwater systems, as well as contaminant cycling in the Great Lakes Region.

Site description

Five eutrophic ponds on islands in eastern Lake Ontario (East Brother Island, Pigeon Island,

Little Galloo Island, High Bluff Island, and Gull Island) were sampled because of the large colonies of

Double-crested Cormorants (Phalacrocorax auritus) and Ring-billed Gulls (Larus delawarensis) that nest there from spring to fall (Figure 3.1). In addition, three other ponds on two islands with no current waterbird colonies (Main Duck Island and Calf Island) were also sampled as non-impacted sites that are naturally eutrophic (Figure 3.1). Pond characteristics, including dimensions and estimated overall bird numbers, are summarized in Table 3.1.

Bird-impacted ponds East Brother Island (44°12’18.43”N, 76°37’28.33”W) is located near Kingston (Ontario) and has been home to a dense colony of Double-crested Cormorants since at least 2001 (Stewart et al. 2015 supplemental material), with current total nest numbers (including some Ring-billed Gulls) equalling

~1500. East Brother Pond is likely impacted by the majority of this colony (predominantly cormorants), as the pond takes up much of the island and is surrounded by cormorant nests. Likewise, nearby Pigeon 63

Island (44°03’59.41”N, 76°32’51.51”W) is further from the northeastern shore of Lake Ontario than East

Brother and currently hosts a breeding colony of ~2100 cormorant nests. In addition, ~5000 Ring-billed

Gulls nested on Pigeon Island until ~1990, after which time they left (Weir 2008). The Pigeon Island pond is ephemeral and typically surrounded by ground-nesting cormorants. Little Galloo Island

(43°53’08.78”N, 76°23’43.93”W) is located on the U.S. side of eastern Lake Ontario near Henderson

Harbor, NY. There were ~2200 cormorant nests on the island in 2015, along with a colony of ~43,000

Ring-billed Gulls (<5% Herring Gulls, Larus argentatus) and ~2000 Caspian Terns (Hydroprogne caspia) (data from New York State Department of Environmental Conservation (NYSDEC), see

Appendix B).

High Bluff Island (43°58’32.03” N, 77°44’47.74” W) and Gull Island (43°59’00.56” N,

77°44’21.43” W) are located close to the northern shore of Lake Ontario in Presqu’ile Provincial Park

(approximately mid-way between Kingston and Toronto, ON). Both islands were colonized by cormorants in 1986, though only High Bluff Island supported cormorants from 2011 onwards (OMNR

2011). Gull Island has supported a colony of ~40,000 Ring-billed Gulls for over 40 years, which can reach upwards of 100,000 birds during the height of breeding season (D. Tyerman, OMNR, pers. commun. 2016). High Bluff Pond is impacted by a portion of the cormorant colony, at times as many as

4000 nests and as little as 1000 nests (D. Tyerman, OMNR, pers. commun. 2016), and Gull Pond is likely impacted by the majority of the Ring-billed Gulls on the island. Only sparse vegetation thrives on all of these bird-impacted islands because of the toxic concentrations of guano deposited by the large colonies of birds during the breeding season.

Non-impacted ponds Main Duck Island (43°55’28.39”N, 76°36’51.34”W) is located approximately half way between the Canadian and US shores of Lake Ontario south of Kingston, Ontario. This island has two large ponds in marshy and wooded areas that are not impacted by large breeding colonies, Main Duck Pond 1 and

Main Duck Pond 2. The third non-impacted site is unofficially named Calf Pond on Calf Island

(43°52’38.92”N, 76°22’02.76”W), close to Little Galloo Island on the U.S. side of Lake Ontario. The

NYSDEC actively manages waterbird populations on Calf Island. In 2008, ~170 cormorants nested on

Calf Island (data from NYSDEC) but were subsequently removed the following year.

Methods

Water chemistry Approximately 1L of pond water was collected from each pond for our analyses of lakewater nutrients, trace elements, and major ions at the National Laboratory for Environmental Testing

(Burlington, ON, Canada) using their standard procedures (Environment Canada 1994a; 1994b). A calibrated Hanna Instruments pH/EC/TDS/Temperature meter (model HI98129) was used to measure specific conductance, temperature, and pH on site. Samples and measurements were taken from East

Brother on June 24th 2013, September 11th 2013, and July 28th 2016; from Pigeon on June 16th 2011

(however, we were unable to collect water due to extremely shallow depth); from Little Galloo on May

23rd 2014; from High Bluff and Gull on April 24th 2014; from Main Duck Pond 1 on September 11th 2013; from Main Duck Pond 2 on September 11th 2013, July 27th 2016, and August 17th 2016; and from Calf on

June 19th 2014, July 21st 2016, and September 22nd 2016. Basic water chemistry data (not including trace elements) were first published for East Brother, Pigeon, and Main Duck Pond 2 in Stewart et al. (2015), but are presented here for comparison with newer samples and trace element concentrations. Water chemistry data were plotted using Sigma Plot 10, and raw data are given as supplemental material

(Supplemental Table 3.1, 3.2).

Dissolved oxygen, specific conductance, temperature, and pH were also measured every hour in

East Brother Pond using a YSI 6-Series sonde logger from September 19th to 25th 2013, and Ecowatch

Lite software version 1.0 was used to calculate dissolved oxygen saturation (%) with an accuracy of ± 2% of the reading. Dissolved oxygen and temperature were also measured using an Onset® HOBO DO logger

(model U26-001) from July 27th to August 17th 2016 in Main Duck Pond 2, and from July 21st to

September 22nd in Calf Pond. Measured dissolved oxygen concentrations (± 0.2 mg/L) from Main Duck

Pond 2 and Calf Pond were processed to give dissolved oxygen saturation (%) using the “Dissolved

Oxygen Assistant” function in the HOBOware program version 3.7.8. Dissolved oxygen profiles were plotted using Sigma Plot 10. Finally, dissolved oxygen measurements from our Lake Ontario sites were compared with measurements taken in July 2014 using the same methods from two eutrophic shallow ponds (Supplemental Figure 3.3) near Resolute Bay (74°41'50.28", 094°49'46.92") on Cornwallis Island

(NU) in the Canadian High Arctic, both of which were also studied for changes in sedimentary chironomid assemblages (Stewart et al. 2014).

Sediment sampling and analysis Sediment cores were taken using a Glew and Smol (2016) shallow water push corer, and sectioned onsite using a Glew (1988) extruder into 0.5-cm intervals. All sediments were processed for biological indicators at Queen’s University. Sediment subsamples from each core were freeze-dried and dated using an Ortec® high purity germanium gamma spectrometer (Oak Ridge, TN, USA). Excess 210Pb activities were modeled into chronologies using the Constant Rate of Supply (CRS) model (Appleby

2001) with ScienTissiME software version 2.0.1 (Barry’s Bay, ON, Canada). Cores used in our chironomid analysis have previously published 210Pb chronologies, with East Brother, Pigeon, and Main

Duck Pond 2 detailed in Stewart et al. (2015), and Little Galloo and Calf in Hargan et al. (in review) - the latter two of which are also presented in Appendix A. Sedimentary metal and trace element concentrations (dry weight) were measured in the cores from High Bluff, Gull, Little Galloo, and Calf using standard ICP-MS procedures at SGS Laboratories (Ottawa, Ontario), and data are presented for surface sediments only as supplementary material (Supplemental Figure 3.1).

Chironomid head capsule remains were isolated from sediments of East Brother, Little Galloo,

Pigeon, Main Duck Pond 2, and Calf using the standard methods of Walker (2001). Insufficient head capsules in sediment cores from High Bluff and Gull ponds prevented chironomid analysis, and Main

Duck Pond 1 was not analysed for chironomids due to a lack of a reliable core chronology.

Approximately 1-6 g (depending on head capsule abundance) of wet sediment from each interval was

processed in 80 mL of 5% KOH solution at ~70-80ºC for 20-30 minutes. This sediment-KOH solution was then rinsed through a 100-µm sieve using double deionized water into a small beaker. Sediments were picked through for a minimum of 50 head capsules when possible (Quinlan and Smol 2001), which were then mounted permanently on slides using Entellan®. Chironomid head capsules were identified to species type or genus when possible using predominantly the taxonomy of Brooks et al. (2007) and

Andersen et al. (2013), as well as the online resources of Walker (2007) for “Podonominae Species 1.”

Relative abundances of all taxa reaching 5% of the total assemblage in at least one interval were plotted as stratigraphies using C2 software version 1.7.

Statistical analysis Chironomid relative abundance data were compared and summarized for 4 of the 5 sites analysed using Principal Component Analysis (PCA) performed in CANOCO 5 with the CANOCO advisor tool.

Raw species counts were transformed using a Hellinger transformation (Legendre and Birks 2012).

Pigeon was excluded from statistical analyses because chironomid remains retrieved from the core were insufficient to calculate representative relative abundances (Quinlan and Smol 2001). For comparison purposes, chironomid species data from three High Arctic sites from Stewart et al. (2014) were presented here using a PCA to compare with species distributions in our temperate sites (Supplemental Figure 3.3).

Results

Water chemistry

Total dissolved nitrogen (TN-f), as well as nitrate plus nitrite (NO2 + NO3) concentrations, were highest in High Bluff and Gull ponds in April of 2014 compared to all sampling events from all bird- impacted and non-impacted sites (Figure 3.2). Interestingly, however, ammonia (NH3) concentrations were higher in East Brother in June 2013 than any other sample (Figure 3.2). All bird-impacted sites had higher nitrogen levels than non-impacted ponds. Total phosphorus (TP) was also greatly elevated in all samples from bird-impacted sites compared to non-impacted sites, with all bird-impacted measurements surpassing 2000 µg/L, except for those taken from East Brother in September 2013 and July 2016 (Figure 67

3.2). However, the non-impacted site, Main Duck Pond 1, was eutrophic in Sept 2013 by North American

Standards (> 30 µg/L, OME 2010), with total unfiltered phosphorus (TP-u) equal to 36.4 µg/L (Figure

3.2). Main Duck Pond 2 and Calf Pond were hypereutrophic by the same standards for all three sampling events with TP-u concentrations of 86.8, 52.8, and 53.9 µg/L in Main Duck Pond 2, and 151, 333, and

449 µg/L in Calf. Chlorophyll-a (chl-a) concentrations were highest in Little Galloo Pond and Gull Pond, and all other sites (regardless of bird impact) had water chlorophyll-a concentrations that were orders of magnitude lower (Figure 3.2).

Pond water pH values were comparable across all sites and sampling dates for both bird-impacted ponds and non-impacted ponds, and both Main Duck ponds had the lowest pH values in September 2013

(Figure 3.2). Specific conductance was elevated in the bird-impacted sites, with all measurements exceeding ~500 µS/cm, except in Little Galloo (Figure 3.2). Non-impacted site, Calf Pond, also had a specific conductance approaching 500 µS/cm in September 2016, when water levels were reduced from

~1 m to 10 cm due to a hot and dry summer. Major ion concentrations were also elevated, especially potassium and sulfate ion concentrations, in all bird-impacted sites, except for Little Galloo, which while still elevated, was the least enriched compared to non-impacted sites (Figure 3.2).

Trace element concentrations in water Many trace metals and elements were elevated in bird-impacted sites compared to non-impacted sites (Figure 3.3, Supplemental Table 3.2). Furthermore, many element concentrations exceeded or approached the freshwater quality guidelines for the protection of aquatic life for chronic effects (CCME

1999). Cd concentrations were 0.24 µg/L in Gull Pond in April 2014, which was more than double the guideline for the protection of aquatic life equal to 0.09 µg/L (Figure 3.3). Al concentrations also exceeded the guideline of 300 µg/L by more than double in Gull Pond with concentrations of 693 µg/L, and also surpassed the guideline in Little Galloo in May 2014 with concentrations of 317 µg/L (Figure

3.3). Fe concentrations exceeded guidelines of 300 µg/L not only in bird-impacted ponds, Little Galloo and Gull, but also in non-impacted sites, Main Duck Pond 2 and Calf Pond, in the summer of 2016. Fe

concentrations were highest across all sampling sites and dates in Main Duck Pond 2 in July and August

2016 with values of ~1200-1300 µg/L, coinciding with extremely reduced water levels (Figure 3.3). Cu concentrations were uniformly higher (or nearly) than guidelines of 2 µg/L in all bird-impacted sites, with concentrations of 1.91 µg/L in East Brother (July 2016), 3.2 µg/L in Little Galloo, 2.51 µg/L in High

Bluff, and 5.11 µg/L in Gull. Se concentrations were higher than recommended guidelines of 1 µg/L in

Little Galloo (1.17 µg/L) and Gull (1.38 µg/L), and other bird-impacted sites were elevated compared to the non-impacted ponds, but below guidelines (Figure 3.3). Finally, Pb concentrations were above guidelines of 1 µg/L particularly in East Brother in July 2016 (2.3 µg/L), but also in Gull Pond (1.15

µg/L) and non-impacted site Main Duck Pond 2 in August 2016 (1.11 µg/L). Pb concentrations approached guidelines in Little Galloo (0.91 µg/L) and in non-impacted Calf Pond in July 2016 (0.97

µg/L), and was the only additional metal other than Fe that was enriched near or above guidelines in a non-impacted site. Concentrations approached but did not exceed guidelines for Cr and As to the greatest extent in Gull Pond. Cr appeared greatly enriched compared to the other sampling sites, but As appeared only mildly enriched with concentrations in the non-impacted sites somewhat comparable to those of the bird-impacted sites (Figure 3.3).

Other elements appeared enriched in the bird-impacted sites compared to the non-impacted sites, but were below the guidelines for the protection of aquatic life (Figure 3.3). These elements included Bi,

Cs, Co, Li, Mo, Ni, Rb, Sr, Sn, Ti, and V. These metals, as well as those that exceeded guidelines, showed patterns of differential enrichment across bird-impacted sites. Trace elements that were the most enriched in Gull Pond included: Cd, Cr, Cs, Ti, Al, Zn, Co, Cu, and Se (Figure 3.3). Of these metals, Al,

Zn, Co, Cu, and Se were also enriched in Little Galloo, as were (to a lesser extent) Cd, Cr, and Ti. Bi and

V were enriched in both Little Galloo and Gull, as well; yet they were higher in Little Galloo than Gull.

Rb, Li, and Pb were most enriched in East Brother for at least one sampling date, and they were second highest in Gull Pond, followed by High Bluff for Rb and Li, and Little Galloo for Pb. Sr and Sn were most enriched in High Bluff, with elevated strontium concentrations also occurring in East Brother and elevated Sn concentrations in Gull (Figure 3.3). Mo and Ni appeared to be the most uniformly elevated in 69

the bird-impacted ponds compared to the non-impacted ponds, with the exception of high concentrations in Main Duck Pond 2 and Calf Pond in the summer of 2016 when water levels in both were severely decreased. Mo was the most enriched in Little Galloo and Gull and not enriched in High Bluff, whereas

Ni was most enriched in Gull, then Little Galloo, and less so in East Brother during September 2013.

Sedimentary trace element chemistry Surface sediments from High Bluff, Gull, Little Galloo, and Calf ponds did not strictly follow patterns of trace element enrichment observed in the water chemistry, as many elements were highest in

Calf Pond, the non-impacted site (Supplemental Figure 3.1). However, Cd and Zn concentrations were elevated in Gull Pond, both of which approached the probable effects level (PEL) of sediment quality guidelines for freshwater (CCME 1999). With the exception of High Bluff, all sites crossed the lower threshold of the interim sediment quality guideline (ISQG) for Cd, including the non-impacted site, however only Gull and Little Galloo had concentrations higher than the ISQG for Zn (Supplemental

Figure 3.1). In addition, Cu, K, Mo, and Se (and possibly V) were notably higher in Gull Pond’s surface sediments relative to the other sites, and Ca and Sr were notably higher in Little Galloo.

Dissolved oxygen concentrations The hourly dissolved oxygen measurements in the bird-impacted pond, East Brother, demonstrate diurnal fluctuations from lows in the early morning (~6 AM) to highs in the early evening (~6 PM) over the duration of our sample from September 19th to 25th 2013 (Figure 3.4A). Highs in dissolved oxygen ranged from ~8 to just over 18 mg/L (~70 - 225% saturation), and lows ranged between ~2-8 mg/L (~20 -

70% saturation). The average water temperature also recorded by the YSI sonde during this time was

17°C, the average pH was 8.8, and the average specific conductance was 597 µS/cm.

Similarly, dissolved oxygen measurements in Main Duck Pond 2 (non-bird-impacted site) also fluctuated diurnally from lows of ~0-3 mg/L (~0 - 80% saturation) in the early morning to highs of ~9 -

33 mg/L (~120 - 400% saturation) in the early evening (Figure 3.4B). Measurements were made from

July 27th to August 17th 2016, and a brief breakdown in diurnal fluctuations occurred over August 3rd to

5th. The average water temperature for the duration of the oxygen measurements was 27°C. Our second eutrophic non-bird-impacted site, Calf, had the longest record of oxygen measurements taken from July

21st to September 22nd 2016, over which time dissolved oxygen also fluctuated from lows of 0 - 11.5 mg/L (0 - 125 %) in the early morning to highs of 2.6 - 20.9 mg/L (32 - 254 %) in the late afternoon or early evening (Figure 3.4C). The average July/August water temperature of Calf Pond was 26 to 27°C and the average September water temperature was 23°C. The logged dissolved oxygen measurements in

Calf Pond clearly showed that diurnal fluctuations shifted upward in concentration, but no longer went to zero at night by late-September, which should be considered when assessing the week-long profile of our bird-impacted site East Brother that was taken in late September of 2013 (Figure 3.4A).

East Brother sediment core Radiometric dates, VRS-chla, and δ15N for East Brother were first published by Stewart et al.

(2015). Radiometric dates were rerun with two additional samples (13.5 - 14.0 cm, 14.5 - 15.0 cm) measured on the gamma spectrometer at Queen’s University, since the next measured interval after 12 -

12.5 cm was 15 - 15.5 cm. With the additional samples, we have confirmed that the core is mixed until

~12 - 13 cm, and the intrinsic time resolution (or the approximate amount of time affected by mixing;

Eisenreich et al. 1989) is ~38 years (Figure 3.5A). Because of the mobile nature of 137Cs peaks in very organic sediments (Blais et al. 1995) and the discrepancy between the newly calculated intrinsic time resolution and the independent 137Cs peak indicating 1963, we have chosen to exclude the 137Cs date from the East Brother profile and instead rely only on the intrinsic time resolution as a rough indication of the multi-decadal time scale of the upper sediments.

Chironomid taxa with relative abundances of 5% or greater in at least one interval were plotted for East Brother Pond (Figure 3.5A). No taxon occurred at greater than 20% relative abundance throughout the core. The assemblage was predominantly Chironomus anthracinus-type, Chironomus plumosus-type, Glyptotendipes, Cricotopus/Orthocladius, Limnophyes/Paralimnophyes, Paratanytarsus, and Pseudochironomus. Taxonomic shifts also cannot be determined because of possible mixing in the

East Brother core, and the subtle nature of changes that are apparent likely reflect this. Chironomid remains were too sparse for enumeration below 8-8.5 cm depth in the East Brother core. However, concentrations of head capsules ranged between ~75-150 per gram of dry sediment in the top 8 cm.

Little Galloo sediment core

Radiometric dates for the Little Galloo core were first used in Hargan et al. (in review), and VRS- chla and δ15N were first prepared for Stewart et al. (Appendix B). Total lead concentrations and stable lead isotopes suggested that the Little Galloo core was disturbed, and thus the intrinsic time resolution

(Eisenreich et al. 1989) for the two plateaus apparent in the supported 210Pb profile were calculated

(Appendix A). The chironomid assemblage of bird-impacted Little Galloo Pond was predominantly comprised of the genus Smittia at ~30 - 40% relative abundance, followed by the group

Cricotopus/Orthocladius at ~20% (Figure 3.5B). A notable synchronized shift among many of the taxa present in the Little Galloo core occurred between approximately 4 and 1 cm depth. This shift was characterized by decreases in Smittia, Propsilocerus cf., and Chironomus anthracinus-type, as well as increases in Limnophyes/Paralimnophyes, Metriocnemus, Parakiefferiella,

Parametriocnemus/Paraphaenocladius, and later (at ~2 cm), Pseudochironomus (Figure 3.5B). This shift is also accompanied by fluctuations in the group Cricotopus/Orthocladius, which recorded small declines at both 1 and 4 cm. Chironomid head capsule abundance was insufficient for enumeration below 7 cm, and the head capsules per gram dry sediment increased from stable values of ~50 to ~75 around the time that species shifts occurred from 4 to 1 cm depth (Figure 3.5B). All shifts must be considered with caution, however, due to confirmed core mixing (Appendix A) and the lack of a reliable core chronology.

Pigeon sediment core Radiometric dates, VRS-chla, and δ15N from the Pigeon core were first published by Stewart et al. (2015). Chironomid remains were sparse in this core, with insufficient head capsules for reliable interpretations according to the minimum requirement of 50 whole head capsules set by Quinlan and

Smol (2001). Thus, chironomid assemblage data from Pigeon was not included in the statistical analysis

comparing bird-impacted and non-impacted assemblages, but were presented as supplementary material

(Supplemental Figure 3.1). For counted intervals, fewer than 30 whole head capsules were retrieved and most head capsules belonged to the genus Smittia. Chironomid subfossils were virtually absent before 6 -

6.5 cm depth in the Pigeon core. The head capsules per gram dry sediment increases from ~10 to 50 between ~6 cm and 1 cm depth (shortly after increasing trends in δ15N and VRS-chla) and then decreases to ~30 in the top 0 - 0.5 cm interval (Supplemental Figure 3.1).

Main Duck Pond 2 sediment core Radiometric dates, VRS-chla, and δ15N were first published for non-bird-impacted Main Duck

Pond 2 in Stewart et al. (2015). The chironomid taxa present in this core (Figure 3.6A) were similar to those from the bird-impacted sites (Figure 3.5). Two large taxonomic groupings and one genus dominated the Main Duck Pond 2 chironomid assemblage, with the Tanytarsini increasing from ~20% at the bottom of the core to ~40% in the top half (from pre-1900s to ~1977). The grouping of Cricotopus/Orthocladius also made up ~20% of the assemblage, followed by 15 - 20% Pseudochironomus. Main Duck Pond 2 had slightly lower relative abundances (~5%) of the two Chironomus species types (C. anthracinus-type and

C. plumosus-type) compared to the bird-impacted assemblages from East Brother and Little Galloo

(Figure 3.5, 3.6A). However, Main Duck Pond 2 had greater abundances of chironomids belonging to the

Tanypodinae, compared to the bird-impacted sites, which occurred at less than 5% relative abundance in these sites and were thus not plotted on the stratigraphies. The head capsule abundance per gram of dry sediment in Main Duck Pond 2 was variable over the record, recording numbers ~4 times as many (~400) as the bird-impacted cores (Figure 3.6A).

Calf sediment core Radiometric dates for the Calf core were first used in Hargan et al. (in review), and VRS-chla and

δ15N were first prepared for Stewart et al. (Appendix B). The distribution of taxa in the Calf Pond chironomid assemblage is largely similar to that of Main Duck Pond 2, with the exception of a few differences in the rare species that still occurred at 5% or greater relative abundance (Figure 3.6). The 73

Cricotopus/Orthocladius group constituted ~40% of the Calf assemblage, and the large grouping of

Tanytarsini made up another ~20% (Figure 3.6B). One of the only notable changes in the Calf chironomid assemblage was a decrease in the relative abundance of Limnophyes/Paralimnophyes from 20% in the bottom interval counted (40 cm, pre-1942) to ~5% of the assemblage above 30 cm (1942 and later), which was then accompanied by the appearance of low abundances of Podonominae species 1. The non- bird-impacted Calf Pond sediment core also had greater abundances of head capsules per gram dry sediment than the bird-impacted cores, as was the case with Main Duck Pond 2. Interestingly, the abundance of head capsules in Calf Pond steadily increased from 40 cm (pre-1940s) to 20 cm (~1983), and then decreased somewhat until ~2 cm (~2013), at which point it sharply increased again with slightly lesser abundances in the surface interval (Figure 3.6B).

Comparison of bird-impacted and non-impacted sediments A Principal Component Analysis was used to summarize the main similarities and differences between the bird-impacted assemblages of East Brother and Little Galloo with the non-impacted assemblages of Main Duck Pond 2 and Calf (Figure 3.7). Both non-impacted sites (Main Duck Pond 2 and Calf) plotted together with some overlap along PCA axis 2, whereas each impact site plotted separately from the non-impacted sites as well as each other. Within each cluster of points, several species arrows appeared closely associated. Bird-impacted East Brother separates from non-impacted sites along axis 2 due to higher abundances or the presence of Paratanytarsus, Endochironomus, Dicrotendipes,

Paracricotopus, Tanytarsus, and to a lesser extent Chironomus plumosus-type and

Limnophyes/Paralimnophyes. However, East Brother appeared fairly similar to the non-impacted sites on axis 1 (Figure 3.7). The opposite was true for bird-impacted Little Galloo, which plotted close to non- impacted sites on axis 2, but far apart on axis 1. The intervals from Little Galloo were distinguished from the other sites by Smittia, Metriocnemus, Parametriocnemus/Paraphaenocladius, Propsilopcerus, and

Psectrocladius. Species arrows for Parakieferiella and Chironomus anthracinus-type fell approximately mid-way between clusters of both bird-impacted sites. Interestingly, both bird-impact sites separate more

from each other than from either non-impacted site. The non-impacted sites were driven in the negative direction along PCA axis 2 by Pseudochironomus, Tanypodinae (which were not present in bird-impacted ponds), as well as undistinguished Tanytarsini species, which were not present in abundance in Little

Galloo, and even including identifiable Tanytarsini species, the Tanytarsini were less abundant in East

Brother.

Discussion

Waterbird-mediated impacts on water chemistry Site-specific and year-to-year differences were apparent in the water chemistry across our bird- impacted and non-impacted ponds (Figure 3.2, 3.3). Bird-impacted ponds had consistently higher nitrogen and phosphorus, chlorophyll-a (i.e. primary production), specific conductance, and ion concentrations (particularly potassium and sulfate), as has been previously shown for these islands in two separate studies (Stewart et al. 2015; Appendix B). One of the most striking differences across all bird- impacted and non-impacted sites is the large difference in total phosphorus (TP-u) concentrations (Table

3.1, Figure 3.2), which exceeded or approached 2000 µg/L in the bird-impacted ponds. Although, the non-impacted ponds still qualified as eutrophic and hypereutrophic by standards of the Ontario Ministry of the Environment (>30 µg/L, OME 2010), the presence of large water bird colonies led to a 1 - 2 magnitude of difference in TP-u. Differences in water chemistry among the bird-impacted sites were likely attributable to many factors including differences in pond morphology, nutrient and element cycling, and sampling date, as well as the species and number of birds contributing guano to the ponds.

Many trace elements in the water of the bird-impacted ponds were elevated compared to the non- impacted ponds (Figure 3.3), including some heavy metals (e.g. Cd, Pb, Zn) that are known to accumulate in the tissues of gull and cormorant species (King and Cromartie 1986; Elliot et al. 1992; Mora and

Anderson 1995; Gochfeld et al. 1996; Custer et al. 2007), as well as some micronutrients that are physiologically required in trace amounts by many animals (e.g. Co, Cr, Cu, Fe, Mn, Mo, Ni, Zn). In general, the largest array of metals and elements were elevated in Gull Pond, and secondarily in Little

Galloo, which may reflect the presence of Ring-billed Gulls on each island. The largely diverse and opportunistic feeding strategy of Ring-billed Gulls often involves foraging in farm fields, garbage dumps, and the benthic zone of lakes (Caron-Beaudoin et al. 2013), leading to a broad exposure to many contaminants, including various heavy metals. In particular, gulls that feed in farm fields may be exposed to large contaminant burdens, as earthworms (the largest component of animal biomass in soils) bioaccumulate heavy metals such as Cd from soils, making them a key route of contaminant transfer up trophic levels (Vandecasteele et al. 2004). Elevated heavy metal tissue concentrations have been shown in other gull species that prey on invertebrates from farm fields, particularly in the case of Pb and Cd

(Struger et al. 1987).

As Gull Pond had Cd concentrations that were twice the guidelines for the protection of aquatic life (CCME 1999), it is likely that the ~40,000 to 120,000 nesting Ring-billed Gulls on the island are the source of metals to the pond. Elevated metal and element concentrations common to both Gull Pond and

Little Galloo Pond (e.g. Cd, Cr, Cs, Ti, Al, Zn, Co, Cu, Se, Bi, and V) may be sourced predominantly from the Ring-billed Gull populations on both islands, as these metals were not similarly elevated in our cormorant-impacted sites, East Brother and High Bluff (Figure 3.3). Our water chemistry data showed that Cd, Cr, Al, Zn, Cu, and Se may be of particular interest when determining the impacts of large Ring- billed Gull colonies, as these metals and elements were above (Cd, Al, Cu, Se) or approaching (Cr, Zn) water quality guidelines for the protection of aquatic life (CCME 1999). Even though most of the contaminant-related research for gulls and cormorants has focused on persistent organic pollutants or organochlorine compounds (Fox et al. 1991; Larson et al. 1996; de Solla et al. 2016; Hughes et al. 2016), research on other North American species of gull and cormorant has demonstrated bioaccumulation of

Cd, Se, and Pb in various bird tissues (King and Cromartie 1986).

In contrast to gulls, cormorants are strictly piscivorous and currently feed on primarily round goby (Neogobius melanostomus) in Lake Ontario (Somers et al. 2003). The round goby, an invasive species in the Laurentian Great Lakes, is known to be pollution-tolerant and preys upon benthic filter- feeders that tend to accumulate metals (Marentette et al. 2010). Cormorants feeding on round goby may 76

constitute another pathway through which heavy metal contaminants can be transferred from Lake

Ontario to cormorants, and thus to our primarily cormorant-impacted study ponds, East Brother and High

Bluff. In fact, cormorants on the Canadian Atlantic coast have been shown to bioaccumulate heavy metals including Cd, Hg, and Pb (Elliot et al. 1992). Our predominantly cormorant-impacted sites appeared to be notably enriched in fewer elements than in the predominantly gull-impacted sites, but some of the enriched elements included Rb, Li, Pb, Sr, and Sn (Figure 3.3). Interestingly, Sr concentrations were elevated at only the cormorant-impacted sites and not at Gull or Little Galloo, even though some of the multi-species waterbird colony on Little Galloo is composed of cormorants. Other elements enriched in

East Brother or High Bluff were sometimes also enriched in the sites impacted by gull species (i.e. on

Gull or Little Galloo), but no other clear associations between species and enriched elements emerged.

Trace element concentrations were likely influenced by the same factors as our other water chemistry parameters, including pond morphology, sampling date, and bird dynamics, as well as regional effects of where birds forage, making site comparisons of water chemistry more complex. Heavy metal contamination can be localized to various areas of the Great Lakes (Hart et al. 1986; Mayer and Manning

1991), particularly in areas with a history of point-source pollution (e.g. Hamilton Harbour in Lake

Ontario), which may also translate to differential contamination of food sources such as round goby

(Marentette et al. 2010). Double-crested Cormorants are also long-range foragers, so sources of heavy metals to our study sites will also depend on bird-specific foraging habits. The trace element water chemistry, in particular, demonstrated the strong effect evaporation has on the chemistry of these shallow ponds (Figure 3.3). This was perhaps most notable in the concentrations of Fe in the non-impacted ponds, which in the summer and early fall of 2016 were concentrated to values above the guideline for the protection of aquatic life (300 µg/L, CCME 1999), when water levels were reduced from their typical depth of ~1 m to 10 cm. The effect of evaporative concentration must also be considered for the bird- impacted ponds, which are much shallower (often <50 cm deep) and especially susceptible to evaporation loss. In fact, in the summer of 2016, all bird-impacted sites were completely desiccated, except for East

Brother, which was sampled in June and therefore not yet dry. Depending on sampling date, water 77

concentrations of metals and elements may vary, and thus determining the concentrations of these substances in the guano of each species would be beneficial for assessing species-specific paths of contamination.

Interestingly, the elements elevated in the water column were not similarly elevated in the surface sediments of High Bluff, Gull, Little Galloo, and Calf (Supplemental Figure 3.1). The discrepancies between the water chemistry and surface sediments may have to do with how metals and other trace elements are deposited and recycled from the sediments, which in turn could also affect water concentrations. This may be particularly true for redox-sensitive elements, as there seem to be diurnal oxygen fluctuations in Calf Pond (Figure 3.4), which may also be the case for High Bluff, Gull, and Little

Galloo, as it was for the bird-impacted pond, East Brother. Cd and Zn concentrations in the surface sediments of Gull Pond approached the probable effects level of sediment quality guidelines for the protection of aquatic life (CCME 1999), suggesting that both metals may cause negative effects on benthic sediment-dwelling organisms in Gull Pond, such as chironomids.

Chironomid assemblage response to eutrophication and contamination Chironomid subfossil remains appeared in abundances sufficient for enumeration around ~8 cm depth and higher for East Brother and Little Galloo (Figure 3.4) and at ~6 cm for Pigeon (Supplemental

Figure 3.1). Diatoms (siliceous algae) analysed from these same cores also appear (in sufficient abundance) in the sediment record at the same time as chironomids, possibly suggesting that the formation of the pond occurred at this time, or that the ponds became more permanent. According to sedimentary VRS-chla, δ15N, and sterol analysis, these uppermost portions of the cores from the bird- impacted ponds also showed evidence of heavy bird influence (Stewart et al. 2015; Hargan et al. in review; Appendix B), and it may be that the onset of bird nesting, or bird population increases, led to the formation of a more permanent pond. For example, the establishment of large colonies of these waterbirds, particularly cormorants, has led to the loss of terrestrial vegetation due to accumulation of toxic amounts of guano (Boutin et al. 2011), leaving areas susceptible to erosion. Eroded depressions of

the island may collect water, especially if they have in the past, and possibly lead to pond formation or greater permanence of ponds. The increase in the abundance of chironomid head capsule remains in the bird-impacted cores was consistent with the findings of other studies of eutrophication in shallow ponds, as bird guano stimulated algal production, increasing food availability (Stewart et al. 2013; 2014). Food availability is also one of the main drivers of chironomid response to eutrophication in deeper lakes as well, alongside changes in dissolved oxygen (reviewed by Brodersen and Quinlan 2006).

As the sediment cores from bird-impacted sites, East Brother and Little Galloo, may be affected by sediment mixing (Stewart et al. 2015; Hargan et al. in review; Appendix A), changes in the chironomid assemblage within the bird-impacted portion of each core (upper ~8 cm) could not be determined (Figure

3.5). However, the species present in East Brother and Little Galloo were typical of shallow productive ponds, including some species that are tolerant of dry or semi-terrestrial conditions (e.g. Smittia,

Andersen et al. 2013). The semi-terrestrial genus, Smittia, made up most of the head capsules retrieved from Pigeon Pond (Supplemental Figure 3.2), which was the shallowest site sampled and regularly dries out. The dominant species found in our hypereutrophic bird-impacted sites (Figure 3.5), as well as the eutrophic non-impacted sites (Figure 3.6), are similar to those found in other temperate shallow eutrophic sites, including Chironomus species, Dicrotendipes, Endochironomus, and Glyptotendipes (Dévai and

Moldován 1983; Brodersen et al. 2001; Langdon et al. 2006).

The subtle differences between the chironomid assemblages of the bird-impacted and non- impacted ponds suggest that the large differences in total phosphorus concentration (Figure 3.3) do not result in drastically different species compositions, which may be partly attributable to the fact that all of our sites classify as eutrophic and have similar oxygen dynamics. During eutrophication, chironomid assemblages are generally governed directly by oxygen concentrations and only indirectly by nutrient concentrations, as in some cases the distribution of species shows high overlap between shallow eutrophic and oligotrophic sites where oxygen levels are decoupled from nutrient concentrations by shallow depth and 24-hours of daylight (Stewart et al. 2013; 2014). Furthermore, in situ experiments with chironomids in streams artificially supplied with phosphorus showed that chironomids cannot make use of additional 79

phosphorus added to the system in both low or high nutrient conditions (Small et al. 2011), supporting the idea that our large gradient of total phosphorus did not have a direct influence on the chironomid assemblages.

Furthermore, the bird-impacted sediments of East Brother and Little Galloo did not support high abundances of typical hypoxia-tolerant eutrophication-indicator taxa (e.g. Chironomus), as they are shallow and thus experience different oxygen dynamics compared to deep stratifying lakes. In East

Brother, oxygen appears to fluctuate diurnally from hypoxic conditions at night when respiration depletes oxygen to supersaturated in the day when photosynthesis supplies oxygen (Figure 3.4). However,

Chironomus did remain present throughout the bird-impacted and non-impacted ponds, both of which experience some periods of low oxygen (Figure 3.5, 3.6). This is in contrast to eutrophic ponds in the

Canadian High Arctic, which remain oxic 24-hours per day in the summer when the period of daylight is also 24-hours long (Supplemental Figure 3.3). Accordingly, the relative abundances of hypoxia-tolerant

Chironomus (which is not restricted by latitude, Walker et al. 1997) in these same High Arctic ponds were slightly lower (<5%) (Stewart et al. 2014) compared to our temperate eutrophic ponds (East Brother,

Main Duck Pond 2, Calf) that have nightly hypoxia/anoxia (Chironomus ~10%, Figure 3.5, 3.6).

The chironomid taxa present in the bird-impacted ponds vary somewhat from the non-impacted ponds, despite both being eutrophic and having similar oxygen dynamics (Figure 3.2, 3.4, 3.5, 3.6). Both bird-impacted ponds (East Brother and Little Galloo) separated out in the PCA ordination from the non- impacted ponds (Main Duck Pond 2 and Calf) based on their full-core chironomid assemblages (Figure

3.7), and only 10 of 22 total taxa were found in all ponds. The differences between impacted and non- impacted assemblages may be partly attributable to pond morphology, as the non-impacted ponds are generally deeper (>1 m) than the bird impacted ponds (>0.5 m), which in turn may support a greater diversity of habitat for chironomids. This was further supported by the observation that four taxa

(Cryptotendipes, Paracricotopus, Polypedilum, and Tanystarsus) were unique to the non-impacted ponds, whereas none of the total 22 taxa found were identified from the bird-impacted ponds alone (Figure 3.5,

3.6). The extreme environment of the impacted ponds, including a proneness to drying out and 80

concentrated water chemistry parameters (e.g. TP-u >2000 µg/L), may have limited chironomid diversity and survival. In the case of Pigeon, High Bluff, and Gull ponds, these factors may have inhibited chironomid production such that subfossils were too scarce for analysis.

In addition to extreme general water chemistry (nutrients and ions) concentrations and potential desiccation, some of the bird-impacted ponds had concentrations of heavy metals and other trace elements that exceeded guidelines for the protection of aquatic life (Figure 3.3, CCME 1999). Gull Pond, in particular, had concentrations of Cd, Cu, and Al that were twice the freshwater guidelines (CCME 1999),

Se levels above guidelines, and concentrations of Cr and Zn that approached guidelines (Figure 3.3).

Furthermore, the dry weight concentration of Zn and Cd in the surface sediments of Gull Pond

(Supplemental Figure 3.1) closely approached the probable effects level (PEL) set by the CCME (1999).

As chironomids are often associated with sediments, it may have been that bird-mediated enrichment of metals in Gull Pond prevented the abundant production of chironomids in otherwise eutrophic and food- rich Gull Pond.

The absence of chironomids in highly metal-polluted sites has been noted in Russian lakes impacted by mining (Brooks et al. 2005), and some of the concentrations of the same metals in Gull Pond

(e.g. Cd, Figure 3.4) exceed those recorded from the heavily polluted Russian sites. Chironomids (of the genera Chironomus and Glyptotendipes) have been assayed for genotoxic effects of many metals, including Al, Cr, Cu, and Pb, demonstrating metal-induced structural and functional deformities in large polytene chromosomes (Michailova and Belcheva 1990; Michailova et al. 2012). In the cases of Al and

Pb, concentrations found in Gull Pond exceeded those found to have negative effects in the toxicological assays. Furthermore, morphological deformities, as well as reduced survivorship, have also been noted in many chironomids (e.g. Chironomus, Polypedilum, Cryptochironomus, and Dicrotendipes) in response to metal pollution in natural settings (Warwick et al 1987; Diggins and Stewart 1993). The presence of

Chironomus, as well as Glyptotendipes, Dicrotendipes, and Psectrocladius, in our bird-impacted ponds may also be partly governed by metal concentrations, since despite possible deformities, these species have been found to tolerate metal pollution (Diggins and Stewart 1993; Ilyashuk et al. 2003; Thienpont et 81

al. 2016; Stewart et al. in review). Many of these taxa, including Chironomus and Dicrotendipes, were also recorded in the heavily metal-contaminated inner harbour of Port Hope on Lake Ontario (Hart et al.

1986).

Previous analyses of diatoms from the bird-impacted ponds showed subfossil algal assemblages unsurprisingly dominated by eutrophic species (Stewart et al. 2015; Appendix B). However, many of these species have also been reported as tolerant of metal pollution, particularly Nitzschia palea, Eolimna minima (reported as Navicula minima), Navicula atomus (reported as Mayamaea atomus), Gomphenema parvulum, and Achnanthidium minutissimum (Ivorra et al. 1999; Morin et al 2008; da Silva et al. 2009;

Chen et al. 2014). For example, in Gull Pond, our site with generally the highest concentrations of metals, the diatom assemblage was dominated by M. atomus (Appendix B), a species that has been found to tolerate high concentrations of Zn and Cd (Ivorra et al. 1999), both of which were elevated in Gull’s water and sediments to concentrations near or above Canadian guidelines (Figure 3.3, Supplementary Figure

3.1). Thus, this possible bird-mediated metal pollution may have been partly responsible for the composition of both metal-tolerant chironomid and diatom species in the bird-impacted ponds.

Conclusions

Our analysis of water chemistry and subfossil chironomids from shallow ponds on nesting islands of Double-crested Cormorants and/or Ring-billed Gulls in Lake Ontario indicated that bird-mediated eutrophication and heavy metal pollution has shaped the biology of these sites. Water concentrations of

Cd, Al, Cu, and Se exceeded the Canadian guidelines for the protection of aquatic life (CCME 1999) in one or more of our bird-impacted sites, suggesting that the presence of large nesting colonies and consequent deposition of large amounts of guano can lead to toxic levels of metals and other trace elements. Some of these metals have already been shown to accumulate in gulls or cormorants, as well as their food sources, and thus the birds are likely the predominant source of contaminant-transfer to the study ponds. However, some of the elements enriched in our bird-impacted sites compared to the non- impacted sites are micronutrients essential to many organisms, including birds, and thus enrichment in the

ponds may have occurred because of the sheer number of birds contributing guano to the landscape, rather than by biomagnification processes up the food chain.

Chironomid assemblages were different in the bird-impacted ponds compared to the non- impacted ponds, though oxygen dynamics, one of the main drivers of chironomid species assemblage changes under eutrophication scenarios, were similar in impacted and non-impacted ponds. Differences in the assemblages may be partly due to pond morphology (particularly depth), but may also be due to toxicological effects of metal contaminants that were higher at the impacted sites. In fact, one of the most metal-contaminated sites (Gull Pond) contained only sparse chironomid remains, similar to research conducted on lakes with mining-related metal contamination (Brooks et al. 2005). Previous analyses of diatom assemblages in our study ponds further support the idea that metal contamination has helped shape the biology of these ponds, as assemblages were dominated by diatoms known to tolerate high concentrations of metals such as Cd and Zn. Down-core analysis of metals could link the arrival of bird colonies to these islands with metal contamination, and the analysis of trace elements in guano could associate species to specific contaminants in the ponds.

Acknowledgements

We would like to thank the Natural Sciences and Engineering Research Council for funding;

Xioawa Wang and the National Laboratory for Environmental Testing for water chemistry analyses; Chip

Weseloh for field assistance and bird data for East Brother and Pigeon Islands; Irene Mazzocchi and the

New York State Department of Environmental Conservation for field access and assistance and bird data for Little Galloo and Calf Islands; Parks Canada for access to Main Duck Island; Presqu’ile Provincial

Park for access to and bird data from High Bluff and Gull; as well as members of the Paleoecological

Environmental Assessment and Research Laboratory at Queen’s University for field assistance throughout.

References

Andersen T, Cranston PS, Epler JH (2013) The larvae of Chironomidae (Diptera) of the Holarctic region -

keys and diagnoses. Insect Syst Evol Supp 66:1–571

Appleby PG (2001) Chronostratigraphic techniques in recent sediments. In: Last WM, Smol JP (eds)

Tracking Environmental Change Using Lake Sediments. Volume 1: Basin Analysis, Coring and

Chronological Techniques. Kluwer Academic Publishers, Dordrecht, pp 172–203

Blais JM, Kalff J, Cornett RJ, Evans RD (1995) Evaluation of 210Pb dating in lake sediments using stable

Pb, Ambrosia pollen, and 137Cs. J Paleolimnol 13:169–178

Blais JM, Kimpe LE, McMahon D, Keatley BE, Mallory ML, Douglas MSV, Smol JP (2005) Arctic

seabirds transport marine-derived contaminants. Science 309:445

Boutin C, Dobbie T, Carpenter D, Hebert CE (2011) Effects of double-crested cormorants

(Phalacrocorax auritus Less.) on island vegetation, seedbank, and soil chemistry: evaluating

island restoration potential. Restor Ecol 19:720–727.

Brodersen KP, Odgaard BV, Vestergaard O, Andersen NJ (2001) Chironomid stratigraphy in the shallow

and eutrophic Lake Søbygaard, Denmark: chironomid–macrophyte co-occurrence. Freshw Biol

46:253–267

Brodersen KP, Quinlan R (2006) Midges as paleoindicators of lake productivity, eutrophication and

hypolimnetic oxygen. Quat Sci Rev 25:1995–2012

Brooks SJ, Langdon PG, Heiri O (2007) The Identification and Use of PalaeArctic Chironomidae Larvae

in Palaeoecology. Quaternary Research Association, London

Brooks SJ, Udachin V, Williamson BJ (2005) Impact of copper smelting on lakes in the southern Ural

Mountains, Russia, inferred from chironomids. J Paleolimnol 33:229–241

Canadian Council of Ministers of the Environment (CCME) (1999) Canadian environmental quality

guidelines. Canadian Council of Ministers of the Environment, Winnipeg

Caron-Beaudoin É, Gentes M-L, Patenaude M, Hélie J-F, Giroux J-F, Verreault J (2013) Combined usage

of stable isotopes and GPS-based telemetry to understand the feeding ecology of an omnivorous

bird, the Ring-billed Gull (Larus delawarensis). Can J Zool 91:689–697

Chen X, Li C, McGowan S, Yang X (2014) Diatom response to heavy metal pollution and nutrient

enrichment in an urban lake: Evidence from paleolimnology. Ann Limnol – Int J Lim 50:121–130

Custer TW, Custer CM, Eichhorst BA, Warburton D (2007) Selenium and metal concentrations in

waterbird eggs and chicks at Agassiz Nation Wildlife Refuge, Minnesota. Archiv Environ

Contam Toxicol 53:103–109 da Silva EF, Almeida SFP, Nunes ML, Luís AT, Borg F, Hedlund M, Marques de Sá C, Patinha C,

Teixeira P (2009) Heavy metal pollution downstream of the abandoned Coval da Mó mine

(Portugal) and associated effects on epilithic diatom communities. Sci Tot Environ 407:5620–

5636 de Solla SR, Weseloh DVC, Hughes KD, Moore DJ (2016) Forty-year decline of organic contaminants in

eggs of Herring Gulls (Larus argentatus) from the Great Lakes, 1974 to 2013. Waterbirds

39(sp1):166–179

Dévai G, Moldován J (1983) An attempt to trace eutrophication in a shallow lake (Balaton, Hungary)

using chironomids. Hydrobiol 103:169–175

Diggins TP, Stewart KM (1993) Deformities of aquatic larval midges (Chironomidae: Diptera) in the

sediments of the Buffalo River, New York. J Gt Lakes Res 19:648–659

Eisenreich SJ, Capel PD, Robbins JA, Bourbonniere R (1989) Accumulation and diagenesis of

chlorinated hydrocarbons in lacustrine sediments. Environ Sci Technol 23:1116-1126

Elliott JE, Scheeuhammer AM, Leighton FA, Pearce PA (1992) Heavy metals and metallothionein

concentrations in Atlantic Canadian seabirds. Archiv Environ Contam Toxicol 22:63–73

Environment Canada Manual of Analytical Methods (1994a) Major Ions and nutrients vol 1. National

Laboratory for Environmental Testing, Canadian Centre for Inland Waters, Burlington

Environment Canada Manual of Analytical Methods (1994b) Trace Metals vol 2. National Laboratory for

Environmental Testing, Canadian Centre for Inland Waters, Burlington

Fox GA, Collins B, Hayakawa E, DV Weseloh, Ludwig JP, Kubiak TJ, Erdman TC (1991) Spatial

variation in the occurrence and prevalence of bill defects in young Double-Crested Cormorants in

the Great Lakes, 1979-1987. J Gt Lakes Res 17:158–167

Glew JR (1988) A portable extruding device for close interval sectioning of unconsolidated core samples.

J Paleolimnol 1:235–239

Glew JR, Smol JP (2016) A push corer developed for retrieving high-resolution sediment cores from

shallow waters. J Paleolimnol 56:67–71

Gochfeld M, Belant JL, Shukla T, Benson T, Burger J (1996) Heavy metals in Laughing Gulls: gender,

age and tissue differences. Environ Toxicol Chem 15:2275–2283

Hargan KE, Stewart EM, Michelutti N, Grooms C, Kimpe L, Mallory M, Smol JP, Blais JM (in review)

Incorporating sterols and stanols as biomarkers for tracking waterbird impacts to temperate

ponds. Proc R Soc Lond B Biol Sci: manuscript ID RSPB-2017-2669

Hart DR, McKee PM, Burt AJ (1986) Benthic community and sediment quality assessment of Port Hope

Harbour, Lake Ontario. J Gt Lakes Res 12:206–220

Hughes KD, de Solla SR, Weseloh DVC, Martin PA (2016) Long-term trends in legacy contaminants in

aquatic wildlife in the Hamilton Harbour Area of Concern. Aquat Ecosys Health Manag 19:171–

180

Ilyashuk B, Ilyashuk E, Dauvalter V (2003) Chironomid responses to long-term metal contamination: a

paleolimnological study in two bays of Lake Imandra, Kola Peninsula, northern Russia. J

Paleolimnol 30:217–230

Ivorra N, Hettelaar J, Tubbing GMJ, Kraak MHS, Sabater S, Admiraal W (1999) Translocation of

microbenthic algal assemblages used for in situ analysis of metal pollution in rivers. Arch

Environ Contam Toxicol 37:19–29

Keatley BE, Douglas MS, Blais JM, Mallory ML, Smol JP (2009) Impacts of seabird-derived nutrients on

water quality and diatom assemblages from Cape Vera, Devon Island, Canadian High Arctic.

Hydrobiologia 621:191–205.

King KA, Cromartie E (1986) Mercury, lead, and selenium in three waterbird species nesting in Gaveston

Bay, Texas, USA. Colonial Waterbirds 9:90–94.

Langdon PG, Ruiz Z, Brodersen KP, Foster IDL (2006) Assessing lake eutrophication using chironomids:

understanding the nature of community response in different lake types. Freshw Biol 51:562–577

Larson JM, Karasov WH, Sileo L, Stromborg KL, Hanbridge BA, Giesy JP, Jones PD, Tillitt DE,

Verbrugge DA (1996) Reproductive success, developmental anomalies, and environmental

contaminants in Double-crested Cormorants (Phalacrocorax auritus). Environ Toxicol Chem

15:553–559

Legendre P, Birks HJB (2012) From classical to canonical ordination. In: Birks HJB, Lotter AF, Juggins

S, Smol JP (eds) Tracking Environmental Change Using Lake Sediments. Volume 5: Data

Handling and Numerical Techniques. Springer, Dordrecht, pp 201–248

Little JL, Smol JP (2001) A chironomid-based model for inferring late-summer hypolimnetic oxygen in

southeastern Ontario lakes. J Paleolimnol 26:259–270

Ludwig JP (1974) Recent changes in the ring-billed gull population and biology in the Laurentian Great

Lakes. Auk 91:575–594

Ludwig JP (2013) The Dismal State of the Great Lakes: An ecologist’s analysis of why it happened and

how to fix the mess we have made. Xlibris, Bloomington

Luoto TP, Salonen V-P (2010) Fossil midge larvae (Diptera: Chironomidae) as quantitative indicators of

late-winter hypolimnetic oxygen in southern Finland: a calibration model, case studies and

potentialities. Boreal Environ Res 15:1–18

Luoto TP, Brooks SJ (2014) Ecological responses to climate change in a bird-impacted High Arctic pond

(Nordaustlandet, Svalbard). J Paleolimnol 51:87–97

Marentette JR, Gooderham KL, McMaster ME, Ng T, Parrott JL, Wilson JY, Wood CM, Balshine S

(2010) Signatures of contaminants in invasive round gobies (Neogobius melanstomus): A double

strike for ecosystem health? Ecotoxicol Environ Safety 73:1755–1764

Mayer T, Manning PG (1991) Evaluating inputs of heavy metal contaminants and phosphorus to Lake

Ontario from Hamilton Harbour. Water Air Soil Pollut 59:281–298

Michailova P, Belcheva R (1990) Different effect of lead on external morphology and polytene

chromosomes of Glyptotendipes barbipes (Staeger) (Diptera, Chironomidae). Folia Biol

(Kraków) 38:83–88

Michailova P, Sella G, Petrova N (2012) Chironomids (Diptera) and their salivary gland chromosomes as

indicators of trace-metal genotoxicity. Ital J Zool 79:218–230

Michelutti N, Keatley BE, Brimble S, Blais JM, Liu H, Douglas MSV, Mallory ML, Macdonald RW,

Smol JP (2009) Seabird-driven shifts in Arctic pond ecosystems. Proc Roy Soc B 276:591–596

Michelutti N, Blais JM, Mallory ML, Brash J, Thienpont J, Kimpe LE, Douglas MSV, Smol JP (2010)

Trophic position influences the efficacy of seabirds as metal biovectors. Proc Natl Acad Sci USA

107:10543–10548

Mora MA, Anderson DW (1995) Selenium, boron, and heavy metals in birds from the Mexicali Valley,

Baja California, Mexico. Bull Environ Contam Toxicol 54:198–206

Morin S, Duong TT, Dabrin A, Coynel A, Herlory AC, Baudrimont M, Delmas F, Durrieu G, Schäfer J,

Winterton P, Blanc G, Coste M (2008) Long-term survey of heavy-metal pollution, biofilm

contamination and diatom community structure in the Riou Mort watershed, South-West France.

Environ Pollut 151:532–542

Morris RD, Weseloh DV, Wires LR, Pekarik C, Cuthbert FJ, Moore DJ (2011) Population trends of Ring-

billed Gulls breeding on the North American Great Lakes, 1976 to 2009. Waterbirds 34:202–212

Ontario Ministry of Natural Resources (OMNR) (2011) Presqu’ile Islands Resource Management

Implementation Plan. Queen’s Printer for Ontario

Ontario Ministry of the Environment (OME) (2010) Lakeshore Capacity Assessment Handbook

Protecting Water Quality in Inland Lakes on Ontario’s Precambrian Shield. Queen’s Printer for

Ontario

Quinlan R, Smol JP (2001) Setting minimum head capsule abundance and taxa deletion criteria in

chironomid-based inference models. J Paleolimnol 26:327–342

Rush SA, Pekarik C, Weseloh DV, Cuthbert F, Moore D, Wires L (2015) Changes in heron and egret

populations on the Laurentian Great Lakes and connecting channels, 1977-2009. Avian Conserv

Ecol 10:7

Small GE, Wares JP, Pringle CM (2011) Differences in phosphorus demand among detritivorous

chironomid larvae reflect intraspecific adaptations to differences in food resource stoichiometry

across lowland tropical streams. Limnol Oceanogr 56:268–278

Somers CM, Lozer MN, Kjoss VA, Quinn JS (2003) The invasive Round Goby (Neogobius melanstomus)

in the diet of nestling Double-crested Cormorants (Phalacrocorax auritus) in Hamilton Harbour,

Lake Ontario. J Gt Lakes Res 29:392-399

Stewart EM, Hargan KE, Sivarajah B, Kimpe LE, Blais JM, Smol JP (in review) A paleoenvironmental

study tracking eutrophication, metal pollution, and climate change in Niven Lake (NT),

Yellowknife’s first sewage lagoon. Arctic: ms#17-134

Stewart EM, McIver RM, Michelutti N, Douglas MSV, Smol JP (2014) Assessing the efficacy of

chironomids and diatoms in tracking eutrophication in High Arctic sewage ponds. Hydrobiologia

721: 251–268

Stewart EM, Michelutti N, Blais JM, Mallory ML, Douglas MSV, Smol JP (2013) Contrasting the effects

of climatic, nutrient, and oxygen dynamics on subfossil chironomid assemablges: a

paleolimnological experiment from eutrophic High Arctic ponds. J Paleolimnol 49: 205–219

Stewart EM, Michelutti N, Shenstone-Harris S, Grooms C, Weseloh C, Kimpe LE, Blais JM, Smol JP

(2015) Tracking the history and ecological changes of rising Double-crested Cormorant

populations using pond sediments from islands in eastern Lake Ontario. PLoS ONE 10:e0134167 89

Struger J, Elliott JE, Weseloh DV (1987) Metals and essential elements in Herring Gulls from the Great

Lakes, 1983. J Gt Lakes Res 13:43–55

Thienpont JR, Korosi JB, Hargan KE, Williams T, Eickmeyer DC, Kimpe LE, Palmer MJ, Smol JP, Blais

JM (2016) Multi-trophic level response to extreme metal contamination from gold mining in a

subarctic lake. Proc Roy Soc B 283:20161125

Vandecasteele B, Samyn J, Quataert P, Muys B, Tack FMG (2004) Earthworm biomass as additional

information for risk assessment of heavy metal biomagnification: a case study for dredged

sediment-derived soils and polluted floodplain soils. Environ Pollut 129:363–375

Walker IR (2001) Midges: Chironomidae and related Diptera. In: Smol, JP, Birks HJB, Last WM (eds)

Tracking Environmental Change Using Lake Sediments volume 4, Zoological Indicators. Kluwer

Academic Publishers, Dordrecht, pp 43-66

Walker IR (2007) The WWW Field Guide to Fossil Midges. (http://www.paleolab.ca/wwwguide/)

Walker IR, Levesque AJ, Cwynar LC, Lotter AF (1997) An expanded surface-water palaeotemperature

inference model for use with fossil midges from eastern Canada. J Paleolimnol 18:165–178

Warwick WF, Fitchko J, McKee PM, Hart DR, Burt AJ (1987) The incidence of deformities in

Chironomus spp. from Port Hope Harbour, Lake Ontario. J Gt Lakes Res 13:88–92

Weir RD (2008) Birds of the Kingston Region 2nd ed. Kingston Field Naturalists, Kingston

Weseloh DVC, Pekarik C, Havelka T, Barrett G, Reid J (2002) Population trends and colony locations of

double-crested cormorants in the Canadian Great Lakes and immediately adjacent areas, 1990–

2000: A manager’s guide. J Gt Lakes Res 28:125–144

Weseloh DV, Ewins PJ, Struger J, Mineau P, Bishop CA, Postupalsky S, Ludwig JP (1995) Double-

crested Cormorants of the Great Lakes: Changes in population size, breeding distribution and

reproductive output between 1913 and 1991. Colonial Waterbirds 18(sp1):48–59

Figure 3.1 Study sites (black dots) in eastern Lake Ontario. Ponds on each bird-impacted island are: East Brother, Pigeon, Little Galloo, High Bluff, and Gull. Non-bird-impacted sites include: Main Duck Ponds 1 and 2, as well as Calf Pond. The inset shows the sampling region within the Laurentian Great Lakes of North America.

Table 3.1 Pond characteristics, pH, specific conductance (cond.), total phosphorus (TP-u), and total dissolved phosphorus (TP-f) for each bird- impacted site (EB=East Brother, PGN=Pigeon, LG=Little Galloo, HB=High Bluff, GU=Gull) and non-impacted site (MD1=Main Duck 1, MD2=Main Duck 2, CF=Calf). The estimated total number of birds nesting on each island represents a composite of species. The pond dimensions (m) were taken from Google Earth, and the approximate maximum depth (m) was estimated at the time of sampling. Pond traits & BIRD-IMPACTED NON-IMPACTED water chemistry EB PGN LG HB GU MD1 MD2 CF Approx. # nests * 1500 1700 50,000 4000 30,000 0 0 0; 170ŧ Dimensions (m) 50×35 20×20 28×120 35×45 125×30 220×70 120×110 406×65 Max. depth (m) 0.5 0.1 0.4 0.15 0.3 1 1 1 pH§ 7.7 7.0 9.5 8.6 8.8 7.8 8.3 8.5 Cond. (µS/cm) § 770 1650 193.4 1350 547 209 218.3 172.5 TP-u (µg/L) § 2205 - 6410 4310 8130 36.4 86.8 151 TP-f (µg/L) § 2050 - 3160 3280 3030 12.1 25.9 136 *Numbers are 2015 field estimates for EB and PGN; from OMNR (2011) for HB and GU; and from NYSDEC 2015 data for LG and 2008 data for CF. §Averaged values from field measures for EB. ŧin 2008.

Figure 3.2 Select water chemistry data from bird-impacted sites, East Brother (EB), Little Galloo (LG), High Bluff (HB), and Gull (GU), as well as for non-impacted sites, Main Duck Pond 1 (MD1), Main Duck Pond 2 (MD2), and Calf (CF). Sampling dates are given in order below graphs, including three measures from each EB, MD2, and CF. TN is total nitrogen and TP is total phosphorus with “-u/-f” indicating unfiltered or filtered values, respectively. TKN is total Kjeldahl nitrogen. Chl-a is chlorophyll- a. DOC and POC are dissolved and particulate organic carbon, respectively. Cond is specific conductance. Raw data are given in Supplemental Table 3.1.

Figure 3.3 Total concentrations of trace elements in water (µg/L) from bird-impacted ponds, East Brother (EB), Little Galloo (LG), High Bluff (HB), and Gull (GU), as well as for non-impacted sites, Main Duck Pond 1 (MD1), Main Duck Pond 2 (MD2), and Calf (CF). Sampling dates are given in order below graphs, including three measures from each EB, MD2, and CF. Parameters are organized by the site in which they are most elevated, and horizontal dashed lines indicate freshwater quality guidelines for the protection of aquatic life (CCME 1999). The chromium guideline is for the hexavalent form. Raw data are given in Supplemental Table 3.2. 94

Figure 3.4 Dissolved oxygen concentrations (mg/L) and saturation (%) in A. bird-impacted East Brother Pond measured hourly from Sept 19th to 25th 2013, B. non-impacted Main Duck Pond 2 measured hourly from July 27th to August 17th 2016, and C. non-impacted Calf Pond measured hourly from July 21st to September 21st 2016. 95

Figure 3.5 Chironomid assemblages of bird-impacted sites: A. East Brother Pond and B. Little Galloo Pond. Total number of head capsules retrieved per gram of dry sediment for analysed intervals (HC/g dry sed), sedimentary ratios of stable nitrogen isotopes (δ15N in ‰), and sedimentary chlorophyll-a concentrations (VRS-chla in mg/g) are also give to the left of species data. δ15N and VRS-chla profiles are modified from Stewart et al. (2015) for East Brother and Pigeon, and from Stewart et al. (Appendix B) for Little Galloo.

Figure 3.6 Chironomid assemblages of non-bird-impacted sites: A. Main Duck Pond 2 and B. Calf Pond. Total number of head capsules retrieved per gram of dry sediment for analysed intervals (HC/g dry sed), sedimentary ratios of stable nitrogen isotopes (δ15N in ‰), and sedimentary chlorophyll-a concentrations (VRS-chla in mg/g) are also given to the left of species data. δ15N and VRS-chla profiles are modified from Stewart et al. (2015) for Main Duck Pond 2 and from Stewart et al. (Appendix B) for Calf.

Figure 3.7 Principal Component Analysis (PCA) of Hellinger-transformed down-core chironomid species data for bird-impacted ponds, East Brother and Little Galloo, as well as non-impacted ponds, Main Duck Pond 2 and Calf. Species abbreviations are: Chaet = Chaetocladius, ChirAnth = Chironomus anthracinus, ChirPlum = Chironomus plumosus, Crico = Cricotopus/Orthocladius, Crypt = Cryptotendipes, Dicro = Dicrotendipes, Endo = Endochironomus, Glypto = Glyptotendipes, Limnophy = Limnophyes/Paralimnophyes, Metrio = Metriocnemus, Microt = Microtendipes, Ortho = Orthocladiinae, Paracr = Paracricotopus, Parakief = Parakiefferiella, Param = Parametriocnemus/Paraphaenocladius, Parat = Paratanytarsus, Polyp = Polypedilum, Psectro = Psectrocladius, Pseudoch = Pseudochironomus, Propc = Propsilocerus cf., Smitt = Smittia, Tanyp = Tanypodinae, TanySpp = Tanytarsini species, and Tanyt = Tanytarsus.

Supplemental Table 3.1 Select water chemistry for three sampling dates from East Brother (EB), one sampling date from each Little Galloo (LG), High Bluff (HB), Gull (GU), and Main Duck Pond 1 (MD1), as well as three sampling dates from each Main Duck Pond 2 (MD2) and Calf (CF). Basic water BIRD-IMPACTED NON-IMPACTED chemistry EB EB EB LG HB GU MD1 MD2 MD2 MD2 CF CF CF Sampling date Jun Sep Jul May Apr Apr Sep Sep Jul Aug Jun Jul Sep 2013 2013 2016 2014 2014 2014 2013 2013 2016 2016 2014 2016 2016 TN-f (mg/L) 9.97 3.11 6.15 3.47 109 30.2 0.896 1.46 1.91 1.89 1.04 1.57 2.22 NO2+NO3 (mg/L) 0.736 0.036 0.095 0.013 104 23.5 0.015 0.014 0.006 0.01 0.018 0.019 0.009 NH3 (mg/L) 7.46 0.101 3.3 0.049 0.209 2.28 0.022 0.056 0.032 0.077 0.047 0.022 0.021 TKN (mg/L) 9.7 3.28 6.8 3.38 5.2 8.25 0.898 1.46 2.73 2.31 1.36 1.75 2.46 TP-u (µg/L) 2970 1440 1940 6410 4310 8130 36.4 86.8 52.8 53.9 151 333 449 Chl-a (µg/L) 1.2 1.2 0.1 221 3.6 210 4.6 3.6 2.3 2.5 2.1 12 2.4 DOC (mg/L) 20.5 31.6 28 31.1 24.4 29.1 13.1 17.6 26.4 24.8 12.4 17.8 28.4 POC (mg/L) 5.29 1.34 0.977 50.2 2.12 29.8 0.921 2.98 1.48 2.43 0.979 5.38 2.61 pH - 9 - 9.5 8.6 8.8 7.8 7.6 9.2 9.3 9.4 8.3 8.3 Cond (µS/cm) 910 600 - 193 1350 547 209 252 153 250 197 148 495 Cl- (mg/L) 22 34.3 33.7 6.7 23.3 21.6 1.83 3.52 6.1 11.4 5.45 19.5 24.8 2- SO4 (mg/L) 76 55.1 65.2 16.2 125 58.6 1.37 1.73 0.5 0.95 1.2 2.17 2.16 K+ (mg/L) 46.5 57.9 57.7 14.3 42.8 54.1 0.24 0.39 0.04 0.7 0.44 1.89 0.69 Na+ (mg/L) 17.6 27.1 25.6 7.87 22.9 19.2 1.17 1.32 3.71 6.07 4.23 11.1 12.7 Ca2+ (mg/L) 110 57.8 81.9 32.2 250 72.2 31 48 22.4 38.7 34.8 48.2 57.5 Mg2+ (mg/L) 11.4 7.99 10.8 2.72 14 9.91 1.99 2.45 1.91 2.75 3.55 4.68 5.45 *TN is total nitrogen and TP is total phosphorus with “-u/-f” indicating unfiltered or filtered values, respectively. TKN is total Kjeldahl nitrogen. Chl-a is chlorophyll-a. DOC and POC are dissolved and particulate organic carbon, respectively. Cond is specific conductance.

Supplemental Table 3.2 Trace element concentrations (µg/L) listed alphabetically for three sampling dates from East Brother (EB), one sampling date from each Little Galloo (LG), High Bluff (HB), Gull (GU), and Main Duck Pond 1 (MD1), as well as three sampling dates from each Main Duck Pond 2 (MD2) and Calf (CF). Trace BIRD-IMPACTED NON-IMPACTED Metals EB EB EB LG HB GU MD1 MD2 MD2 MD2 CF CF CF Date  Jun Sep Jul May Apr Apr Sep Sep Jul Aug Jun Jul Sep Conc. (µg/L) 2013 2013 2016 2014 2014 2014 2013 2013 2016 2016 2014 2016 2016 Al 15.6 9.2 35.3 317 116 693 25.4 46.6 32.8 60.4 38.5 138 181 As 3.92 2.74 2.68 4.25 3.75 4.82 0.82 1.67 3.09 3.58 1.53 2.18 3.3 Bi 0.002 0 0.005 0.016 0.003 0.01 0.004 0.005 0.004 0.004 0.001 0.003 0.005 Cd 0.022 0.003 0.054 0.073 0.025 0.24 0.026 0.004 0.009 0.029 0.025 0.013 0.015 Co 0.23 0.11 0.175 0.613 0.184 1.63 0.035 0.067 0.059 0.076 0.112 0.124 0.215 Cr 0.07 0.06 0.14 0.33 0.19 0.86 0.05 0.03 0.06 0.11 0.11 0.19 0.24 Cs 0.017 0.013 0.009 0.015 0.007 0.04 0.001 0.003 0.002 0.005 0.004 0.008 0.008 Cu 0.61 0.31 1.91 3.2 2.51 5.11 0.44 0.47 0.19 0.46 1.69 0.46 0.71 Fe 40.8 24.7 46.2 462 112 733 68.4 211 1240 1360 264 484 323 Li 1.74 1.9 1.62 0.26 1.06 1.43 0.15 0.21 0.28 0.4 0.13 0.3 0.23 Mo 0.753 0.848 0.633 1.02 0.311 0.96 0.167 0.491 0.184 0.343 0.294 0.519 0.837 Ni 0.76 0.59 1.01 1.48 0.91 1.65 0.6 0.38 0.25 0.31 0.3 0.41 0.56 Pb 0.254 0.529 2.3 0.914 0.468 1.15 0.225 0.228 0.793 1.11 0.13 0.97 0.75 Rb 19.3 23.3 22.3 2.81 4.47 12 0.316 1.03 0.151 1.47 0.514 1.25 0.602 Se 0.53 0.46 0.73 1.17 0.58 1.38 0.1 0.19 0.21 0.22 0.15 0.16 0.25 Sn 0 0.019 0.048 0.041 1.2 0.79 0.032 0 0.012 0.016 0.068 0 0.009 Sr 279 231 127 40.6 352 68.2 67.4 107 60 105 93.5 129 142 Ti 1.4 0.35 1.7 5.8 4 27.2 0.2 0.9 0.65 1.57 0.8 5.15 6.86 V 0.291 0.3 0.737 10.8 0.907 5.73 0.504 0.489 0.74 1.23 1.03 1.95 5.01 Zn 1.8 1.1 4 10 4.9 25.7 1.1 0.7 0.5 1.7 0.6 1.5 1.4 *Below detection limit values were taken as equal to zero.

100

Supplemental Table 3.3 Trace element concentrations (µg/L) not presented in Figure 3.3, listed alphabetically for three sampling dates from East Brother (EB), one sampling date from each Little Galloo (LG), High Bluff (HB), Gull (GU), and Main Duck Pond 1 (MD1), as well as three sampling dates from each Main Duck Pond 2 (MD2) and Calf (CF). Trace BIRD-IMPACTED NON-IMPACTED Metals EB EB EB LG HB GU MD1 MD2 MD2 MD2 CF CF CF Date  Jun Sep Jul May Apr Apr Sep Sep Jul Aug Jun Jul Sep Conc. (µg/L) 2013 2013 2016 2014 2014 2014 2013 2013 2016 2016 2014 2016 2016 Ag 0.001 0 0.003 0.009 0.004 0.011 0.001 0.001 0.007 0.001 0.002 0.002 0.002 B 33.4 19.3 62.3 16.5 6.5 10.1 14.9 18.2 23.9 29.3 20.5 31.4 5.6 Ba 20.8 12.2 5.5 4.93 47.2 6.69 6.02 17.4 1.63 9.02 7.23 20.8 8.77 Be 0.004 0 0.005 0.014 0.006 0.026 0.003 0.005 0.005 0.01 0.005 0.011 0.014 Ce 0.025 0.026 0.052 0.593 0.276 1.59 0.037 0.1 0.139 0.313 0.102 0.6 0.778 Ga 0.006 0.018 0.025 0.392 0.047 0.307 0.053 0.085 0.057 0.072 0.06 0.117 0.172 Ge 0.013 0.013 0.025 0.246 0.136 0.673 0.021 0.046 0.064 0.134 0.05 0.279 0.361 La 10.3 7.99 3.66 78.7 6.25 65.3 24 71 45.5 69.7 106 208 133 Mn 0.753 0.848 0.633 1.02 0.311 0.956 0.167 0.491 0.184 0.343 0.294 0.519 0.837 Nb 0.007 0 0.006 0.023 0.013 0.073 0.001 0.004 0.003 0.003 0.009 0.008 0.016 Sb 0.179 0.161 0.254 0.3 0.14 0.202 0.078 0.135 0.158 0.188 0.077 0.126 0.219 Tl 0.003 0.003 0.002 0.007 0.024 0.013 0.001 0.002 0.002 0.004 0 0.004 0.003 U 0.074 0.070 0.046 0.167 0.117 0.144 0.085 0.095 0.053 0.078 0.138 0.217 0.457 W 0.017 0.007 0.082 0.031 0.045 0.058 0.043 0.004 0.005 0.02 0.044 0.006 0.02 Y 0.009 0.008 0.015 0.268 0.104 0.477 0.016 0.048 0.067 0.142 0.046 0.226 0.303 *Below detection limit values were taken as equal to zero.

101

Supplemental Figure 3.1 Trace metal and element concentrations in surface sediments (0-0.5 cm) of cores from High Bluff (cormorant-impacted), Gull (gull-impacted), Little Galloo (mixed species impact), and Calf (non-impacted) ponds. Where available, the interim sediment quality guideline (ISQG) (dotted grey line) and the probable effects level (PEL) (dotted black line) from the sediment quality guidelines for the protection of aquatic life (CCME 1999) were given.

102

Supplemental Figure 3.2 The sedimentary chironomid remains retrieved from Pigeon Pond as concentrations (head capsules per gram dry sediment), as well as the absolute total number of head capsules counted per interval (Total HC), the overall number of head capsules per gram of dry sediment (HC/g dry sed). Sedimentary concentrations (mg/g) of chlorophyll-a inferred by visual reflectance spectroscopy (VRS-chla) and ratios of stable nitrogen isotopes (δ15N in ‰) were first published in Stewart et al. (2015).

103

Supplemental Figure 3.3 The first two axes of a Principal Component Analysis (PCA) of the Helinger-transformed chironomid species data of sediment cores from two sewage ponds (R-12 and R-13) and a reference pond (R-2) near Resolute Bay, Cornwallis Island (NU) modified from Stewart et al. (2014). The temporal dissolved oxygen concentration (mg/L) and saturation (%) profiles of B. R-12 and C. R-13 taken in each pond in July of 2014.

104

Chapter 4

A paleoenvironmental study tracking eutrophication, metal pollution, and

climate change in Niven Lake (NT), Yellowknife’s first sewage lagoon

Formatted for:

Stewart, E.M., Hargan, K.E., Sivarajah, B., Kimpe, L.E., Blais, J.M., and Smol, J.P. In review. A paleoenvironmental study tracking eutrophication, metal pollution, and climate change in Niven Lake

(NT), Yellowknife’s first sewage lagoon. Arctic: submission number 17-134.

Abstract

Niven Lake was the City of Yellowknife’s (Northwest Territories, Canada) first wastewater disposal site, receiving domestic sewage for more than 30 years. Here, we used a high-resolution sediment core to track past sewage inputs to Niven Lake by comparing changes in sedimentary sterols and three diagnostic ratios for human fecal contamination, as well as biological assemblages and overall lake production, with the known history of sewage inputs to the lake from 1948 to 1981. Coprostanol, often considered the best indicator of human fecal contamination, increased by ~10% between 7.5 cm and 5 cm

(~1950 to 1981). Trends in diagnostic sterol ratios also tracked sewage contamination, best represented by coprostanol/(coprostanol+5α-cholestanol). Muted responses in subfossil diatom and chironomid assemblages were noted during the time of sewage inputs, as have been reported in other Arctic sites, as well as in many macrophyte-dominated shallow lakes in general. More marked ecological changes occurred a decade after the end of sewage inputs, in the 1990s, which closely aligned with the warmest years on record for Yellowknife. Recent climate warming may be partly responsible for the more recent expression of limnological symptoms associated with eutrophication, including anoxia and the positive feedback of possible internal phosphorus loading, as indicated by our biological proxies. Changes in the

105

biology of Niven Lake were also indicative of metal pollution, suggesting that the lake has experienced the compounding effects of arsenic contamination from nearby gold mining.

Keywords: sewage lagoon, Arctic, paleolimnology, sterols, stable nitrogen isotopes, diatoms, chironomids, shallow lakes

Introduction

Sewage lagoons, or water bodies used to decontaminate sewage, are an inexpensive and often cost-efficient method of wastewater treatment for communities that cannot make use of large wastewater treatment facilities. This is especially true for northern communities that, due to their remote location and extreme climate, experience greater difficulty implementing more complex waste-management infrastructure. Thus, untreated sewage may be released directly into sewage lagoons, coastal bays, or other freshwater systems. The effectiveness of sewage lagoons in treating the wastewater of Arctic communities has been of interest for decades (e.g. Miyamoto and Heinke, 1979), and is still closely monitored today as new risks and challenges arise (Gunnarsdóttir et al., 2013). Despite this, however, few intensive limnological studies of wastewater-related cultural eutrophication have been undertaken in

Arctic regions, even though populations have been expanding as interest in northern resources has grown in recent decades and will likely continue into the future (Arctic Monitoring and Assessment Programme,

2010). After their useful lifetime as sewage receptacles has passed, abandoned lagoons can be assessed for recovery from eutrophication, as well as from contamination by sewage-related pollutants and pathogens (Squires, 1982; Heinke and Smith, 1986; Ferguson Simek Clark, 1990a, b).

Niven Lake (62.4612° N, 114.3695° W; Figure 4.1) was the primary receptacle of wastewater for the community of Yellowknife from 1948 to 1981, after which time the current lagoon system at Fiddlers

Lake was constructed (Squires, 1982). Throughout the 1950s and into the 1960s, the lake was considered effective in providing what was deemed to be acceptably clean effluent that flowed down a ravine and into “Back Bay,” a portion of Yellowknife Bay in Great Slave Lake (reviewed by Heinke and Smith,

106

1986). However, after territorial government operations were moved from Ottawa to Yellowknife in

1967, the population of Yellowknife more than tripled from ~3 000 to ~10 000 residents by 1977. Niven

Lake became overloaded, and it was determined that the effluent was no longer being sufficiently detoxified (Grainge, 1971; Yamomoto, 1975; Bell et al., 1976). Once Niven Lake was no longer used as a sewage lagoon, studies of the public and environmental health risks associated with it were initiated, and remediation options were considered, since the area around Niven Lake was attractive for residential development (Squires, 1982; Heinke and Smith, 1986; Ferguson Simek Clark, 1990a, b). Due to logistical and cost constraints, the City of Yellowknife opted to allow Niven to recover naturally from sewage inputs, after confirming that this course of remediation posed no serious risk to public or environmental health (Ferguson Simek Clark, 1990a). The Niven Lake area was developed shortly after 2008 and is now in the centre of a prosperous and growing residential district.

For this study, we used the stanols, coprostanol and epicoprostanol, in dated lake sediments for tracing human sewage inputs to Niven Lake over time, as they have been shown to track human fecal contamination (Leeming et al., 1996; Bull et al., 2002; Carreira et al., 2004; Campos et al., 2012). We also explored whether changes in stable nitrogen isotope ratios, which can be enriched by human waste

(Kendall et al., 2007; Vane et al., 2010), were recorded in the sediment profile. In addition, we tracked the limnological responses of the lake biota to past sewage inputs, as well as other environmental stressors, using sub-fossil algal and invertebrate assemblages. Finally, we estimated past trends of overall lake production using spectrally-inferred sedimentary chlorophyll-a concentrations (Michelutti and Smol,

2016).

In higher mammals, cholesterol is converted in the gut by anaerobic bacteria to coprostanol, which once in the aquatic environment can undergo further microbial reduction to yield epicoprostanol

(Bull et al., 2002). Furthermore, approximately 10 or more diagnostic ratios using sterols and stanols have been published for identifying human fecal contamination in different media, including sediments, river and lake water, marine water, and landfill leachate (reviewed in Furtula et al., 2012). Here we use a

107

combination of sterol ratios to provide a robust analysis of the extent of sewage contamination in Niven

Lake. There is a relatively limited understanding of the conditions that favour the preservation of sterols and stanols in sediments (Korosi et al., 2015), with the exception that anaerobic conditions may slow degradation processes and that degradation may also be especially restricted in cold climates (Ogura et al., 1990). Our application of multiple ratios may inform future studies using sterols as a proxy for sewage contamination, especially in Arctic lakes, by comparing temporal trends in the ratios with the known history of sewage discharge into Niven Lake, as well as information about the limnological conditions under which they were likely deposited.

Subfossil diatom and chironomid assemblages have commonly been examined for studies of eutrophication, as diatom species are responsive to changing nutrient (phosphorus) conditions (Hall and

Smol, 2010), and chironomid assemblages often shift with altered oxygen dynamics, specifically the decrease in bottom-water dissolved oxygen that typically accompanies eutrophication (Little et al., 2000;

Luoto and Salonen, 2010). However, the response of subarctic and Arctic lakes to excess nutrient inputs can be complex and less predictable than in temperate latitudes, with a notable absence of large cyanobacterial blooms (Schindler et al., 1974), a more subdued and delayed response in diatom taxa that favours periphytic types (Douglas and Smol, 2000), and increases in chironomid production only after climate-mediated reductions in ice cover (Antoniades et al., 2011). The relatively muted responses to cultural eutrophication in northern lakes has been attributed to the dominant influence of prolonged ice cover and short growing seasons (Michelutti et al., 2007; Smol and Douglas, 2007), reinforcing the idea that the extreme climate restricts many aspects of environmental change in Arctic freshwaters.

Tracking Niven Lake’s environmental past using sediment archives allowed us to assess eutrophication and contamination, as well as possible trajectories of recovery, from a continuous and time-integrated perspective. In addition, the possibility of multiple environmental stressors acting on

Niven Lake was considered. Giant Mine and Con Mine were gold mines that operated near Yellowknife’s city limits between 1948-2004 and 1938-2003, respectively. Both mines released copious amounts of

108

arsenic trioxide dust from the roasting of gold-bearing arsenopyrite ore until roasting activities ceased in

1999 (Hocking et al., 1978; Indian and Northern Affairs Canada, 2007). Using the sediment record, it was determined that the resultant metal pollution in a nearby lake caused the disappearance of keystone invertebrate grazers and planktonic algal groups from the time of peak operational activities, with little- to-no biological recovery following closure of the mines (Thienpont et al., 2016). Therefore, we attempted to disentangle the compounding effects of arsenic contamination on diatoms and chironomids from the effects of sewage inputs in Niven Lake, as well as the additional stressor of recent climate warming known to affect this area (Schindler and Smol, 2006; Coleman et al., 2015).

Since only sparse qualitative observations of Niven Lake’s pre-impact conditions exist, these paleoenvironmental data will position the possible recovery of Niven Lake into important historical context. We hypothesized that the relatively short growing season and shallow depth of Niven Lake were important in determining the response of diatoms and chironomids to eutrophication, and the compounding effects of recent warming and Giant Mine may result in a unique response not often recorded elsewhere in studies of Arctic eutrophication.

Site description Niven Lake (62.4612° N, 114.3695° W) in Yellowknife (NT) is approximately 0.07 km2 in open- water surface area with a maximum depth of just over ~1.5 m, as measured in July of 2015 (Figure 4.1).

Niven is currently in the middle of a residential subdivision and is surrounded by a walking path separated from the open water by some trees and 5-10 m of thick marshland. Niven’s main basin is >95% covered by macrophytes from the lake bottom to the water surface, making it entirely littoral with very little pelagic habitat and only some small pockets of bare sediment. During the July 2015 sampling, Niven also had a bloom of filamentous green algae floating throughout the lake surface. Niven Lake currently freezes nearly to the bottom in the winter and is fishless. The lake also currently hosts many migrating birds in the summer months on their way to Arctic breeding grounds, and some ducks nest in the marshes around the lake.

109

Niven Lake drains a relatively large catchment (compared to its surface area) of almost 4 km2, including Frame Lake from the west and runoff from the northern part of Yellowknife’s downtown from the south. Niven flowed into Back Bay (a portion of Yellowknife Bay of Great Slave Lake between old town and the mainland to the west) through a natural ravine at its northeastern edge that was modified when it became a sewage lagoon in 1948 (Heinke and Smith, 1986). After a narrow reach, a dam was implemented in the first few years of Niven’s use as a sewage lagoon to raise water levels in order to increase holding capacity, but the dam was later removed after sewage inputs ended in 1981. The sewage inlet built in 1948 consisted of an utilidor that carried wastewater out into the lake from the approximate middle of the southeastern edge. In 1963, in response to deteriorating water quality and impending anoxia, a primary cell (32 × 18 × 2.4 m) was built between Niven Lake and the sewage inlet in order to allow sludge to settle with clear overflow to the main basin (Heinke and Smith, 1986). Niven Lake is in close proximity to both historically-operated gold mines in Yellowknife, with Giant Mine ~4 km to the north and Con Mine ~3 km to the south (Figure 4.1).

Methods

Water chemistry On 23 July 2015, water samples were taken from just below the surface of Niven Lake at the approximate centre of the basin using Nalgene bottles. Surface water pH and specific conductance were also measured on site using an YSI probe. Water chemistry parameters were measured at the Taiga

Environmental Laboratory in Yellowknife using the standard US Environmental Protection Agency

Methods and the Standard Methods for the Examination of Water and Wastewater (American Public

Health Association, 2005). Dissolved oxygen, specific conductance, and water temperature profiles were also taken at approximately 0 m (just below the surface), 0.5 m, and 0.9 m (near the sediment-water interface) using a YSI meter on 23 July 2015.

+ Ammonium (NH4 ) concentrations for the 2016 data set were estimated using available parameters and compared with ammonium-nitrogen measures from Ferguson Simek Clark (1990a) to 110

estimate lake recovery from sewage inputs. Nitrate plus nitrite concentrations were subtracted from total

+ dissolved nitrogen to yield total Kjeldahl nitrogen, from which NH4 can be calculated by subtracting dissolved organic nitrogen (DON). DON was estimated using measured dissolved organic carbon (DOC) concentrations and the range of DOC:DON ratios common in rivers (8-41, with a mean of 20) given by

Wetzel (2001). This yielded an estimate of possible recovery based on Niven Lake’s current nitrogen levels, since total nitrogen concentrations were not given in Ferguson Simek Clark (1990a) and ammonium concentrations were not measured in our data set.

Sediment sampling and dating A 30-cm sediment core was retrieved from Niven Lake from the centre of the southwestern half of the basin on 23 July 2015 using a UWITEC© gravity corer (Uwitec, Mondsee, Austria). The core was extruded into 0.5-cm intervals using a modified Glew (1988) extruder. Subsamples of every second centimetre from 0-20 cm were freeze dried and used for 210Pb dating with an Ortec high purity

Germanium gamma spectrometer (Oak Ridge, TN, USA) at the University of Ottawa. The resultant radioactivity profiles were developed into a chronology using the Constant Rate of Supply (CRS) model

(Appleby, 2001) with ScienTissiME software (Barry’s Bay, Ontario). 210Pb ages were verified using the circa 1963 peak in 137Cs from the height of atomic bomb testing.

Sterols and stanols

Sedimentary sterols and stanols were examined every one cm until 12 cm, and then every two cm until 20 cm to fully capture pre-sewage conditions according to the 210Pb dates. Analytical methods were modified from Birk et al. (2010) and Cheng et al. (2016). Freeze-dried sediment (∼0.1 g dry weight) samples were sonicated with 10 mL dichloromethane for 10 minutes with activated cleaned copper, repeated three times. All extracts were combined and concentrated to 1.0 mL under a gentle flow of nitrogen at room temperature. The concentrated extract was transferred to a 6-mL LC-Si SPE column

(liquid chromatography solid phase extraction), which was preconditioned with 6 mL of dichloromethane.

The SPE columns were eluted with 30 mL dichloromethane for cleanup, which was concentrated to 1.0 111

mL under nitrogen. For derivatization, the eluted sterols and stanols were dried completely under a gentle flow of nitrogen, before adding 100 µL 99:1 BSTFA+TMCS (N,O-Bis(trimethylsilyl) trifluoroacetamide and Trimethylchlorosilane), then heated at 60°C for 2 hours. We added 900 μL of toluene to the derivatized samples, and 10 μL of an internal standard. Sterols and stanols were quantified by GC-MSD

(gas chromatography mass selective detector) with a capillary column of Agilent 19091J-433 HP-5 5% phenyl methyl siloxane, and details of the oven temperature ramp and times can be found in Cheng et al.

(2016). For quality control, all GC-MSD results were amended to the internal standard d14-p- terphenyl done by MSD ChemStation D.02.00.275.

For every five samples, an experimental blank was run simultaneously. Cholesterol and stigmastanol were detected in some blanks with concentrations no more than 10% of samples. The blank values were subtracted from corresponding samples. Limit of quantification was defined as a signal-to- noise ratio of three. Signals below that ratio were regarded as not quantified and were discarded. The nine sterols measured were coprostanol (5β-cholestan-3β-ol), epicoprostanol (5β-cholestan-3α-ol), coprostanone (5β-cholestan-3-one), cholesterol (cholest-5-en-3-ol), 5α-cholestanol (5α-cholestan-3β -ol), cholestanone (5α-cholestan-3-one), stigmastanol (5α-stigmastan-3β-ol), and sitosterol (β-sitosterol).

Diagnostic sterol ratios and their associated threshold values for determining human fecal contamination are described in Table 4.1 (reviewed by Furtula et al., 2012), where values between these criteria cannot be readily apportioned on the basis of this ratio alone. Ratios for the Niven Lake sediment core were determined using sterol concentrations, and the sterol changes through time were described in percent relative abundances of the total sterol concentration.

Elemental analysis and stable nitrogen isotopes Elemental and isotopic analyses of all samples were performed at the G.G. Hatch Stable Isotope

Laboratory at the University of Ottawa, Ottawa, ON. Samples and standards were submitted to an elemental analysis (EA) to determine the elemental composition of carbon and nitrogen using the CE

EA1110 Elemental Analyzer, with the detailed methods outlined in Brazeau et al. (2013). Amounts

112

needed for the isotopic analyses were based on the results of the EA. Sediments were weighed accordingly into tin capsules (~3 mg) with 5 mg tungsten oxide (WO3). Calibrated internal standards were prepared as a reference with every batch of samples. Samples were combusted at 1800°C in an elemental analyzer (EA 1110, CE Instruments, Italy) and the gases produced were run through an isotope ratio mass spectrometer (Delta-Plus Advantage IRMS, ThermoFinnigan, Germany) using a Conflo III Interface

(Thermo, Germany). The data were reported using Delta notation (δ) in per mil (‰), which is defined as δ

= [(Rx -Rstd)/Rstd]*1000, where R is the ratio of the abundance of the heavy to light isotope, “x” denotes sample, and “std” is an abbreviation for standard. The routine precision of the analyses was 0.20 ‰.

Spectrally-inferred chlorophyll a Sedimentary concentrations of chlorophyll-a, estimated using visual reflectance spectroscopy

(VRS), were used to track trends in overall primary production of Niven Lake and thus its eutrophication during use as a sewage lagoon, as well as possible recovery since 1981. The methodology generally followed Wolfe et al. (2006), as reviewed by Michelutti and Smol (2016). Freeze-dried sediments were sieved using a 120-µm mesh and placed in cuvettes to be run for absorbance using a FOSS NIR System

Model 6500 spectrometer. The spectra in the range of 650-700 nm were analyzed for the area under the curve using RStudio© (version 1.0.136) every one cm from 0 to 15 cm, which approximates the concentration of chlorophyll-a and its main diagenetic products in mg/g (hereafter VRS-chla, Michelutti and Smol, 2016).

Diatoms and chironomids Approximately 0.2 g of wet sediments of every one cm from 0 to 15 cm were digested for subfossil diatom assemblages using standard methods described by Battarbee et al. (2001). Subsamples were treated with a 1:1 molar ratio solution of concentrated sulfuric acid (H2SO4) and nitric acid (HNO3) and heated in a ~80ºC water bath for two hours. Concentrated acid in the diatom slurries was aspirated over several days to reach a neutral pH, and various concentrations of the neutral slurry were plated and mounted on slides using Naphrax. A minimum of 400 diatom valves were enumerated from each sample 113

at 1000-x magnification using primarily Krammer and Lange-Bertalot (1986-1991). Zones of the diatom assemblage were determined using the cluster analysis technique, constrained incremental sum of squares

(CONISS, Grimm, 1987), and the number of important groupings were identified using broken stick analysis (Bennett, 1996). CONISS analyses were performed using the rioja package (Juggins, 2015) in

RStudio© (version 1.0.136).

Subfossil chironomid assemblages were analyzed on the same sediment intervals as were the diatoms in order to capture pre- and post-sewage conditions. Following the methods described in Walker

(2001), ~0.02-0.05 g of wet sediment were deflocculated in 80 mL of 5% potassium hydroxide (KOH) solution at ~70ºC for 20-30 minutes and then rinsed on a 100-µm sieve into a beaker using deionized water. Sediments were picked for chironomid head capsules and Chaoborus mandibles, which were placed on slides that were mounted permanently using Entellan. A minimum of 50 head capsules, as suggested by Quinlan and Smol (2001), was exceeded for all intervals. Chironomid head capsules were identified to the species-type where possible using Larocque and Rolland (2006), Brooks et al. (2007), and Andersen et al. (2013). Taxa were plotted using C2 (version 1.7.4) and arranged by tribe or subtribe.

CONISS analyses were also performed on the chironomid assemblage profile, but groups were found to be not important by broken stick analysis (Bennett, 1996).

Results

Water chemistry Niven Lake was eutrophic in July of 2015 and 2016 (Table 4.2) with an average unfiltered total phosphorus (TP-u) of 96 µg/L and average total filtered nitrogen (TN-f) value of 1.72 mg/L. Water concentrations of chlorophyll-a were measured in July 2015 with a value of 2.35 µg/L, demonstrating the lake’s productive status. The average pH of both sampling times was 9.9, and the average specific conductance was 565 microSiemens per cm (µS/cm), which was also reflected in the elevated concentrations of major ions in the water (Table 4.2). Finally, arsenic concentrations were notably high, with values between 40-50 µg/L in 2015 and 2016 compared to the drinking water quality guideline of 10 114

µg/L (Health Canada, 2006) and the 5-µg/L guideline for the protection of aquatic life (Canadian Council of Ministers of the Environment, 2001). Niven Lake drains Frame Lake to the west, which also appeared to be relatively productive in 2014 (Table 4.2).

The dissolved oxygen (DO) depth profiles from Niven Lake on 23 July 2015 showed a gradient from supersaturation at the water surface (DO = 10.3 mg/L, 105% saturation) to saturated at mid-depth

0.5 m below surface (DO = 7.7 mg/L, 85% saturation) to hypoxic near the sediment-water interface 0.9 m below the surface (DO = 1.6 mg/L, 18% saturation) (Table 4.3). Specific conductance also showed a differential from the surface downwards, increasing by 200 µS/cm from 266 µS/cm to 466 µS/cm, with the bottom waters approximately equal to mid-depth (0.5 m) values. A gradual change in temperature was also notable in Niven Lake with temperatures of 18℃, 17℃, and 16℃ at 0, 0.5, and 0.9 m depth, respectively.

Recent water chemistry for Niven Lake was compared to data from an environmental assessment undertaken by Ferguson Simek Clark (1990a, b) in the summer of 1990 that determined the public and environmental health risks posed by the lake to the community of Yellowknife approximately ten years after sewage was last discharged. Measures of nitrogen were reported as ammonium-nitrogen (Ferguson

Simek Clark, 1990a, b) and, though we did not have measures of ammonium-N in our 2015-2016 water chemistry, we estimated the possible range of ammonium-N concentrations from our total nitrogen concentrations using nitrate plus nitrite concentrations in conjunction with estimates of DON (dissolved organic nitrogen) calculated using our dissolved organic carbon (DOC) measurements and the range of

DOC:DON ratios for rivers (8 - 41, mean 20) given by Wetzel (2001). Furthermore, the TP concentrations from Ferguson Simek Clark (1990a, b) were extremely high (~3000 µg/L), even for Arctic sewage lagoons (Douglas and Smol, 2000), and therefore were likely unrepresentative of the true concentrations in the water column. We believe that arsenic interferes with the measurement of phosphorus in these systems, as arsenate and phosphate have the same colorimetric effect, yielding TP values that are artificially high when measured using colorimetric assays. For this reason, our TP

115

concentrations for Niven Lake were determined by ICP-MS (inductively coupled plasma mass spectroscopy) in 2015 and 2016, and the concentrations of phosphorus in Frame Lake from 2014 that were measured using colorimetric methods (Table 4.2) were not used. We hypothesize that the TP values of Niven Lake reported in Ferguson Simek Clark (1990a, b) likely suffered from this effect as well, and they were therefore not used for comparison with our 2015-2016 measures.

From a comparison of the ammonium-N concentrations in 1990 and the estimated ammonium-N concentration in 2016, Niven has possibly experienced an ~80-93% decrease (using the maximum and mean possible ratios of DOC:DON = 41 or 20, respectively) in ammonium-N from 10 years since sewage inputs (1990) to 35 years since sewage inputs (2016). The 2016 ammonium-N concentration estimated using the minimum possible DOC:DON of 8 equaled zero, and thus was unlikely to represent Niven’s

+ true water chemistry. In fact, a DOC:DON ratio of ~16 or less yields a 2016 NH4 -N concentration of zero. Water alkalinity (hardness) also decreased by approximately half in Niven Lake from 1990 to

2015/2016, though pH increased from circumneutral or slightly alkaline to very alkaline (with modern values of nearly 10 pH units). Finally, arsenic concentrations in Niven Lake have moderately declined from ~65 µg/L in 1990 to ~40 µg/L in 2015/2016.

Sediment chronology The Niven sediment core was dated using the radioactivities of 210Pb, as well as 214Bi and 214Pb

(averaged to give supported 210Pb activity), and the circa-1963 peak in 137Cs was used as an independent dating marker (Supplemental Figure 4.1). Total 210Pb activity was low as is common for sediments from northern regions (Wolfe et al., 2004), however there was good exponential decay observed over the core, and the CRS model (Appleby, 2001) was applied to determine the core chronology (Supplemental Figure

4.1). The peak in 137Cs activity was difficult to determine, as is common in highly organic sediments

(Blais et al., 1995). However, the 137Cs peak modeled by the ScienTissiME software (Barry’s Bay,

Ontario) had good agreement with CRS-derived dates, giving a date of 1979 ± 12 years for 6.25 cm, corresponding to a CRS-date of 1971 ± 9 years. The CRS-derived sedimentation rate sharply increased

116

from 8.25 to 6.25 cm (1939 to 1971) and then stayed approximately stable for the upper 6.25 cm of the core (Supplemental Figure 4.1). The increase in sedimentation rate had good agreement (within error) with the period of known sewage inputs to Niven (1948 to 1981). Therefore, based on our chronology the sediment intervals that represent the duration of Niven Lake’s use as a sewage lagoon are from approximately 7.5 cm (~1950) to 5 cm (1981).

Sterols and stable nitrogen isotopes Both the pre- and post-industrial sterol and stanol profiles of Niven Lake were dominated by the quantified plant sterols: sitosterol and stigmastanol (Figure 4.2). Over time, and particularly marked at

~1999 (3 cm), there was a switch in the dominance of these plant sterols from stigmastanol to sitosterol.

Prior to ~1950 (7.5 cm), the sterol and stanol sedimentary profile was relatively consistent, represented on average by 13% coprostanol, 2% epicoprostanol, 3% coprostanone, 7% cholesterol, 14% 5α-cholestanol,

1% cholestanone, 15% sitosterol and 45% stigmastanol (Figure 4.2). At ~1971 (6 cm), epicoprostanol and coprostanol both increased and represented ~22.5% of the total sterols, with coprostanol increasing from

10% to ~18%. This peak was short and declines in both stanols occurred by ~1981 (5 cm). Additionally, coprostanone peaked at 1971 reaching its maximum abundance of 4.2% through the sedimentary profile.

A correction to sedimentary total carbon content did not alter these trends in human fecal markers.

The ratio of coprostanol/(coprostanol+5α-cholestanol) peaked at ~1971 (6 cm) reaching a maximum value of 0.56 (Figure 4.2). The mean for this ratio prior to sewage disposal in Niven Lake was

0.48 and after sewage diversion in the 1980s (5 cm) was 0.36, with all values remaining between the indicator values for the presence or absence of human fecal contamination (>0.7 and <0.3, respectively,

Table 4.1). The coprostanol/cholesterol ratio was more variable through time, with ratios varying between

1.5 and 3.05, and all values prior to 4 cm (~1999) were indicative of human fecal contamination (>0.5,

Table 4.1). However, from a coprostanol/cholesterol ratio equal to ~1.5 at 10 cm, there was a peak in values at 6 cm, consistent with peaks in other human fecal marker ratios (Figure 4.2). Only the most modern sediments at 3 cm, 1 cm, and 0 cm had coprostanol/cholesterol values below the lower threshold

117

indicating no human fecal contamination (<0.3, Table 4.1). The epicoprostanol/coprostanol diagnostic ratio is less clear in capturing the period of sewage input to Niven Lake, with values of 0.15 in the pre- sewage part of the core and increasing to almost 0.2 during sewage inputs. Ratios <0.2 indicate human fecal contamination (Table 4.1), and values exceeded 0.2 at ~1991 (4 cm), with a sharp spike to just over

0.3 and a decrease in the upper sediment intervals back to a value of ~0.2-0.25Stable isotopes of nitrogen

(δ15N) increased gradually across the entire core from the bottom (~1.5‰) to the surface (~4‰), with no notable changes during the period of wastewater inputs to Niven Lake (Figure 4.2). This trend did not coincide with trends in the relative abundances of animal sterols, but was more similar to the gradual increase of the phytosterol, sitosterol.

Diatoms and chironomids The diatom assemblage of Niven Lake was characterized by two distinct zones that are separated by the point of greatest turnover at 3-4 cm depth or approximately the mid-1990s, as determined using

CONISS and broken stick analysis (Figure 4.3A). However, changes in diatom relative abundances begin as early as 7-8 cm (~1940s). The earliest assemblage (8-15 cm) can be considered pre-impact, as this portion of the core was not affected by mining or municipal waste disposal. The pre-impact diatom assemblage of Niven Lake was co-dominated by the cosmopolitan species Achnanthidium minutissimum, as well as the epiphytic species Brachysira neoexilis and Encyonopsis microcephala. Beginning around 8 cm, some of these taxa and some of the rare species began to show subtle relative decreases, and by 3 to 4 cm, all of these species virtually disappeared from the record. Navicula taxa remained stable for the entire record.

Simultaneously, at ~8 cm, several species appeared (or appeared more commonly) at low relative abundances in the diatom assemblage, including Fragilaria mesolepta, Nitzschia amphibia, Eolimna minima, Sellaphora seminulum, Planothidium lanceolata, and Stephanodiscus species (primarily S. hantzschii). Most of these species dramatically increased after the 3-4 cm depth and became the dominant species that make up the second zone of the diatom stratigraphy from 3-4 cm (the mid-1990s) to the

118

surface (2015). For example, Fragilaria mesolepta reached 40% relative abundance in the topmost interval. Stephanodiscus species had a peak relative abundance of ~5% in the 3-cm depth interval and decreased again thereafter.

Changes in sedimentary VRS-chla concentrations were subtle throughout the core (Figure 4.3A).

In the pre-impact sediments, VRS-chla concentrations maintained values around ~0.095 mg/g at 15 cm and slowly increased to ~0.11 mg/g by the approximate time of the onset of sewage inputs around 7.5 cm, or ~1950 (Figure 4.3A). This slow increase continues through the period of sewage inputs, with the highest values throughout the entire record occurring at ~5 cm, or 1981, at the end of sewage discharge into Niven Lake. Subsequently, VRS-chla concentrations decreased back to values of ~0.095 mg/g by 3.5 cm (early 2000s) and then remained stable until 1.5 cm (~2008-2009) at which point a secondary increase to just over 0.1 mg/g was noted in the most recent sediments.

The chironomid assemblage of Niven Lake was not dominated by any particular taxon, but instead had low abundances (~5-10%) of 21 species types or groups (Figure 4.3B). CONISS determinations showed that the largest groupings of the chironomid assemblage occurred between 3 and 4 cm, which was the same for the diatom assemblage. However, broken stick analysis (Bennett, 1996) indicated that groups in the chironomid assemblage were not more probable than random distributions, highlighting the subtle and gradual nature of any visible changes in the chironomid assemblage compared to the drastic changes seen in the diatoms. The most notable changes were decreases in the Tanytarsini by

3-4 cm (1990s), especially the cold-indicator, Micropsectra insignobilous-type. At the same time, an increase in Chironomus plumosus-type to 5% relative abundance and a moderate increase in

Psectrocladius sordidellus/psilopterus-type occurred. The abundance of chironomid remains in the core was calculated as head capsules per gram of dry sediment (HC/g dry sed) and gave an idea of the overall production of chironomids in Niven Lake through time. Chironomids were extremely abundant in the

Niven Lake sediment core with ~3000 head capsules in 1 gram of dry sediment. A notable increase in

119

chironomid production from ~3000 HC/g dry sed to ~5000 HC/g dry sed occurred at 6 cm (1969), followed by a sharp decline back to initial concentrations.

Discussion

Tracking sewage using sterols and δ15N The phytogenic sterol and stanol compounds, sitosterol and stigmastanol (10-20% and 50-60%, respectively), dominated the early sedimentary record (Figure 4.2), indicating that Niven Lake was a naturally productive system before it was used for sewage disposal from 1948 to 1981. Surprisingly, coprostanol comprised on average 13% of the measured sterol composition of Niven’s presumed pre- impact sediments (before 7.5 cm, or ~1950 according to our 210Pb dates). This may be considered high for pristine and uncontaminated sediments, where in situ production of coprostanol has been shown to be negligible (Nishimura, 1982; Leeming et al., 1996). Coprostanol is a product of cholesterol formed in the intestines, and humans have been found to excrete approximately 10 times as much coprostanol as other animals (Leeming et al., 1996), making large increases in coprostanol an effective indicator of human fecal contamination in some instances. In addition, the values of two of the three diagnostic sterol ratios

(coprostanol/cholesterol and epicoprostanol/coprostanol) also suggest that humans were impacting Niven

Lake in the pre-sewage sediments (see Table 4.1 for threshold criteria). Our sedimentary sterol evidence of early human impact may be corroborated by the history of the Wedeleh Yellowknives Dene First

Nation, who documented the occupation of the area around Great Slave Lake by the T’satsaot’ine people for centuries before Yellowknife became a mining town in the 1930s (Wedeleh Yellowknives Dene,

1997).

The duration of sewage inputs to Niven Lake was also possibly tracked by sedimentary sterols and stanols (Figure 4.2). The 8% increase in the relative importance of coprostanol in Niven’s sediments

(to ~18% of the overall composition) suggests that sewage contamination occurred between 7.5 and 5 cm

(the 1950s to the 1980s), which coincides with the main use of the lake as the primary (and, until 1975, the only) sewage lagoon of the City of Yellowknife from 1948 to 1981 (Heinke and Smith, 1986). 120

Smaller increases (1-2%) in coprostanone and epicoprostanol (Figure 4.2) also occurred at this time, both of which are naturally reduced under low dissolved oxygen and thus were unsurprising as hypoxia had been reported in Niven Lake during the sewage era (Heinke and Smith, 1986). Decreases in the relative abundances of coprostanone, coprostanol, and epicoprostanol occurred by 5 cm (~1981), likely marking the cessation of sewage inputs to Niven Lake.

The ratio of coprostanol/(coprostanol+5α-cholestanol) best recorded fecal contamination of Niven

Lake, peaking in value at ~1971 (6 cm) and decreasing thereafter. The increase coincided with the height of sewage disposal to Niven Lake (Heinke and Smith, 1986), and the decrease may be indicative of the distribution of 50% of the sewage load into nearby Kam Lake from 1975 to 1981 (Squires, 1982). It should be noted that, throughout the sediment record, coprostanol/(coprostanol+5α-cholestanol) values fell between the threshold values considered ambiguous for determining human fecal matter contamination (Grimalt et al., 1990; Vane et al., 2010). This is likely due to the similar concentrations of coprostanol and 5α-cholestanol in the sediments, yet the fairly stable relative abundance of 5α-cholestanol during sewage inputs and the nearly 10% increase in coprostanol, signifies that the coprostanol/(coprostanol+5α-cholestanol) ratio still likely tracked sewage inputs to Niven Lake.

Changes in the other two sterol ratios examined were less clear, and the lack of changes in these ratios may owe to variable concentrations of cholesterol in the case of the coprostanol/cholesterol ratio and to increases in both epicoprostanol and coprostanol in the case of the epicoprostanol/coprostanol ratio. As cholesterol is the most ubiquitous animal sterol and may come from various sources (including birds, Cheng et al., 2016), this ratio may be driven by other aspects of the environment in and around

Niven Lake. The epicoprostanol/coprostanol ratio has also traditionally been used to indicate either the degree of treatment of sewage or its age in the environment (Mudge and Ball, 2006), as coprostanol can be reduced to epicoprostanol in anoxic conditions over time. Thus, sediments with combined high coprostanol/cholesterol and low epicoprostanol/coprostanol ratios should be a strong indication of active high sewage input, and such a combination of ratios peaks at 6 cm (~1971) in the sediments. Therefore,

121

overall, our analysis of sedimentary sterols in Niven Lake highlights that it is important to consider both relative changes in different sterol compounds, as well as changes in multiple diagnostic ratios, as much is still unknown about the dynamics of sterol compounds under various conditions in lakes, especially in northern regions (Korosi et al., 2015).

Stable nitrogen isotopes did not track sewage inputs in Niven Lake, showing no apparent trends during the period of sewage inputs, even though it has been shown to track sewage contamination in some freshwater systems (Vane et al., 2010). The δ15N of Niven Lake’s sediments remained low with values common in pristine and non-impacted sediments (Blais et al., 2005). Even though animal or sewage waste can have a relatively high δ15N signature of >10‰ (Heaton, 1986), overall municipal wastewater can have a complex δ15N signal because human excrement makes up only a component of the entire discharge, which also includes domestic waste, food scraps, etc. The δ15N in a lake is also mediated by primary production and other aspects of the N-cycle (Leng et al., 2005), which may drastically change with eutrophication. The unresponsive sedimentary δ15N suggests that the use of sterol-based proxies may be more effective in tracking human waste inputs. However, we acknowledge that the sterols signal in Niven

Lake was also muted, which may relate to the construction of the primary cell at the sewage inflow in

1963 (Heinke and Smith, 1986), where solids settled out with a relatively clear flow of water to the lake.

As sterols are hydrophobic and will rapidly attenuate by binding to organic matter (reviewed by Korosi et al., 2015), it is likely that much of the sterol input may have resided in the sludge of the primary cell.

Effects of sewage, mining, and climate warming on Niven Lake The early sediments of Niven Lake were dominated by high abundances of benthic epiphytic diatoms (Brachysira neoexilis and Encyonopsis microcephala) and macrophyte-associated littoral chironomids (Polypedilum nubeculosum-type, Psectrocladius species, and Dicrotendipes), substantiating the high proportions of phytogenic sterols in this portion of the core (Figure 4.2). Slowly increasing relative abundances of Cocconeis placentula, a reliable diatom indicator of macrophytes and higher total phosphorus (Reavie and Smol, 1997; Vermaire et al., 2011), at the bottom of the core suggested that

122

Niven was a naturally macrophyte-dominated system, which according to our sterol analysis may have been the result of humans in the area (Figure 4.2).

The species characteristic of the pre-impact diatom assemblage in Niven Lake began decreasing in relative importance in the 1940s (7-8 cm), and more substantially by the mid-1990s (3-4 cm) (Figure

4.3A). During the sewage era, the pre-impact diatom assemblage was replaced by taxa characteristic of higher nutrient concentrations and/or heavy organic matter pollution, such as Nitzschia amphibia,

Eolimna minima, Sellaphora seminulum, and Stephanodiscus hantzschii (Kelly et al., 2005). S. hantzschii, the main species comprising Niven’s Stephanodiscus group and a well-established indicator of eutrophication (Hall and Smol, 1992, 2010; Hadley et al., 2010; Reavie and Kireta, 2015) appeared at

~5% relative abundance around the time of sewage inputs and declined thereafter. However, the increase in planktonic Stephanodiscus species in Niven Lake was somewhat muted, as is common for diatoms experiencing eutrophication in the Arctic, where ice cover, a short growing season, and colder temperatures may override the effects of nutrient additions (Douglas and Smol, 2000; Michelutti et al.,

2007; Stewart et al., 2014). In addition, Niven Lake is very shallow and dominated by aquatic macrophytes – a limnological setting where, due to various feedback mechanisms, the more striking eutrophication-related ecological changes that are typically recorded in deeper lakes are often delayed until new thresholds are crossed (Scheffer, 1998). In fact, given Niven Lake’s shallow and macrophyte- dominated nature, the increase in the planktonic S. hantzschii was perhaps more ecologically significant than the percentage data may indicate. Similarly, there were virtually no changes in the chironomid species assemblage during the period of sewage inputs, which has also been noted elsewhere in shallow eutrophic Arctic sites (Stewart et al., 2013, 2014), and only very subtle increases in VRS-chla were detected.

In contrast, the major turnover in the diatom assemblage occurred later in the 1990s a decade after the end of sewage inputs, between 3 and 4 cm (Figure 4.3A), according to our CONISS (Grimm,

1987) and broken stick (Bennett, 1996) analyses. This change consisted of increases in the meso- to

123

eutrophic taxon, Fragilaria mesolepta, as well as hypereutrophic species, N. amphibia, E. minima, and S. seminulum (Kelly et al., 2005). The turnover of the assemblage to these species suggest that Niven Lake became more eutrophic after the cessation of sewage inputs in 1981, a change that was more drastic than those during the onset or height of wastewater discharge. Additional evidence of the post-sewage eutrophication of Niven Lake include recent increases in sedimentary VRS-chla (overall primary production) in the early 2000s, as well as increases in Psectrocladius, a strongly macrophyte-associated chironomid genus, that suggests further expansion of macrophyte growth in Niven Lake from 1991 (4 cm) onwards. Furthermore, modern phosphorus concentrations from Niven Lake in 2015 and 2016 show the lake was still hypereutrophic (Table 4.2), and the current lakewater pH of nearly 10 was elevated compared to the circumneutral pH of Niven Lake in 1990 (Ferguson Simek Clark, 1990a). A higher pH recently compared to 1990 may be a direct result of increased primary production and associated CO2 uptake that would have shifted the pH buffering reaction to the alkaline.

The continued eutrophication of Niven Lake into the 1990s and 2000s could be linked to internal phosphorus loading, which occurs under low oxygen conditions and creates a positive feedback cycle.

The more than 2-fold increase in the hypoxia- and anoxia-tolerant species, Chironomus plumosus-type, by

1991 (4 cm) suggests that oxygen depletion in Niven was likely occurring after the end of sewage inputs.

Our July 2015 water column depth profile of dissolved oxygen demonstrated decreasing concentrations and saturation with depth, where hypoxic conditions (1.6 mg/L) prevailed near the sediment-water interface (Table 4.3). Based on biological oxygen demand (BOD) measurements in the 1960s, Niven

Lake had also previously been reported as on the verge of becoming anaerobic (Heinke and Smith, 1986).

Furthermore, the increase in the anaerobic degradation products of coprostanol, such as epicoprostanol, in the sediment record at this time (Figure 4.2) supports that Niven Lake likely had low oxygen conditions during the height of sewage inputs and in the decades after. As Niven was dominated by macrophyte cover from the sediments to the water surface during sampling, it may be that, despite its shallow depth,

Niven experienced reduced water-column mixing because of the stagnating effect of extensive

124

macrophyte growth throughout the lake, potentially leading to oxygen depletion (and possibly internal phosphorus loading) lower in the water column. Even though eutrophication-related oxygen depletion may be unexpected in a shallow system (unlike in deep lakes), evidence of reduced water-column mixing was also supported by the directional gradients in specific conductance measured through the water column in July 2015.

The response of diatoms and chironomids in the 1990s indicative of post-sewage eutrophication could have been facilitated and/or exacerbated by recent climate warming. The mean annual air temperature (MAAT) of the City of Yellowknife from 1993 to 2015 was consistently above the average

MAAT (n = 47 years) from 1944 to 2015, and an increase in MAAT over this period of time was also

2 significantly described by a positive linear regression (R adj = 0.4, p<0.0001, Supplemental Figure 4.2).

The same was also true for mean summer temperature of Yellowknife with summer temperatures from

1996 onwards exceeding the ~70-year average (n = 58 years) (Supplemental Figure 4.2). Recent climate warming has reduced ice cover in Arctic lakes and prolonged the growing season in sensitive Arctic regions, leading to related changes in lake biota such as diatoms (Douglas et al., 1994; Griffiths et al.,

2017). A longer and warmer growing season for Niven Lake may have promoted lake stagnation (i.e. reduced mixing of dissolved oxygen) and algal production (increasing BOD), exacerbating the cycle of anoxia and possible internal phosphorus loading that may be inferred from our multiple paleo-proxies.

The timing of the major shifts in diatoms and chironomids aligns with the warmest years on record for

Yellowknife, suggesting that climate may play a role in Niven Lake’s recent late-onset symptoms of eutrophication.

To further complicate matters, some biological changes in Niven Lake were also indicative of metal pollution, namely from gold mining and its consequent release of arsenic trioxide dust to the atmosphere (Indian and Northern Affairs Canada, 2007). Niven Lake was reported to be likely affected by gold mining, as the arsenic concentration in the water was elevated at values of ~60 µg/L in 1990

(Ferguson Simek Clark, 1990a). Our modern water chemistry data document legacy pollution from

125

regional gold mining, with total arsenic concentrations of 41-44 µg/L (Table 4.2), which were elevated relative to the guideline for arsenic in water for the protection of aquatic life at 5 µg/L (Canadian Council of Ministers of the Environment, 2001). For Niven Lake, legacy arsenic contamination may also be controlled by the hypoxia/anoxia associated with its use as a sewage lagoon, since arsenic can be released from the sediments under anoxic conditions (Andrade et al., 2010).

Niven Lake was first used as a sewage lagoon during the same year that Giant Mine opened, in

1948. The onset of the increases in E. minima and S. seminulum, which may indicate eutrophication because of their high TP optima (Kelly et al., 2005), may also reflect mining pollution, as both taxa have been previously associated with tolerating metal pollution (Morin et al., 2008). Furthermore, Niven Lake showed decreases in the Tanytarsini, which also declined in Pocket Lake on Giant Mine property, a lake severely impacted by atmospheric metal contamination (Thienpont et al., 2016), as well as in other studies of mining contamination (Ilyashuk et al., 2003; Doig et al., 2015). Tanytarsus species, in particular, have been shown to be intolerant to trace metal contamination (Johnson et al., 1992), all of which were effectively lost to the Niven Lake assemblage by the end of mining operations in 1999 (Figure 4.3B,

Indian and Northern Affairs Canada, 2007). Interestingly, however, many of the Tanytarsus species-types in the Niven core also have high dissolved oxygen optima (Luoto and Salonen, 2010) and/or low temperature optima (Larocque and Rolland, 2006), suggesting that they likely experienced the compounding stressors of not only trace metal pollution, but anoxia and recent climate warming as well.

Finally, the relative increase in Psectrocladius may indicate that macrophyte growth increased in Niven

Lake in the 1990s, but it was also noted to increase and co-dominate the chironomid assemblage of highly impacted Pocket Lake during mining operations (Thienpont et al., 2016). This makes Psectrocladius a likely candidate for tolerating the trace metal pollution of Giant Mine in this region.

Recovery in Niven Lake Some recovery from eutrophication and arsenic contamination is evident in the sediment record and water chemistry of Niven Lake. For example, the comparison of our 2016 estimates of ammonium-N

126

concentration with measures taken in 1990 demonstrated that ammonium-N levels may have decreased by

80-93% from 1990 to 2016 (Table 4.2). However, it must be noted that the 2016 ammonium-N concentrations were not measured in our data set, but were estimated based on measured parameters (TN- f, nitrate + nitrite, DOC) and a range of published DOC:DON (Wetzel, 2001). According to these estimates, the minimum reduction in ammonia was ~80%, but if Niven Lake had a DOC:DON ratio of greater than published maximum of 41 for unpolluted rivers (Wetzel, 2001), recovery could be less.

However, since the 2016 total dissolved nitrogen concentrations (which includes ammonium-N) were lower than ammonium-N concentrations from 1990, Niven Lake has clearly experienced some chemical recovery since the 1990s. The lake also had a decrease in the concentration of arsenic from ~60 to 40

µg/L (Table 4.2). The increase in the diatom F. mesolepta to ~40% relative abundance near the surface of the Niven Lake sediment core may indicate moderate recovery from eutrophication, as this species has a lower TP optimum compared to hypereutrophic species, N. amphibia, E. minima, and S. seminulum

(Kelly et al., 2005), in the years immediately prior (the 1990s). This recovery was also supported by a subtle decrease in the VRS-chla concentrations after 5-cm or ~1981 (Figure 4.3A), tracking a decrease in the overall primary production; though it did again increase in the early 2000s. However, Niven Lake has clearly not returned to pre-impact conditions in the past 35 years since the end of sewage inputs. Here, the interplay of eutrophication, anoxia, and arsenic contamination is evident in the sediment record, and the added “threat multiplier” of climate warming (Smol, 2010) may be a barrier to the full recovery of Niven

Lake from its use as Yellowknife’s first sewage lagoon.

Conclusions Niven Lake’s use as a sewage lagoon for three decades was likely tracked, albeit subtly, in the sediments by the deposition of higher amounts of coprostanol, which is considered the best indicator of human fecal contamination (Korosi et al., 2015). The nearly 10% rise in the relative abundance of coprostanol between 7.5 and 5 cm (~1950 to 1981) aligns with the period of sewage discharge into Niven

Lake according to our 210Pb CRS-derived dates. This increase was also possibly reflected in diagnostic

127

ratios used to determine fecal contamination in water and sediments, best represented by changes in coprostanol/(coprostanol+5α-cholestanol). However, based on previously published thresholds for this diagnostic ratio, our signal was muted and did not exceed the sewage contamination criterion, perhaps in part due to presence of the primary cell, where sterol compounds may have been largely deposited with other solid organic matter.

During the period of sewage inputs, subfossil diatom assemblages appeared to track sewage inputs with subtle increases in planktonic Stephanodiscus species, and in particular the well-known eutrophication indicator, S. hantzschii (Reavie and Kireta, 2015). The chironomid assemblage demonstrated little change during the time of sewage inputs, which together with the subtle shifts in the diatoms, equated to a muted response typical of Arctic systems that are generally controlled by the overarching effect of a short growing season and prolonged ice cover (Douglas and Smol, 2000;

Antoniades et al., 2011). In addition, shallow lakes dominated by aquatic macrophytes, such as Niven lake, are also known to often be resistant to threshold-type changes linked to eutrophication, due to a variety of feedback mechanisms (Scheffer, 1998). However, despite the muted nature of the responses of sterols and biological assemblages in the sediment record, the synchronous changes in our multiple paleo- proxies occurred between the radiometric dates of ~1950-1980, which align with the known period of sewage inputs to Niven Lake.

Interestingly, the greatest period of biotic change in Niven Lake occurred in the 1990s, approximately a decade after sewage inputs ended, aligning with the onset of the warmest years on record for Yellowknife. The nature of the marked changes in diatoms were consistent with lagging symptoms of eutrophication and, together with an increase in the hypoxia-tolerant chironomid, Chironomus plumosus- type, suggest that Niven was experiencing increased anoxia and potential internal phosphorus loading from the sediments. These changes may have been climate-mediated, as a longer growing season and warmer temperatures may have promoted the development of anoxia, and thus the possible positive feedback cycle of internal phosphorus release, through reduced water-column mixing, which was also

128

evident in the directional gradients of our 2015 water column depth profiles of dissolved oxygen and conductance. In fact, despite being shallow, Niven Lake had hypoxic bottom waters in July of 2015.

Finally, changes in both the diatoms and chironomids were also consistent with heavy metal pollution, as

Niven Lake was affected by arsenic contamination from nearby Giant Mine, and the chironomids in particular evinced the combined impact of multiple stressors on Niven Lake. Overall, Niven Lake may be showing some signs of chemical recovery from sewage loading since the early 1990s, but the additional stressor of recent climate warming may complicate biological recovery as the lake experiences a novel climatic regime, as well as legacy metal pollution from gold mining in the region.

Acknowledgements

We would like to thank members of the JMB lab and Jennifer Korosi for field help and the

Cumulative Impacts Monitoring Program (Government of the Northwest Territories) for equipment.

Thank you to Taiga Environmental Laboratories (Yellowknife, NT) for water chemistry analyses. Thank you also to David Jessiman (Government of the Northwest Territories) for providing monitoring reports and other resources concerning Niven Lake. Many thanks to Dr. Peter Dillon (Trent University) for generous advice concerning water chemistry comparisons. This work was supported by grants from the

Natural Sciences and Engineering Research Council of Canada and the Polar Continental Shelf Program awarded to JMB and JPS, as well as fieldwork funding from the Northern Scientific Training Program.

References

American Public Health Association. 2005. Standard methods for the examination of water & wastewater,

21st edition. American Public Health Association Publications.

Andersen, T., Cranston, P.S., and Epler, J.H., eds. 2013. The larvae of Chironomidae (Diptera) of the

Holarctic region - keys and diagnoses. Insect Systematics and Evolution Supplemental 66:1–571.

Andrade, C.F., Jamieson, H.E., Kyser, T.K., Praharaj, T., and Fortin, D. 2010. Biogeochemical redox

cycling of arsenic in mine-impacted lake sediments and coexisting pore waters near Giant Mine,

129

Yellowknife Bay, Canada. Applied Geochemistry 25(2):199–211.

http://dx.doi.org/10.1016/j.apgeochem.2009.11.005

Antoniades, D., N. Michelutti, R. Quinlan, J.M. Blais, S. Bonilla, M.S.V. Douglas, R. Pienitz, J.P. Smol

and W.F. Vincent. 2011. Cultural eutrophication, anoxia, and ecosystem recovery in Meretta

Lake, High Arctic Canada. Limnology and Oceanography 56(2): 639-65.

http://dx.doi.org/10.4319/lo.2011.56.2.0639

Appleby, P.G. 2001. Chronostratigraphic techniques in recent sediments. In: Last, W.M., and Smol, J.P.,

eds. Tracking Environmental Change Using Lake Sediments Volume 1: Basin Analysis, Coring

and Chronological Techniques. Dordrecht: Kluwer Academic Publishers. 172–203.

Arctic Monitoring and Assessment Programme. 2010. Assessment 2007: Oil and gas activities in the

Arctic – Effects and potential effects. Volume 1. Oslo: Arctic Monitoring and Assessment

Programme (AMAP). vii + 429.

Battarbee, R.W., Jones, V.G., Flower, R.J., Cameron, N.G., Bennion, H., Carvalho, L., et al. 2001.

Diatoms. In: Last W.M., and Smol, J.P., eds. Tracking Environmental Change Using Lake

Sediments Volume 3: Terrestrial, Algal, and Siliceous Indicators. Dordrecht: Kluwer Academic

Publishers. 155–202.

Bell, J.B., Zaal, J.F.J., and Vanderpost, J.M. 1976. The bacterial quality of lake waters at Yellowknife,

Northwest Territories. Arctic 29:138–146.

Bennett, K.D. 1996. Determination of the number of zones in a biostratigraphical sequence. New

Phytologist 132(1):155–170. http://dx.doi.org/10.1111/j.1469-8137.1996.tb04521.x

Birk, J.J., Dippold, M., Wiesenberg, G.L.B., and Glaser, B. 2010. Combined quantification of faecal

sterols, stanols, stanones and bile acids in soils and terrestrial sediments by gas chromatography–

mass spectrometry. Journal of Chromatography A 1242:1–10.

http://dx.doi.org/10.1016/j.chroma.2012.04.027

130

Blais, J.M., Kalff, J., Cornett, R.J., and Evans, R.D. 1995. Evaluation of 210Pb dating in lake sediments

using stable Pb, ambrosia pollen, and 137Cs. Journal of Paleolimnology 13(2):169-178.

http://dx.doi.org/10.1007/BF00678105

Blais, J.M., Kimpe, L.E., McMahon, D., Keatley, B.E., Mallory, M.L., Douglas, M.S.V., et al. 2005.

Arctic seabirds transport marine-derived contaminants. Science 309:445.

http://dx.doi.org/10.1126/science.1112658

Brazeau, M.L., Poulain, A.J., Paterson, A.M., Keller, W., Sanei, H., and Blais, J.M. 2013. Recent changes

in mercury deposition and primary productivity inferred from sediments in lakes from the Hudson

Bay Lowlands, Ontario, Canada. Environmental Pollution 173:52–60.

https://doi.org/10.1016/j.envpol.2012.09.017

Brooks, S.J., Langdon, P.G., and Heiri, O. 2007. The Identification and Use of PalaeArctic Chironomidae

Larvae in Palaeoecology. QRA Technical Guide No. 10. London: Quaternary Research

Association.

Bull, I.D., Lockheart, M.J., Elhmmali, M.M., Roberts, D.J., and Evershed, R.P. 2002. The origin of faeces

by means of biomarker detection. Environnent International 27(8):647–654.

https://doi.org/10.1016/S0160-4120(01)00124-6

Campos, V., Fracácio, R., Fraceto, L.F., and Rosa, A.H. 2012. Fecal sterols in estuarine sediments as

markers of sewage contamination in the Cubatão area, São Paulo, Brazil. Aquatic Geochemistry

18(5):433−443. https://doi.org/10.1007/s10498-012-9167-2

Canadian Council of Ministers of the Environment. 2001. Canadian water quality guidelines for the

protection of aquatic life: arsenic. Canadian Environmental Quality Guidelines.

Carreira, R.S., Wagener, A.L.R., and Readman, J.W. 2004. Sterols as markers of sewage contamination in

a tropical urban estuary (Guanabara Bay, Brazil): space−time variations. Estuarine, Coastal Shelf

Science 60(4):587−598. https://doi.org/10.1016/j.ecss.2004.02.014

131

Cheng, W., Sun, L., Kimpe, L.E., Mallory, M.L., Smol, J.P., Gallant, L.R., Li, J., and Blais, J.M. 2016.

Sterols and stanols preserved in pond sediments track seabird biovectors in a High Arctic

environment. Environmental Science and Technology 50(17):9351−9360.

https://doi.org/10.1021/acs.est.6b02767

Coleman, K., Palmer, M.J., Korosi, J.B., Kokelj, S.V., Jackson, K., Hargan, K.E., Courtney Mustaphi,

C.J., et al. 2015. Tracking the impacts of recent warming and thaw of permafrost peatlands on

aquatic ecosystems: a multi-proxy approach using remote sensing and lake sediments. Boreal

Environmental Research 20(3):363–377.

Doig, L.E., Schiffer, S.T., and Liber, K. 2015 Reconstructing the ecological impacts of eight decades of

mining, metallurgical, and municipal activities on a small boreal lake in northern Canada.

Integrated Environmental Assessment and Management 11(3):490–501.

https://doi.org/10.1002/ieam.1616

Douglas, M.S.V., and Smol, J.P. 2000. Eutrophication and recovery in the High Arctic: Meretta Lake

(Cornwallis Island, Nunavut, Canada) revisited. Hydrobiologia 431(2-3):193–204.

https://doi.org/10.1023/A:1004000530997

Douglas, M.SV., Smol, J.P., and Blake Jr, W. 1994. Marked post-18th century environmental changes in

High-Arctic ecosystems. Science 266:416–419. https://doi.org/10.1126/science.266.5184.416

Fattore, E., Benfenati, E., Marelli, R., Cools, E., and Ranelli, R. 1996. Sterols in sediment samples from

Venice Lagoon, Italy. Chemosphere 33(12):2383–2393. https://doi.org/10.1016/S0045-

6535(96)00340-2

Ferguson Simek Clark. 1990a. Niven Lake environmental survey City of Yellowknife, NWT. Final

Report 89-1380. Yellowknife, Northwest Territories: Ferguson Simek Clark Engineers &

Architects.

132

Ferguson Simek Clark. 1990b. Niven Lake restoration remediation alternatives, City of Yellowknife,

NWT. Report 89-1380. Yellowknife, Northwest Territories: Ferguson Simek Clark Engineers &

Architects.

Froehner, S., Martins, R.F., and Errera, M.R. 2009. Assessment of fecal sterols in Barigui River

sediments in Curitiba, Brazil. Environmental Monitoring and Assessment 157(1-4):591–600.

https://doi.org/10.1007/s10661-008-0559-0

Furtula, V., Osachoff, H., Derksen, G., Juahir, H., Colodey, A., and Chambers, P. 2012. Inorganic

nitrogen, sterols and bacterial source tracking as tools to characterize water quality and possible

contamination sources in surface water. Water Research 46(4):1079−1092.

https://doi.org/10.1016/j.watres.2011.12.002

Glew, J.R. 1988. A portable extruding device for close interval sectioning of unconsolidated core

samples. Journal of Paleolimnology 1(3):235–239. https://doi.org/10.1007/BF00177769

Grainge, J. 1971. Study of the sewage oxidation ponds, City of Yellowknife, Northwest Territories.

Edmonton, Alberta: Department of the Environment, Public Health Engineering Division.

Griffiths, K., Michelutti, N., Sugar, M., Douglas, M.S.V., and Smol, J.P. 2017. Ice-cover is the principal

driver of ecological change in High Arctic lakes and ponds. PLoS ONE 12(3):e0172989.

https://doi.org/10.1371/journal.pone.0172989

Grimalt, J.O., Fernández, P., Bayona, J.M., Albaigés, J. 1990. Assessment of fecal sterols and ketones as

indicators of urban sewage inputs to coastal waters. Environmental Science and Technology

24(3):357–365. https://doi.org/10.1021/es00073a011

Grimm, E.C. 1987. CONISS: A FORTRAN 77 program for stratigraphically constrained cluster analysis

by the method of incremental sum of squares. Computers and Geosciences 13(1):13–35.

https://doi.org/10.1016/0098-3004(87)90022-7

Gunnarsdóttir, R., Jenssen, P.D., Jensen, P.E., Villumsen, A., and Kallenborn, R. 2013. A review of

wastewater handling in the Arctic with special reference to pharmaceuticals and personal care

133

products (PPCPs) and microbial pollution. Ecological Engineering 50:76–85.

https://doi.org/10.1016/j.ecoleng.2012.04.025

Hadley, K.R., Douglas, M.S.V., McGhee, R.H., Blais, J.M., and Smol, J.P. 2010. Ecological influences of

Thule Inuit whalers on high Arctic pond ecosystems: a comparative paleolimnological study from

Bathurst Island (Nunavut, Canada). Journal of Paleolimnology 44(1):85–93.

https://doi.org/10.1007/s10933-009-9388-6

Hall, R.I., and Smol, J.P. 1992. A weighted-averaging regression and calibration model for inferring total

phosphorus concentration from diatoms in British Columbia (Canada) lakes. Freshwater Biology

27:417–434. https://doi.org/10.1111/j.1365-2427.1992.tb00551.x

Hall, R.I., and Smol, J.P. 2010. Diatoms as indicators of lake eutrophication. In Smol, J.P., and Stoermer,

E.F., eds. The Diatoms: Applications for the Environmental and Earth Sciences 2nd edition.

Cambridge: Cambridge University Press. 122–151.

Health Canada. 2006. Guidelines for Canadian drinking water quality: guideline technical document –

arsenic. Ottawa, Ontario: Water Quality and Health Bureau, Healthy Environments and

Consumer Safety Branch.

Heaton, T.H.E. 1986. Isotopic studies of nitrogen pollution in the hydrosphere and atmosphere: a review.

Chemical Geology 59:87–102. https://doi.org/10.1016/0168-9622(86)90059-X

Heinke, G.W., and Smith, D.W. 1986. Niven site renovation study. Yellowknife, Northwest Territories:

Daniel W. Smith and Associates Ltd.

Hocking, D., Kuchar, P., Plambeck, J.A., and Smith RA. 1978. The impact of gold smelter emissions on

vegetation and soils of a sub-arctic forest-tundra transition ecosystem. Journal of the Air Pollution

Control Association 28(2):133–137. https://doi.org/10.1080/00022470.1978.10470580

Ilyashuk, B., Ilyashuk, E., and Dauvalter, V. 2003. Chironomid responses to long-term metal

contamination: a paleolimnological study in two bays of Lake Imandra, Kola Peninsula, northern

Russia. Journal of Paleolimnology 30(2):217–230. https://doi.org/10.1023/A:1025528605002

134

Indian and Northern Affairs Canada. 2007. Giant Mine remediation plan. Yellowknife, Northwest

Territories: Mackenzie Valley Land and Water Board.

Johnson, R.K., Eriksson, L., and Wiederholm, T. 1992. Ordination of profundal zoobenthos along a trace

metal pollution gradient in northern Sweden. Water, Air, and Soil Pollution 65(3-4):339–351.

https://doi.org/10.1007/BF00479897

Juggins, S. 2015. rioja: Analysis of Quaternary science data, R package version (0.9-9). (http://cran.r-

project.org/package=rioja).

Kelly, M.G., Bennion, H., Cox, E.J., Goldsmith, B., Jamieson, J., Juggins, S., Mann, D.G., and Telford,

R.J. 2005. Common freshwater diatoms of Britain and Ireland: an interactive key. Bristol, United

Kingdom: Environment Agency.

Kendall, C., Elliott, E.M., and Wankel, S.D. 2007. Tracing anthropogenic inputs of nitrogen to

ecosystems. In: Michener, R., and Lajtha, K., eds. Stable Isotopes in Ecology and Environmental

Science 2nd edition. Oxford: Blackwell Publishing. 375–449.

Korosi, J.B., Cheng, W., and Blais, J.M. 2015. Organic pollutants in sediment core archives. In: Blais,

J.M., Rosen, M.R., and Smol, J.P., eds. Developments in Paleoenvironmental Research volume

18: Environmental Contaminants. Dordrecht: Springer. 161–185.

Krammer, K., and Lange-Bertalot, H. 1986-1991. Bacillariophyceae. Stuttgart: Gustav Fisher Verkag.

Larocque, I., and Rolland, N. 2006. A visual guide to sub-fossil chironomids from Québec to Ellesmere

Island. Rapport de recherche R-900. Québec, Québec: Institut National de la Recherche

Scientifique.

Leeming, R., Ball, A., Ashbolt, N., and Nichols, P. 1996. Using faecal sterols from humans and animals

to distinguish faecal pollution in receiving waters. Water Research 30(12):2893−2900.

https://doi.org/10.1016/S0043-1354(96)00011-5

135

Leng, M.J., Lamb, A.L., Heaton, T.H.E., Marshall, J.D., Wolfe, B.B., Jones, M.D, Holmes, J.A., and

Arrowsmith, C. 2005. Isotopes in lake sediments. In: Leng, M.J., ed. Isotopes in

Palaeoenvironmental Research. Dordrecht: Springer. 148-176.

Little, J.L., Hall, R.I., Quinlan, R., and Smol, J.P. 2000. Past trophic status and hypolimnetic anoxia

during eutrophicaton and remediation of Gravenhurst Bay, Ontario: comparison of diatoms,

chironomids, and historical records. Canadian Journal of Fisheries and Aquatic Sciences

57(2):333–341. https://doi.org/10.1139/f99-235

Luoto, T.P., and Salonen, V-P. 2010. Fossil midge larvae (Diptera: Chironomidae) as quantitative

indicators of late-winter hypolimnetic oxygen in southern Finland: a calibration model, case

studies and potentialities. Boreal Environmental Research 15(1):1–18.

Michelutti, N., and Smol, J.P. 2016. Visible spectroscopy reliably tracks trends in paleo-production.

Journal of Paleolimnology 56(4):253–265. https://doi.org/10.1007/s10933-016-9921-3

Michelutti, N., Hermanson, M.H., Smol, J.P., Dillon, P.J., and Douglas, M.S.V. 2007. Delayed response

of diatom assemblages to sewage inputs in an Arctic lake. Aquatic Sciences 69(4):523–533.

https://doi.org/10.1007/s00027-007-0928-8

Miyamoto, H.K., and Heinke, G.W. 1979. Performance evaluation of an arctic sewage lagoon. Canadian

Journal of Civil Engineering 6(2):324–328. https://doi.org/10.1139/l79-033

Morin, S., Duong, T.T., Dabrin, A., Coynel, A., Herlory, A.C., Baudrimont, M., Delmas, F., et al. 2008.

Long-term survey of heavy-metal pollution, biofilm contamination and diatom community

structure in the Riou Mort watershed, South-West France. Environmental Pollution 151(3):532–

542. https://doi.org/10.1016/j.envpol.2007.04.023

Mudge, S.M., and Ball, A.S. 2006. Sewage. In: Morrsion, R.D., and Murphey, B.I., eds. Environmental

Forensics, Contaminant Specific Guide. Amsterdam: Elsevier. 36–53.

136

Nishimura, M. 1982. 5β-Isomers of stanols and stanones as potential markers of sedimentary organic

quality and depositional paleoenvironments. Geochimica Cosmochimica Acta 46(3):423–432.

https://doi.org/10.1016/0016-7037(82)90233-2

Ogrua, K., Machihara, T., and Takada, H. 1990. Diagenesis of biomarkers in Biwa Lake sediments over 1

million years. Organic Geochemistry 16(4-6):805–813. https://doi.org/10.1016/0146-

6380(90)90119-K

Patton, D., Reeves, A.D., 1999. Sterol concentrations and temporal variations on the north shore mudflats

of Firth of Tay, Scotland. Marine Pollution Bulletin 38(7)613–618.

https://doi.org/10.1016/S0025-326X(99)00068-5

Quinlan, R., and Smol, J.P. 2001. Setting minimum head capsule abundance and taxa deletion criteria in

chironomid-based inference models. Journal of Paleolimnology 26(3):327–342.

https://doi.org/10.1023/A%3A1017546821591

Reavie, E.D., and Kireta, A.R. 2015. Centric, araphid and eunotioid diatoms of the coastal Laurentian

Great Lakes. Bibliotheca Diatomologica 62:1–184.

Reavie, E.D., and Smol, J.P. 1997. Diatom-based model to infer past littoral habitat characteristics in the

St. Lawrence River. Journal of Great Lakes Research 23(3):339–348.

https://doi.org/10.1016/S0380-1330(97)70917-1

Scheffer, M. 1998. Ecology of shallow lakes. London: Chapman and Hall.

Schindler, D.W., and Smol, J.P. 2006. Cumulative effects of climate warming and other human activities

on freshwaters of Arctic and Subarctic North America. AMBIO 35(4):160–168.

https://doi.org/10.1579/0044-7447(2006)35[160:CEOCWA]2.0.CO;2

Schindler, D.W., Kalff, J.,Welch, H.E., Brunskill, G.J., Kling, H., and Kritsch, N. 1974. Eutrophication in

the High Arctic – Meretta Lake, Cornwallis Island (75° N Lat.). Journal of the Fisheries Research

Board of Canada 31(5):647–662. https://doi.org/10.1139/f74-096

137

Smol, J.P. 2010. The power of the past: using sediments to track the effects of multiple stressors on lake

ecosystems. Freshwater Biology 55(s1):43–59. https://doi.org/10.1111/j.1365-2427.2009.02373.x

Smol, J.P., and Douglas, M.S.V. 2007. From controversy to consensus: making the case for recent climate

change in the Arctic using lake sediments. Frontiers in Ecology and the Environment 5(9):466–

474. https://doi.org/10.1890/060162

Squires, M. 1982. Examination of Niven Lake sludge quality. 1B Work Report. Yellowknife, Northwest

Territories: Environmental Protection Service.

Stewart, E.M., Michelutti, N., Blais, J.M., Mallory, M.L., Douglas, M.S.V., and Smol, J.P. 2013.

Contrasting the effects of climatic, nutrient, and oxygen dynamics on subfossil chironomid

assemablges: a paleolimnological experiment from eutrophic High Arctic ponds. Journal of

Paleolimnology 49(2):205–219. https://doi.org/10.1007/s10933-012-9658-6

Stewart, E.M., McIver, R., Michelutti, N., Douglas, M.S.V., and Smol, J.P. 2014. Assessing the efficacy

of chironomid and diatom assemblages in tracking eutrophication in High Arctic sewage ponds.

Hydrobiologia 721(1): 251–268. https://doi.org/10.1007/s10750-013-1667-6

Thienpont, J.R. Korosi, J.B., Hargan, K.E., Williams, T., Eickmeyer, D.C., Kimpe, L.E., Palmer, M.J.,

Smol, J.P., and Blais, J.M. 2016. Multi-trophic level response to extreme metal contamination

from gold mining in a subarctic lake. Proceedings of the Royal Society B 283(1836):20161125.

https://doi.org/10.1098/rspb.2016.1125

Tse, T.J., Codling, G., Jones, P.D., Thoms, K., Liber, K., Giesy, J.P., Wheater, H., and Doig, L.E. 2014.

Reconstructing long-term trends in municipal sewage discharge into a small lake in northern

Manitoba, Canada. Chemosphere 103:299–305.

https://doi.org/10.1016/j.chemosphere.2013.12.019

Vane, C.H., Kim, A.W., McGowan, S., Leng, M.J., Heaton, T.H., Kendrick, C.P., Coombs, P., Yang, H.,

and Swann, G.E. 2010. Sedimentary records of sewage pollution using faecal markers in

138

contrasting periurban shallow lakes. Science of the Total Environment 409(2):345−56.

https://doi.org/10.1016/j.scitotenv.2010.09.033

Vermaire, J.C., Prairie, Y.T., and Gregory-Eaves, I. 2011. The influence of submerged macrophytes on

sedimentary diatom assemblages. Journal of Phycology 47(6):1230−1240.

https://doi.org/10.1111/j.1529-8817.2011.01069.x

Walker, I.R. 2001. Midges: Chironomidae and related Diptera. In: Smol, J.P., Birks, H.J.B., and Last,

W.M., eds. Tracking Environmental Change Using Lake Sediments Volume 4: Zoological

Indicators. Dordrecht: Kluwer Academic Publishers. 43–66.

Wedeleh Yellowknives Dene. 1997. Wedeleh Yellowknives Dene: a history. Dettah, Northwest

Territories: Yellowknives Dene First Nations Council.

Wetzel, R.G. 2001. Limnology: Lake and River Ecosystems, 3rd edition. San Diego: Academic Press.

Wolfe, A.P., Miller, G.H., Olsen, C.A., Forman, S.L., Doran, P.T., and Holmgren, S.U. 2004.

Geochronology of high latitude lake sediments. In: Pienitz, R., Douglas, M.S.V., and Smol, J.P.,

eds. Tracking Environmental Change Using Lake Sediments Volume 8: Long-term

Environmental Change in Arctic and Antarctic Lakes. Dordrecht: Springer. 19–52.

Wolfe, A.P., Vinebrooke, R.D., Michelutti, N., Rivard, B., Das, B. 2006. Experimental calibration of

lake-sediment spectral reflectance to chlorophyll a concentrations: methodology and

paleolimnological validation. Journal of Paleolimnology 36(1):91–100.

https://doi.org/10.1007/s10933-006-0006-6

Yamomoto, T. 1975. Preliminary analysis of enteric virus content of wastewaters in selected communities

in the Northwest Territories and Alberta. Research Report. Edmonton, Alberta: University of

Alberta.

139

Table 4.1 Sterol ratios used to compare with sewage contamination in the sediments of Niven Lake, Yellowknife (NT), with the range of values typically used to classify sediments that have received human fecal inputs. Indication of human Sterol ratio Literature references fecal contamination yes no Coprostanol/(coprostanol+ >0.7 <0.3 Grimalt et al. 1990, Vane et al. 2010 5α-cholestanol) Fattore et al. 1996, Patton & Reeves Coprostanol/cholesterol >0.5 <0.3 1999, Tse et al. 2014 Epicoprostanol/coprostanol <0.2 >0.8 Froehner et al. 2009

140

Figure 4.1 Map showing the location of Niven Lake. The lake is located in a residential neighbourhood of Yellowknife (NT). Niven Lake drains the lake immediately to the west (Frame Lake), and water from Niven flows into Back Bay, which is a part of Yellowknife Bay of Great Slave Lake. Giant Mine (1948- 2004) and Con Mine (1938-2003) likely influenced Niven Lake with the release of arsenic trioxide dust. The inset shows the location of Great Slave Lake within Canada.

141

Table 4.2 Lake water concentrations of nutrients, major ions, and trace metals for Niven Lake and Frame Lake, Yellowknife (NT), are shown in the leftmost portion of the table. Water samples were taken from Niven Lake on 23 July 2015 and on 12 July 2016. Frame Lake was sampled on 30 March 2014. Data modified from Ferguson Simek Clark (FSC) (1990a) show averages of four sampling sites from Niven Lake and two sampling sites from Frame Lake taken on 26 June 1990 (except for arsenic, taken later on 23 October 1990). Dashes indicate unavailable data. Niven Lake 2015 2016 Frame FSC 1990 Niven Frame (2014) TN-f (mg/L)* 1.78 1.66 4 TN-u (mg/L)* 2.00 1.64 4.02 + Ŧ Est. NH4 (mg/L) min - ~0 ~0 Ammonium-N (mg/L) 4.575 0.134 + Ŧ Est. NH4 (mg/L) max - 0.97 2.80 + Ŧ Est. NH4 (mg/L) mean - 0.32 1.70 TP-f (µg/L)* 93 49 - TP-u (µg/L)* 119 73 - Chlorophyll a (µg/L) 2.35 - - DOC (mg/L) 28.2 25.3 42.8

Alkalinity (mg CaCO3/L) 89.6 86 308 Hardness (mg/L) 176 159 pH 9.98 9.89 7.59 pH 7.52 8.40 Cond. (µS/cm) 569 561 1210 Calcium (mg/L) 35.6 36.9 120 Sodium (mg/L) 52.6 51.9 56.4 Magnesium (mg/L) 19.7 23.6 52.2 Potassium (mg/L) 3.4 4.8 18 Sulfate (mg/L) 35 40 174 Chloride (mg/L) 110 104 110 Dissolved arsenic (µg/L) 47.0 44.1 277 Total arsenic (µg/L) 41.7 44.2 343 Total arsenic (µg/L) 65 190 *where TN is total nitrogen, TP is total phosphorus, “-u” indicates unfiltered measures, “-f” indicated filtered measures, TKN is total Kjeldahl nitrogen, DOC is dissolved organic carbon, and Cond. is specific conductance. Ŧ + Estimated ammonium (NH4 ) concentration, calculated using TN-f, nitrate+nitrite concentrations, DOC, and the minimum (8), maximum (41), or mean (20) DOC:DON (dissolved organic nitrogen) for streams from Wetzel (2001).

142

Table 4.3 Dissolved oxygen (DO) concentration (mg/L) and percent saturation (%), specific conductance (µS/cm), and water temperature (℃) over water depth (maximum depth = 0.9 m at sampling location) in Niven Lake from 23 July 2015. Depth (m) DO (mg/L) DO sat Cond (µS/cm) Temp (℃) (%) 0 10.3 105 266 18 0.5 7.7 85 466 17 0.9 1.6 18 463 16

143

Figure 4.2 Relative abundances (%) of sterols and stanols, as well as stable nitrogen isotopes (δ15N in ‰) measured in the sediment core from Niven Lake (Yellowknife, NT). The three lines in white show different diagnostic ratios for determining human-sourced fecal contamination in lake sediments (reviewed by Furtula et al., 2012). The shaded grey area represents the approximate duration of sewage inputs to Niven (1948- 1981).

144

Figure 4.3 A. The sub-fossil diatom assemblage (%) in the sediment core from Niven Lake (Yellowknife, NT). Right-hand line graphs show the chrysophyte cyst to diatom valve ratio and sedimentary concentrations of chlorophyll a (mg/g) measured using visual reflectance spectroscopy (VRS-chla). The dotted line indicates the transition between CONISS groupings. B. The sub-fossil chironomid assemblage (%) and the number of head capsules per gram of dry sediment (HC/g dry sed). The shaded grey area on each stratigraphy represents the approximate duration of sewage inputs to Niven (1948-1981), and the grey box outline shows the overlapping operational periods of Con Mine (1938-2003) and Giant Mine (1948-2004).

145

Supplemental Figure 4.1 210Pb gamma spectrometry dating was used to develop a chronology for the Niven Lake (NT) sediment core. A. The radioactivity in becquerels per kilogram (Bq/kg) for total 210Pb (with error), 214Pb, 214Bi, and 137Cs, with an inset showing unsupported 210Pb and supported 210Pb (estimated using an average of 214Pb and 214Bi). B. The chronology (years AD) and associated error derived using the Constant Rate of Supply (CRS) model. C. The CRS-derived sedimentation rate (g/cm2/year). All x-axes are plotted as the midpoint of the 0.5-cm interval used for analysis.

146

Supplemental Figure 4.2 Mean air temperatures for the City of Yellowknife (NT) including A. annual means for select years between 1944-2015 and B. summer means of June, July, and August for select years between 1943-2015. Data were included for years that had all monthly mean air temperatures for annual averages (n = 47 years) or all June/July/August monthly means for summer averages (n = 58 years), and overall averages the time periods are shown as grey dashed lines. Data were sourced from Environment Canada.

147

Chapter 5

General Discussion

One of the main aims of this thesis was to determine the relative influences of nutrients, oxygen, and climate on chironomid species distributions in shallow freshwater ecosystems. Each chapter examined subfossil chironomid assemblages in sediments from shallow lakes that have undergone eutrophication, providing evidence of the indirect influence of nutrients on species distributions. The comparison of impacted (eutrophic) ponds and reference sites from the High Arctic (Chapter 2) to those from the Great Lakes Region (Chapter 3) highlighted the difference in how nutrients and oxygen interact to affect chironomid assemblages across latitude. Comparing data from Chapters 2 and 3 with the detailed study of Niven Lake (Chapter 4) demonstrated the importance of depth and macrophytes to chironomid response during eutrophication, emphasizing that there can be differential responses even among shallow systems due to these factors. Collectively, the studies in this thesis show general patterns about the overall functioning of shallow systems during eutrophication, as well as the often co-occurring stressor of pollutant contamination (e.g. metals).

Chironomids as indicators of eutrophication in shallow freshwaters The Resolute Bay sewage ponds described in Chapter 2 acted as a “natural laboratory” where it could be further demonstrated that chironomids, unlike diatoms, do not respond physiologically to phosphorus in the water column, making paleolimnological reconstructions of phosphorus using chironomid assemblages problematic. Chironomid-based paleolimnological reconstructions of nutrients or primary production were popular over a decade ago (Brodersen and Lindegaard 1999; Brodersen et al.

2001; Brooks et al. 2001; Zhang et al. 2006), and more recently they have been used as nutrient indicators in a lake management capacity (e.g. Ruse 2010). These studies prompted my investigation to explicitly test the chironomid response to eutrophication in the absence of nutrient-driven oxygen depletion. The comparison of sewage ponds and oligotrophic reference sites showed that chironomids do not experience

148

typical temperate eutrophication-related species shifts when oxygen is decoupled from nutrients in this

High Arctic environment (due to the 24-hours of daylight and effective wind-mixing of the entire shallow water columns), as has also been demonstrated for seabird-impacted ponds (Michelutti et al. 2011;

Stewart et al. 2013). This suggests that, in shallow sites where nutrients do not affect oxygen concentrations, chironomids should not be used to reconstruct phosphorus, as has also been previously suggested for several shallow temperate sites (Brodersen and Quinlan 2006; Langdon et al. 2006).

Complacent chironomid assemblages during the eutrophication of the Resolute sewage ponds was contrasted by the shifts in the diatom assemblages of the sewage ponds and the dissimilar diatom assemblages in the reference sites, emphasizing the fact that the diatoms in the ponds responded to eutrophication despite the extreme climate.

My studies suggest that the relative influence of nutrients on chironomid species assemblages appears to be indirect, since the increase in head capsule abundance was the sole indication of eutrophication via chironomids in the sediment records, which was attributable to greater food availability from increased algal production. Similarly, the physiological response of chironomids to phosphorus concentration in tropical streams has been shown to be: 1) mediated through detrital food quality; and 2) based around local adaptation to food quality rather than species differences, since the genetic similarity of chironomids from high-phosphorus and low-phosphorus streams was also demonstrated (Small et al.

2011). In fact, it has been directly shown that phosphorus additions resulted in faster larval chironomid growth rates and turnover (from larva to adult) due to enhanced detrital-based food resources in streams

(Ramírez and Pringle 2006). The importance of food quality and quantity to larval chironomids is widely recognized (Berg 1995; Brodersen and Quinlan 2006). Thus, the lack of taxonomic shifts in the Resolute

Bay sewage ponds and the increases in overall chironomid production reinforce previous studies on chironomid ecology.

Using a similar impact-reference comparison, Chapter 3 investigated similar relationships in temperate sites that were impacted by large colonies of waterbirds. However, here the reference sites were

149

also eutrophic in contrast to ultra-oligotrophic High Arctic reference sites of Resolute Bay. This changed the dynamics of the impact-reference pond comparison such that the response of chironomids was examined along a gradient of eutrophication (from ~100 to >2000 µg/L total phosphorus) in the temperate sites of Chapter 3. In the Lake Ontario island ponds, the difference in chironomid assemblage between bird-impacted and non-impacted ponds appeared to be predominantly due to habitat (e.g. depth of pond, presence of macrophytes) rather than the large gradient in phosphorus. In fact, the non-impacted ponds on the low end of the phosphorus gradient (~100 µg/L total phosphorus) appeared to promote greater chironomid abundance compared to the hypereutrophic (>2000 µg/L total phosphorus) bird-impacted ponds, with up to four times the number of head capsules per gram dry sediment (Figures 3.6, 3.7). The greater abundance of chironomid remains in the non-impacted ponds may suggest that there is a limit (i.e. threshold) to the growth-enhancing role of phosphorus (via increased food availability) for larval chironomids, though other possible factors may have also led to lower chironomid abundance in the bird- impacted ponds (e.g. pond desiccation, metal pollution). Interestingly, as was the case with the Resolute sewage ponds, the sedimentary diatom assemblages showed a greater distinction in species composition between bird-impacted and non-impacted sites (Stewart et al. 2015; Appendix B), closely reflecting the higher phosphorus concentrations in the impacted sites. This difference in ecological response to eutrophication between diatoms and chironomids has also be shown by using both proxies to quantitatively reconstruct total phosphorus in the same core, demonstrating that diatom species have a stronger (and more direct) relationship with total phosphorus than chironomids, due to the presence of large secondary gradients of other environmental variables that influence the chironomid assemblage

(Lotter et al. 1998).

The temperate environment of the Chapter 3 ponds does not foster similar dynamics of oxygen and nutrients as was recorded in the High Arctic sites, since night-time lows of oxygen in the temperate sites resulted from periods where decomposition consumes oxygen without concurrent oxygen production from photosynthesis. The difference in oxygen dynamics between the High Arctic and temperate sites

150

may be reflected as differences in the relative abundance of Chironomus plumosus-type, which is the typical eutrophication indicator in deep profundal habitats due to its tolerance of low oxygen levels (i.e. a

“blood worm”), but is still also often found in shallow eutrophic sites as well (Brodin 1982; Salonen et al.

1993; Hall et al. 1999; Stewart et al. 2013). Furthermore, C. plumosus-type is unrestricted by latitude

(Walker et al. 1997), making it useful for comparing temperate and High Arctic sites and avoiding the confounding effect of the distinct latitudinal distribution of many chironomid species. Consistent low relative abundances of C. plumosus-type were found across all sites in Chapters 2 and 3 regardless of trophic status, including eutrophic impacted ponds, as well as oligotrophic and eutrophic reference ponds, highlighting the fact that in shallow sites, C. plumosus-type does not act as an indicator of eutrophication that accurately reflects the intensity of the event. Interestingly, however, the relative abundance of C. plumosus-type in the temperate sites was almost double that of the Arctic sites (~10% vs 5%, respectively). This could possibly be because the temperate eutrophic ponds (both impacted and reference) had periods of hypoxia or anoxia at night, even though the majority of the time the ponds were oxic.

The findings of Chapters 2 and 3, along with those of previously published studies (e.g. Clerk et al. 2000; Little et al. 2000; Langdon et al. 2006; Michelutti et al. 2011; Stewart et al. 2013), suggest that referring to chironomid species as “eutrophic” or “oligotrophic” is a misnomer, since trophic status is generally delineated based on total phosphorus concentration, which does not directly influence chironomid species distributions (as in the case of C. plumosus-type). However, chironomids are affected by changes in primary production (which is directly related to total phosphorus), showing an increase in overall abundance due to increased food availability, as discussed above (Brodersen and Quinlan 2006;

Michelutti et al. 2011; Stewart et al. 2013; Chapters 2 and 3). Furthermore, chironomid taxon distributions appear to be related to habitat (particularly the presence or type of macrophytes) (Brodersen et al. 2001), which can change with eutrophication, as is the case in the alternate stable states of shallow sites (Scheffer et al. 1993). The differences in chironomid assemblage between the bird-impacted and

151

non-impacted ponds of Chapter 3 may further demonstrate this relationship. The extremely shallow depth

(<50 cm) of the impacted ponds, together with the hypereutrophic concentrations of total phosphorus, may have precluded the occurrence of abundant submerged macrophytes (i.e. the turbid state, Scheffer et al. 1993). The deeper (>1 m) and less eutrophic reference sites did support macrophyte growth, leading to a more diverse littoral assemblage in the reference sites compared with the bird-impacted ponds. For example, the reference sites included several littoral taxa (e.g. Dicrotendipes, Paracricotopus,

Polypedilum) that were not recorded in the bird-impacted sites, which instead had greater abundances of semi-terrestrial species (tolerant of ephemeral conditions), such as Smittia.

Both the findings of Chapters 2 and 3 support the growing body of literature investigating chironomid ecology in shallow eutrophic sites, which suggests that chironomid response to increasing nutrients is generally not based on the increasing prevalence of hypoxia or anoxia, as it is in deep stratifying lakes (reviewed by Brodersen and Quinlan 2006). However, the latitudinal differences in the decoupling of the relationship between nutrients and oxygen, had yet to be explicitly shown. Thus, contrasting Chapters 2 and 3 highlights that oxygen is independent of nutrients in the High Arctic due to

24 hours of daylight in the summer, but daylight patterns in temperate regions result in diurnal fluctuations in oxygen concentration in eutrophic sites. Though each study provides a different perspective on chironomid ecology in shallow sites, they each add insight into the conditions that may contribute to species distributions under very diverse scenarios of eutrophication.

Chapter 4, the investigation of sewage-related eutrophication of Niven Lake in Yellowknife (NT), presented a slightly different perspective on eutrophication compared to Chapters 2 and 3, as an impact- reference methodology was not used. Overall, the sediment record from Niven Lake once again demonstrates a stronger response to sewage inputs in the diatom assemblages than the chironomids, as might be expected for a shallow system where oxygen concentrations are not as closely correlated to nutrients, such as was the case in the ponds of Chapters 2 and 3. Even though the Niven Lake study did not include a reference site for comparison, nearby Tathlina Lake (NT) could act as a reference site for

152

Niven Lake since it is approximately the same depth as Niven Lake (~1 – 2 m), is oligo-to-mesotrophic, and is located close to Yellowknife, approximately 60 km southwest of Great Slave Lake (Stewart et al.

2016). However, the major difference between Niven Lake and Tathlina Lake is the surface area, which is

573 km2 for Tathlina compared to Niven’s 0.07 km2, as well as the greater macrophyte growth in Niven

Lake, likely owing to its eutrophic status. Despite these differences, some useful inferences can still be made based on the similar depth and climate. The most notable difference in the subfossil assemblages of

Niven and Tathlina lakes are noted among the dominant diatom taxa of each, where Eolimna submuralis and small benthic Fragilaria dominate Tathlina Lake with little change throughout the record (Stewart et al. 2016), compared to the large turnover to eutrophic species such as Nitzschia amphibia, Eolimna minima, Sellaphora seminulum, and Stephanodiscus hantzschii (Kelly et al. 2005) in Niven Lake following sewage inputs (Figure 4.3). In contrast, the chironomid assemblages of Niven Lake and

Tathlina Lake appear largely similar with low relative abundances of many taxa including Chironomus plumosus-type, Cladopelma, Polypedilum, Psectrocladius, Cricotopus/Orthocladius, Procladius and

Tanytarsini species. The large difference in diatom assemblage between sewage-impacted Niven Lake and oligo-mesotrophic Tathlina Lake, along with the similarity of their chironomid assemblages, may constitute further evidence that chironomids do not always dependably track nutrients such as phosphorus.

The chironomid assemblage in Niven Lake may have resisted change during the onset and early years of sewage inputs, in part due to the resiliency of shallow systems, but also perhaps due to a lack of hypoxia that generally accompanies eutrophication in deeper systems, as was the case in the Resolute Bay sewage ponds. However, the increase in the relative abundance of Chironomus plumosus-type after the end of sewage inputs indicates that Niven Lake was indeed experiencing hypoxia or anoxia (Figure 4.3), a finding supported by the dissolved oxygen measures made in July 2015 showing hypoxia near the lake bottom. The evidence of low oxygen in Niven Lake suggested that the effects of anoxia on chironomid assemblages cannot be ruled out entirely just because the system is shallow, as may have also been the

153

case for the bird-impacted sites of Chapter 3, where diurnal cycles of hypoxia may have promoted higher relative abundances of C. plumosus-type compared to the oxic sewage ponds of Resolute Bay. It is thought that oxygen depletion is relatively unimportant to chironomids in shallow eutrophic sites because wind mixing keeps oxygen levels high, except for occasional hypoxia that might, for example, occur under ice (Brodersen and Quinlan 2006). However, the research in this thesis cumulatively demonstrated that oxygen depletion may be important to chironomid assemblages in shallow eutrophic systems under certain conditions (such as night time lows or reduced mixing due to excess macrophyte growth). Overall, this speaks to the complexity of shallow systems and the inherent risk in over-simplifying the ecological functioning of shallow lakes during eutrophication, especially when making inferences from the sediment record.

General patterns in shallow lake and pond ecology

Comparing all of the eutrophic sites from each chapter demonstrates some general patterns about the functioning of shallow systems during eutrophication. For example, Niven Lake is approximately 2 m deep, making it 4 times the average depth of the other study sites from Chapters 2 and 3. While all study sites were shallow, the greater depth of Niven Lake likely produced a more complex habitat compared to the sites of Resolute or Lake Ontario, especially in the case of abundant submerged macrophytes. The investigation of Niven lake seemed to highlight the resiliency of shallow systems to eutrophication, likely due to the presence of dense macrophyte growth, as was proposed for the concept of alternate stable states in shallow lakes (Scheffer et al. 1993). This is evident in the muted and delayed response to eutrophication in all proxies examined in the Niven Lake sediment record, which is not only common in

Arctic systems (e.g. Douglas and Smol 2000; Michelutti et al. 2007), but also could be due to the direct competition that macrophytes pose to planktonic blooms for nutrients, as well as indirectly by providing habitat for grazers and stabilizing sediments. However, in contrast to the extremely shallow ponds in

Chapters 2 and 3, this resiliency likely does not hold when water depth is too shallow to support abundant aquatic macrophyte growth (as was evident in field observations). The lack of resiliency was possibly

154

reflected in the sediment records of the Resolute and Lake Ontario sites such that they may have immediately responded to eutrophication with increased primary production and shifts in diatom assemblages according to monitoring records and our 210Pb dates (when applicable) (Figure 2.7, Stewart et al. 2015; Appendix B). Ultimately, this emphasizes that the shallow ponds (that are typically unique to

Arctic or alpine regions - where desiccation is less likely) do not follow the same principles as shallow macrophyte-rich lakes during eutrophication, as was the case for the High Arctic ponds and the ephemeral temperate ponds on bird nesting islands in Lake Ontario.

Effects of metal pollution on chironomid assemblages

Contamination by metals or other trace elements often accompanies eutrophication, such as, for example, when industrial (e.g. mining) effluent reaches freshwater systems (Gagneten et al. 2007; Church et al. 2006). The aerial deposition of arsenic trioxide from nearby gold mining operations to Niven Lake

(Chapter 4) seemed to also influence the biology of the lake in addition to eutrophication by sewage disposal. A more recently discovered source of both nutrients and contaminants to freshwaters is biovector transport, or animals that move substances across ecosystem boundaries (e.g. from marine to freshwaters), including species of seabirds or salmon (Blais et al. 2005; Blais et al. 2007). The most prominent example of biovector transport in this thesis are the large waterbird colonies from Chapter 3, though a less obvious example could be human biovector transport from sewage effluent to the ponds near Resolute Bay (Chapter 2), since biomagnification of heavy metals (e.g. mercury) in marine food webs and subsequent consumption by people in northern communities has been an area of interest in recent decades (Tian et al. 2011). Though metal or trace element contaminants were not examined in the

Resolute Bay sewage ponds, this could be done since not only can sediment records be examined for trace element concentrations, but the span of 30-year monitoring records that include water chemistry are some of the longest from the Canadian High Arctic.

The possible response of chironomids to arsenic contamination in Niven Lake may be difficult to separate from the effects of nutrient and oxygen dynamics; however, comparison with another nearby

155

polluted lake in Yellowknife gives some insight into how chironomids may respond depending on habitat.

For example, the response of chironomids in Pocket Lake, situated on the Giant Mine Property (a formerly long-running gold mine in Yellowknife), was drastic with a switch from Tanytarsus to

Cricotopus and Psectrocladius at the height of arsenic pollution (Thienpont et al. 2016). In contrast, similar shifts involving the same genera appear to occur in the Niven Lake sediment record, however the responses are muted and subtle. This is likely in part due to the severity of contamination experienced by each lake, since the recent water concentrations of arsenic are much lower in Niven Lake than Pocket

Lake; however, the abundant macrophyte growth in Niven Lake might also provide resistance to pollution because of the ability of many macrophytes to take up contaminants (Lizama et al. 2011). In this way, the abundant macrophyte growth that resulted from the early stages of sewage inputs to Niven (Heinke and

Smith 1986) may have aided in the resistance of the lake’s biota to the negative effects of arsenic contamination. Furthermore, Niven Lake is shallower than Pocket Lake and has no pelagic habitat.

Thienpont et al. (2016) concluded that planktonic species were the most susceptible to contamination in

Pocket Lake, and thus the benthic and littoral species supported in Niven Lake might also be less susceptible to the negative effects of arsenic contamination.

Comparing the metal contamination found in the Lake Ontario bird-impacted ponds to that of

Niven Lake further highlights the differences between shallow lakes and ponds, as well as possibly points to chironomid species that may be particularly tolerant of metal pollution. Though different contaminants were identified in the bird-impacted ponds versus Niven Lake, one might note that the contamination in the shallow ponds possibly led to the near-absence of chironomids in the most contaminated bird- impacted site, Gull Pond, as well as in High Bluff (Chapter 3). While many factors likely contributed to this absence (e.g. potential drying out of the pond), the lack of macrophytes could also contribute to the susceptibility of these ponds (and therefore of their chironomid assemblages) to metal contamination, as there would be no macrophyte-based buffering against pollution. Both the bird-impacted sites, as well as

Niven Lake, had assemblages with the taxa Psectrocladius and Chironomus, and while Chironomus is

156

known for its tolerance to pollution, especially of heavy metals (Warwick et al 1987; Hart et al. 1986;

Johnson et al. 1992; Diggins and Stewart 1993; Ilyashuk et al. 2003), Psectrocladius is less often reported as such. However, in other studies of metal pollution, increases in Psectrocladius concurrent with contamination have occurred (Brooks et al. 2005; Thienpont et al. 2016), suggesting that Psectrocladius is robust to metal pollution, which may have to do with the fact that it is commonly found associated with vegetation in benthic or littoral areas (Lindegaard 1992; Brodersen et al. 2001).

Summary

Overall, this thesis has demonstrated the response of chironomid assemblages to eutrophication in shallow systems under diverse conditions across a broad latitudinal gradient. As shallow systems often prove difficult for paleolimnological study due to possible disturbance of the sediment record, they are rarely used for the analysis of long-term environmental change (Smol 2008). However, small shallow lakes are often common on the landscape, and thus represent a relatively untapped source of information.

This thesis examined freshwater ecosystems on the extreme end of the shallow lake spectrum, with ponds that were often less than 50 cm deep. Though the sediment records from these sites posed some issues, it was demonstrated here that they can still provide useful information about ecological relationships, as in the case of chironomids with nutrients and oxygen. As such, very shallow ponds are relatively understudied in most places (except perhaps in the Arctic, where such systems are common), and so this thesis contributes to the growing body of knowledge concerning the long-term development of such shallow systems, especially in the context of environmental issues, such as eutrophication or pollution by other contaminants.

References

Berg MB (1995) Larval food and feeding behaviour. In: Armitage PD, Pinder LC, Cranston P (eds) The

Chironomidae: Biology and Ecology of Non-biting Midges. Chapman & Hall, London

Blais JM, Kimpe LE, McMahon D, Keatley BE, Mallory ML, Douglas MSV, Smol JP (2005) Arctic

seabirds transport marine-derived contaminants. Science 309:445 157

Blais JM, Macdonald RW, Mackey D, Webster E, Harvey C, Smol JP (2007) Biologically mediated

transport of contaminants to aquatic systems. Environ Sci Technol 41:1075–1084

Brodersen KP, Lindegaard C (1999) Classification, assessment and trophic reconstruction of Danish lakes

using chironomids. Freshw Biol 42:143–157

Brodersen KP, Odgaard BV, Vestergaard O, Andersen NJ (2001) Chironomid stratigraphy in the shallow

and eutrophic Lake Søbygaard, Denmark: chironomid–macrophyte co-occurrence. Freshw Biol

46:253–267

Brodersen KP, Quinlan R (2006) Midges as paleoindicators of lake productivity, eutrophication and

hypolimnetic oxygen. Quat Sci Rev 25:1995–2012

Brodin Y (1982) Palaeoecological studies of the recent development of Lake Växjösjön. IV.

Interpretation of the eutrophication process through the analysis of subfossil chironomids. Arch

Hydrobiol 93:313–326

Brooks SJ, Bennion H, Birks HJB (2001) Tracing lake trophic history with a chironomid-total phosphorus

inference model. Freshw Biol 46:513–533

Brooks SJ, Udachin V, Williamson BJ (2005) Impact of copper smelting on lakes in the southern Ural

Mountains, Russia, inferred from chironomids. J Paleolimnol 33:229–241

Church TM, Sommerfield CK, Velinsky DJ, Point D, Benoit C, Amouroux D, Plaa D, Donard OFX

(2006) Marsh sediments as records of sedimentation, eutrophication and metal pollution in the

urban Delaware Estuary. Mar Chem 102:72–95

Clerk S, Hall R, Quinlan R, Smol JP (2000) Quantitative inferences of past hypolimnetic anoxia and

nutrient levels from a Canadian Precambrian Shield lake. J Paleolimnol 23:319–336

Diggins TP, Stewart KM (1993) Deformities of aquatic larval midges (Chironomidae: Diptera) in the

sediments of the Buffalo River, New York. J Gt Lakes Res 19:648–659

Douglas MSV, Smol JP (2000) Eutrophication and recovery in the High Arctic: Meretta Lake (Cornwallis

Island, Nunavut, Canada) revisited. Hydrobiologia 431:193–204

158

Gagneten AM, Gervasio S, Paggi JC (2007) Heavy metal pollution and eutrophication in the lower Salado

River Basin (Argentina). Water Air Soil Pollut 178:335–349

Hall RI, Leavitt PR, Quinlan R, Dixit AS, Smol JP (1999) Effects of agriculture, urbanization, and

climate on water quality in the northern Great Plains. Limnol Oceanogr 44:739–756.

Hart DR, McKee PM, Burt AJ (1986) Benthic community and sediment quality assessment of Port Hope

Harbour, Lake Ontario. J Gt Lakes Res 12:206–220

Heinke GW, Smith DW (1986) Niven site renovation study. Daniel W. Smith and Associates Ltd.,

Yellowknife

Ilyashuk B, Ilyashuk E, Dauvalter V (2003) Chironomid responses to long-term metal contamination: a

paleolimnological study in two bays of Lake Imandra, Kola Peninsula, northern Russia. J

Paleolimnol 30:217–230

Johnson RK, Eriksson L, Wiederholm T (1992) Ordination of profundal zoobenthos along a trace metal

pollution gradient in northern Sweden. Water Air Soil Pollut 65:339–351

Kelly MG, Bennion H, Cox EJ, Goldsmith B, Jamieson J, Juggins S, Mann DG, Telford RJ (2005)

Common freshwater diatoms of Britain and Ireland: an interactive key. Environment Agency,

Bristol

Langdon PG, Ruiz Z, Brodersen KP, Foster IDL (2006) Assessing lake eutrophication using chironomids:

understanding the nature of community response in different lake types. Freshw Biol 51:562–577

Lindegaard C (1992) Zoobenthos ecology of Thingvallavatn: vertical distribution, abundance, population

dynamics, and production. Oikos 64:257–304

Little JL, Hall RI, Quinlan R, Smol JP (2000) Past trophic status and hypolimnetic anoxia during

eutrophicaton and remediation of Gravenhurst Bay, Ontario: comparison of diatoms,

chironomids, and historical records. Can J Fish Aquat Sci 57:333–341

Lizama AK, Fletcher TD, Sun G (2011) Removal processes for arsenic in constructed wetlands.

Chemosphere 84:1032–1043

159

Lotter AF, Birks HJB, Hofmann W, Marchetto A (1998) Modern diatom, cladocera, chironomid, and

chrysophyte cyst assemblages as quantitative indicators for the reconstruction of past

environmental conditions in the Alps. II. Nutrients. J Paleolimnol 19:443–463

Michelutti N, Hermanson MH, Smol JP, Dillon PJ, Douglas MSV (2007) Delayed response of diatom

assemblages to sewage inputs in an Arctic lake. Aquat Sci 69:523–533

Michelutti N, Mallory ML, Blais JM, Douglas MSV, Smol JP (2011) Chironomid assemblages from

seabird-affected high Arctic ponds. Polar Biol 34:799–812

Ramirez A, Pringle CM (2006) Fast growth and turnover of chironomid assemblages in response to

stream phosphorus levels in a tropical lowland landscape. Limnol Oceanogr 51:189–196

Ruse L (2010) Classification of nutrient impact on lakes using the chironomid pupal exuvial technique.

Ecol Indic 10:594–601

Salonen V-P, Alhonen P, Itkonen A, Olander H (1993) The trophic history of Enäjärvi, SW Finland, with

special reference to its restoration problems. Hydrobiologia 268:147–162

Scheffer M, Hosper SH, Meijer M-L, Moss B, Jeppesen E (1993) Alternative equilibria in shallow lakes.

Trends Ecol Evolut 8:275–278

Small GE, Wares JP, Pringle CM (2011) Differences in phosphorus demand among detritivorous

chironomid larvae reflect intraspecific adaptations to differences in food resource stoichiometry

across lowland tropical streams. Limnol Oceanogr 56:268–278

Smol JP (2008) Pollution of Lakes and Rivers: A Paleoenvironmental Perspective. Blackwell Publishing,

Malden

Stewart EM, Michelutti N, Blais JM, Mallory ML, Douglas MSV, Smol JP (2013) Contrasting the effects

of climate, nutrient, and oxygen dynamics on subfossil chironomid assemblages: a

paleolimnological experiment from eutrophic High Arctic ponds. J Paleolimnol 49:205–219

160

Stewart EM, Michelutti N, Shenstone-Harris S, Grooms C, Weseloh C, Kimpe LE, Blais JM, Smol JP

(2015) Tracking the history and ecological changes of rising Double-crested Cormorant

populations using pond sediments from islands in eastern Lake Ontario. PLoS ONE 10:e0134167

Stewart EM, Coleman KA, Korosi JB, Thienpont JB, Palmer MJ, Blais JM, Smol JP (2016) Assessing

environmental stressors on a commercial walleye fishery from a large northern ecosystem

(Tathlina Lake) using water chemistry and paleolimnology. J Gt Lakes Res 42:217–222

Thienpont JR, Korosi JB, Hargan KE, Williams T, Eickmeyer DC, Kimpe LE, Palmer MJ, Smol JP, Blais

JM (2016) Multi-trophic level response to extreme metal contamination from gold mining in a

subarctic lake. Proc R Soc B 283:20161125

Tian W, Egeland GM, Sobol I, Chan HM (2011) Mercury hair concentrations and dietary exposure

among Inuit preschool children in Nunavut, Canada. Environ Int 37:42–48

Walker IR, Levesque AJ, Cwynar LC, Lotter AF (1997) An expanded surface-water palaeotemperature

inference model for use with fossil midges from eastern Canada. J Paleolimnol 18:165–178

Warwick WF, Fitchko J, McKee PM, Hart DR, Burt AJ (1987) The incidence of deformities in

Chironomus spp. from Port Hope Harbour, Lake Ontario. J Gt Lakes Res 13:88–92

Zhang E, Bedford A, Jones R, Shen J, Wang S, Tang H (2006) A subfossil chironomid-total phosphorus

inference model for lakes in the middle and lower reaches of the Yangtze River. Chinese Sci Bull

51:2125–2132

161

Appendix A Lake Ontario bird ponds radiometric dating profiles

The following radiometric dating data pertain to four study sites from Chapter 3 and Appendix B (High

Bluff, Gull, Little Galloo, and Calf). The radioactivity and total lead data were included as supplementary material in:

Hargan KE, Stewart EM, Michelutti N, Grooms C, Kimpe L, Mallory M, Smol JP, Blais JM (in review)

Incorporating sterols and stanols as biomarkers for tracking waterbird impacts to temperate

ponds. Proc R Soc Lond B Biol Sci: manuscript ID RSPB-2017-2669

References in figure captions:

Eisenreich SJ, Capel PD, Robbins JA, Bourbonniere R (1989) Accumulation and diagenesis of

chlorinated hydrocarbons in lacustrine sediments. Environ Sci Technol 23:1116-1126

Graney JR, Halliday AN, Keeler GJ, Nriagu JO, Robbins JA, Norton SA (1995) Isotopic record of lead

pollution in lake sediments from the northeastern United States. Geochim Cosmochim Acta

59:1715-1728

162

Figure A1. Sediment dating profiles for cores from A. High Bluff, B. Calf, C. Gull and D. Little Galloo Ponds. The upper panel for each site depicts the radioactivites (Bq/kg) for unsupported 210Pb, 214Pb, and 137Cs, as well as Constant Rate of Supply (CRS) chronologies for High Bluff and Calf. The bottom panel of each site shows stable lead concentrations (ppm) of the study core compared with data for Lake Ontario sediments (Graney et al. 1995). Due to probable mixing in Gull and Little Galloo, the intrinsic time resolution (Eisenreich et al. 1989) was used to estimate the approximate number of years contained in the plateau of unsupported 210Pb.

163

Figure A2. The stable lead isotope ratios (206Pb/207Pb) shown with open circles and plotted by CRS-derived date for A. High Bluff, B. Calf, C. Gull, and D. Little Galloo islands in eastern Lake Ontario. Also shown are data from the Great Lakes (two cores taken in different years from Lake Erie and one core from each Lake Michigan and Lake Ontario) modified from Graney et al. (1995). 164

Figure A3. Cumulative excess (unsupported) 210Pb inventories (Bq/m2) over core depth (cm) for all sites, including bird-impacted sites (High Bluff, Gull, Little Galloo) and the non-impacted site (Calf), as well as the sediment accumulation rate calculated using the Constant Rate of Supply model.

165

Appendix B Diatom data for Lake Ontario island ponds

Diatom subfossil assemblage data, water chemistry data, and bird species data for High Bluff, Gull, Little

Galloo and Calf Ponds from Chapter 3 are presented below.

Table B1. Pond characteristics for each site include the estimated number of nests on each island in 2015 (where DCCO = Double-crested Cormorants, RBGU = Ring-billed Gulls, HERG = Herring Gulls, and CATE = Caspian Terns), as well as the pond dimensions (m) taken from Google Earth and approximate maximum depth (m) estimated at the time of core sampling. Water samples were taken from High Bluff and Gull on April 24th 2014, from Little Galloo on May 23rd 2014, and from Calf on June 19th 2014. Pond traits & BIRD IMPACTED REFERENCE water chemistry High Bluff Gull Little Galloo Calf DCCO nests ~4000 0 2264 0 RBGU nests 0-few ~30,000 ~50,000 - HERG nests 0-few few 784 - CATE nests - - 2084 - Dimensions (m) 35×45 125×30 28×120 406×65 Depth (m) 0.15 0.3 0.4 1 TP-u (µg/L) 4310 8130 6410 151 TN-f (mg/L) 109 30.2 3.47 1.04 Chl-a (µg/L) 6.9 337 243 3.5 DOC (mg/L) 24.4 29.1 31.1 12.4 pH 8.6 8.8 9.5 8.5* Cond. (µS/cm) 1350 547 193.4 172.5* [Cl-]-u (mg/L) 23.3 21.6 6.7 5.45 2- [SO4 ]-u (mg/L) 125 58.6 16.2 1.2 [Ca2+]-f (mg/L) 250 72.2 32.2 34.8 [K+]-f (mg/L) 42.8 54.1 14.3 0.44 [Mg2+]-f (mg/L) 14 9.91 2.72 3.55 [Na+]-f (mg/L) 22.9 19.2 7.87 4.23 Above “-u” indicates unfiltered measures and “-f” indicates filtered measures. Parameter abbreviations are TP = total phosphorus, TN = total nitrogen, Chl-a = chlorophyll-a, DOC = dissolved organic carbon, and Cond = specific conductance. *Averaged values from field measures taken on 19/06/14 and 21/07/16.

166

Figure B1. Subfossil diatom assemblages (% relative abundance), visual reflectance spectroscopy chlorophyll-a concentrations (VRS-chla in mg/g dry weight), and stable nitrogen isotopes (δ15N in ‰) of sediment cores from A) cormorant-impacted High Bluff Pond and B) predominantly gull-impacted Gull Pond. Radiometric dating information is given in Hargan et al. (in review). The horizontal dotted grey lines and corresponding zone names indicate important CONISS groups in the diatom assemblages. Significant breakpoints (BP) for VRS-chla and δ15N are shown with grey dotted-dashed lines. The arrival of Double-crested Cormorants to High Bluff Island in 1986 (OMNR 2011, see Table B4 for reference) is also indicated.

167

Figure B2. Subfossil diatom assemblages (% relative abundance), visual reflectance spectroscopy chlorophyll-a concentrations (VRS-chla in mg/g dry weight), and stable nitrogen isotopes (δ15N in ‰) of sediment cores from A) mixed-impacted Little Galloo Pond and B) reference Calf Pond. Radiometric dating information is given in Hargan et al. (in review). The horizontal dotted grey lines and corresponding zone names indicate important CONISS groups in the diatom assemblages. Significant breakpoints (BP) for VRS-chla and δ15N are shown with grey dotted-dashed lines. The brief occupation of Double-crested Cormorants (120 nests) on Calf Island in 2008 (data from NYSDEC) is also indicated.

168

Table B2. Double-crested Cormorant nest counts for High Bluff Island and Gull Island in Presqu’ile Provincial Park modified from OMNR (2011). From 2003-2006 active culling took place under the Resource Management Implementation Plan (OMNR 2011*). Year High Bluff Gull 1984 0 0 1985 0 116 1986 32 132 1987 42 447 1988 168 471 1989 439 498 1990 888 704 1991 746 1050 1992 822 1596 1993 898 2143 1994 1724 1435 1995 1956 771 1996 2442 1257 1997 3196 819 1998 5126 1042 1999 6741 1074 2000 8105 867 2001 9532 789 2002 10384 1689 2003 7567 1113 2004 5339 1601 2005 3281 1328 2006 2615 204 2007 3795 60 2008 4152 0 2009 3872 0

*OMNR (2011) Presqu’ile Islands Resource Management Implementation Plan. Queen’s Printer for Ontario

169

Table B3. Numbers of bird nests (unless otherwise specified) on Little Galloo Island near Henderson Harbor (New York, USA). Census data are modified from Blokpoel and Weseloh (1982)ŧ for 1935-1981 and from the NYSDEC for 1986-2015. Species are: Double-crested Cormorant (DCCO), Black-crowned Night Heron (BCNH), Caspian Tern (CATE), Herring Gull (HERG), Great Black-backed Gull (GBBG), Ring-billed Gull (RBGU), and Common Tern (COTE). Dashes indicate years in which the species was not censused, but was known to be present on the island. Year DCCO BCNH CATE HERG GBBG RBGU COTE 1938 nesting 1945 200 birds 1800 1950 19,200 1951 11 - 1955 50 pairs 45,000 1961 50 63,000 1963 80 75,000 1965 75 100 85,000 1966 Nesting 100 pairs 100-120,000 pairs 1967 >75 pairs 85,000 1974 22 25 - 1975 28 29 - 1976 68 121 200 30,000 1977 130 130 27,308 1978 192 77 93 77,000 1980 264 - 1981 461 96 350 4 73,780 1986 1468 80 - 1987 2233 85-100 - 1988 2556 - - 1989 3910 26 - 1990 4121 23 - 1991 5428 12 - 1992 5433 2 - 1993 4743 0 - 1994 3745 0 - 1995 7585 0 - 1996 8410 2 - 1997 7591 1 - 1998 5839 0 - 1999 4570 1 1445 275 8 53,820 2000 5119 1 1350 - 2001 4875 1 1590 19 - 2002 4780 1 1585 15 - 2003 4251 3 1658 313 12 60,000 2004 3967 3 1560 - - - 2005 3179 4 1788 - - - 2006 2692 0 1589 367 4 -

170

(Table B3 continued) Year DCCO BCNH CATE HERG GBBG RBGU COTE 2007 2702 0 1580 - 1 - 2008 2492 0 1376 375 1 37,465 2009 1899 0 1499 356 0 - 2010 1758 0 1472 364 0 - 2011 2831 0 1934 459 0 - 2012 1729 0 2332 512 0 43,324 2013 2387 0 1848 645 0 - 20 2014 2283 0 2436 979 0 - 34 2015 2264 0 2084 784 0 - 30

ŧ Blokpoel H, Weseloh DV (1982) Status of Colonial Waterbirds Nesting on Little Galloo Island, Lake Ontario. Kingbird 32: 149-158

171

Appendix C Chapter 2 (Resolute Bay) raw counts

Table C1. Diatom relative abundance (%) data for R-12 core.

striata

cf cf

R-12

Depth

(cm) Amphora copulata striata Cyclotella Cyclotella (pinched) menighinianaCyclotella Stephanodiscus minutulus delicaticula Cymbella Diatoma moniliformis tenera Fragilaria vulpinaNavicula minusculaNavicula Nitzschia perminuta Nitzschia alpina 1 axis PCA 0.125 8.01 0.22 0.00 0.00 0.00 0.00 0.22 0.22 5.41 2.81 28.79 44.37 -0.80 0.375 10.35 0.24 0.00 0.00 0.00 0.00 0.47 0.00 4.00 5.41 21.88 38.82 -0.83 0.625 7.91 0.00 0.00 0.00 0.00 0.00 0.23 0.00 4.65 5.35 21.86 44.42 -0.92 0.875 5.59 0.00 0.00 0.56 0.00 0.00 0.28 1.12 13.69 0.00 25.70 30.17 -0.53 1.125 3.04 0.61 0.40 0.00 0.00 0.00 0.61 0.40 11.13 2.63 37.85 27.53 -0.55 1.625 7.51 0.64 0.00 0.00 0.00 0.00 0.21 0.00 17.60 1.50 33.26 22.53 -0.65 2.125 4.09 4.81 0.00 2.40 0.00 0.00 1.44 15.87 14.90 0.00 16.35 16.11 0.51 2.625 16.77 0.30 0.00 0.00 0.00 0.00 0.00 0.00 52.40 0.00 7.19 6.89 -0.50 3.125 3.87 18.15 8.04 6.55 1.19 6.25 11.01 2.38 11.01 0.00 15.18 8.63 1.96 3.625 1.61 16.28 0.00 0.00 0.00 2.98 0.69 10.78 4.36 0.00 17.89 11.47 0.92 4.125 6.34 28.40 3.63 5.44 8.16 18.43 5.14 0.00 6.65 0.00 9.06 12.99 2.10 4.625 15.41 0.34 0.00 0.34 0.00 1.03 1.37 0.00 27.05 0.00 29.79 14.73 -0.30 5.125 2.28 1.14 0.00 0.00 0.00 0.00 1.14 0.00 5.25 0.68 34.25 34.93 -0.42

172

Table C2. Diatom relative abundance (%) data for R-13 core.

eulerai

R-13

Depth

(cm) Amphora copulata cleve Cymbella botellus Cymbella phylleptaNavicula vulpinaNavicula minusculaNavicula Nitzschia perminuta Nitzschia alpina 1 axis PCA

0.125 3.67 3.02 0.00 8.42 7.34 8.42 29.16 31.53 -1.10 0.375 5.48 2.38 0.00 8.33 6.19 6.43 33.33 31.90 -0.99 0.625 5.05 3.27 0.00 4.46 9.66 8.02 27.04 36.11 -0.91 0.875 6.23 2.60 0.00 7.44 9.17 5.88 32.87 29.58 -0.96 1.125 4.37 4.96 0.00 6.94 15.28 6.75 28.57 25.20 -1.09 1.625 6.18 3.62 0.00 7.89 16.20 3.20 32.62 21.96 -0.83 2.125 2.01 2.01 0.00 8.04 14.29 4.91 38.84 23.88 -0.67 2.625 1.65 2.60 0.00 2.13 15.13 1.42 41.37 27.90 0.24 3.125 2.18 0.87 0.44 1.09 13.07 2.18 45.10 26.58 0.53 3.625 2.93 1.17 0.78 1.17 13.87 1.37 55.47 17.19 0.62 4.125 0.91 1.13 0.68 0.91 9.07 0.00 33.11 40.82 1.27 4.625 1.53 0.00 1.02 0.00 10.74 0.00 41.18 34.53 1.77 5.125 2.20 1.22 1.47 1.96 6.36 0.49 37.65 33.74 0.85 5.625 1.67 0.28 5.01 1.95 8.64 0.56 30.08 39.28 1.25

173

Table C3. Diatom relative abundance (%) data for R-1 core.

)

R-1

Nedium affineNedium

Depth spitsbergensis spitsbergensis (cm)

Achnanthes minutissimum Achnanthes minutissimum silicula Caloneis Encynopsis cesatii microcephala Cymbella designata Cymbella botellus Cymbella Cymbelopleura angustata var Cymbelopleura incerta var Eucocconeis flexella Eucocconeis leptostriata chiaraeNavicula 1 (cf. Sp 1 axis PCA 0.25 10.69 9.03 4.99 11.64 3.33 3.09 4.51 4.99 2.38 11.16 5.94 2.85 -0.90 1.25 9.91 8.99 4.77 9.17 4.40 3.67 6.42 3.30 6.79 5.50 4.04 4.77 0.05 2.25 5.14 10.88 2.42 9.06 9.37 4.23 5.74 6.34 6.34 1.51 3.02 3.63 1.61 3.25 14.80 6.55 6.77 11.63 2.96 2.33 6.34 4.02 5.29 2.96 4.86 5.29 -0.76

Table C4. Diatom relative abundance (%) data for R-2 core.

)

R-2

Nedium affineNedium

spitsbergensis Depth

(cm) var

Achnanthes minutissimum Achnanthes minutissimum microcephala Cymbella cf microcephalaCymbella botellus Cymbella Cymbelopleura angustata Nitzschia perminuta Nitzschia alpina 1 (cf. Sp 1 axis PCA 0.25 23.91 11.13 0.00 10.77 4.20 5.11 7.12 11.68 1.30 1.25 15.31 8.86 1.11 17.16 5.17 11.25 3.14 8.86 -0.16 2.25 12.13 6.64 4.58 18.99 5.72 10.07 2.97 8.70 -1.13

174

Table C5. Epilithic diatom relative abundance (%) data for R-10.

cf R-10 alpina

Sample

Year Diatoma moniliformis Nitszchia perminuta Nitszchia frustulum Nitszchia Nitszchia incognita silesiacum Cymbella Cymbella microcephala fogediiCymbella bottelusCymbella capucina Fragilaria 2011 0.00 8.46 3.12 21.60 0.45 12.25 35.63 0.67 7.13 0.67 2009 29.95 12.16 1.27 15.43 2.18 7.99 2.18 12.16 2.00 3.81 2008 0.00 8.77 0.78 29.24 0.19 14.04 7.99 4.68 18.13 1.17 2006 31.17 11.30 1.05 5.65 1.46 9.00 2.30 0.84 14.23 7.32 2002 11.76 18.41 0.95 12.71 13.28 19.73 1.14 2.66 0.95 11.01 1993 0.00 7.29 6.83 47.15 0.00 20.50 9.57 0.68 1.14 0.68

Table C6. Epilithic diatom relative abundance (%) data for R-11.

eulerai

- R-11

Sample Year

Diatoma moniliformis Nitzschia perminuta Nitzschia frustulum Nitzschia alpina Nitzschia palea Nitzschia incognita cleve Cymbella saprophila Fistulifera 2011 0.62 33.89 3.12 47.82 0.21 0.42 2.49 0.00 2009 29.93 29.25 5.22 7.94 0.45 9.07 7.94 4.31 2008 0.00 35.63 11.81 16.93 0.00 0.79 9.45 13.78 2006 3.88 63.50 11.65 6.80 1.17 2.33 0.39 0.58 2002 8.24 47.22 8.02 6.24 1.11 9.35 8.24 6.68 1993 0.00 26.05 1.74 20.10 6.95 13.90 0.00 29.28

175

Table C7. Epilithic diatom relative abundance (%) data for R-12.

var

R-12

Sample

Year Diatoma moniliformis Nitzschia perminuta Nitzschia frustulum Nitzschia alpina Nitzschia incognita Nitzschia linearis subtilis Amphora dusenii saprophila Fistulifera 2011 0.00 18.12 1.92 71.00 0.21 0.85 1.07 0.00 2009 0.61 45.64 19.68 25.35 0.00 0.00 0.00 3.85 2008 0.00 61.52 10.94 16.60 0.78 0.20 5.27 0.59 2006 3.64 58.43 5.75 4.02 10.73 6.51 0.00 4.79 2002 6.95 40.05 8.39 8.39 1.68 0.96 0.00 21.58 1993 0.00 9.56 0.00 5.15 0.00 0.00 0.00 79.41

Table C8. Epilithic diatom relative abundance (%) data for R-13.

R-13 cleve

Sample

Year Diatoma moniliformis Nitzschia perminuta Nitzschia frustulum Nitzschia alpina Nitzschia incognita Cymbella eulerai Achnanthes minutissima vulpinaNavicula 2011 0.46 37.53 3.20 27.00 3.20 2.29 17.39 1.14 2009 1.24 42.44 14.29 11.59 0.83 10.56 4.35 5.18 2008 0.60 70.78 6.76 5.77 0.00 5.37 0.80 3.78 2006 4.26 67.14 10.75 3.45 4.87 0.00 0.41 1.83

176

Table C9. Raw chironomid counts (whole head capsules) for R-12 core.

type

group

Oliveridia

Tvetenia

Paralimnophyes

gracilentus R-12

Depth (cm)

Corynoneura arctica Psectrocladius Hydrobaenus Eukiefferiella Limnophyes Metriocnemus hygropetricus Tanytarsus spp. Tanytarsini Procladius Chironomus plumosus Indeterminable 0.625 1 1 0 0 0 0 10 17.5 1 2 1.5 1.125 2 1.5 0.5 0 0.5 1.5 14.5 57 4 6.5 2 1.625 2 2 0 0 0 6 20 21 1 5 2.5 2.125 2 3 2.5 0 0 7.5 26 31 0 6 0 2.625 0 1.5 8.5 0 2 8 57.5 78 0 22 6 3.125 1 2 0 2 0 18.5 54.5 90 0 9 4.5 4.125 0 1 0 4 0 8.5 38 50 0 7 0.5 5.125 0 0 0 0 0 9.5 20 13.5 0 3 1.5 6.125 0 0 0 0 0 0 1 2 0 1 0 8.125 0 0 0 0 0 0 1 2.5 0 1 0 10.25 0 0 0 0 0 1 0 0 0 0 0 20.25 0 0 0 0 0 0 0 0 0 0 0

177

Table C10. Raw chironomid counts (whole head capsules) for R-13 core.

type

group

R-13

Depth (cm)

Corynoneura arctica Psectrocladius Hydrobaenus/Oliveridia Limnophyes/Paralimnophyes Metriocnemus hygropetricus Micropsectra radialis Tanytarsus gracilentus spp. Tanytarsini Procladius Chironomus plumosus Indeterminable 0.125 2.5 2.5 2 0 2 0 6 13.5 5.5 6.5 1 0.625 5 9.5 0 0 1.5 0 13 16.5 11 6 0 1.125 2 8.5 5 1 5.5 1 21 47 8 14 1.5 1.625 2 3 5 0 7.5 0 11 48.5 5 9 0 2.125 0 5.5 3.5 0 0 0 19 48 2.5 3 1.5 2.625 1 3.5 3.5 0 5.5 0 28.5 56 3 6 1 3.125 1 1.5 4 0 7 0 37 66.5 1 10.5 0 4.125 1 3 6 0 7 0 86.5 135 0 11 1 5.375 1 8 8.5 0 11.5 0 81 176 0 13.5 0.5 6.125 0 1 1 0 1 0 6 11 0 2 0

178

Table C11. Raw chironomid counts (whole head capsules) for R-1 core.

type

group

R-1

Depth (cm)

Corynoneura arctica Psectrocladius Hydrobaenus/Oliveridia Limnophyes/Paralimnophyes Metriocnemus hygropetricus Eukiefferiella/Tventnia spp. Tanytarsini Procladius Chironomus plumosus 0.25 1 1 1 0 1.5 2 0 0 0 1 0 0 0 0 0 0 0 0 1 2 1 1 0 0 1.5 3 0 0 1.5 3 0 0 0 0 2 5 0.5 1 3 4.5 1 0 0 0 5.5 2 0 0 1 6.5 0 3 0 0 3.5 1 0 0 0 7 0 4.5 0 3.5 5.5 1 0 2 2 8 0 2 1.5 1 9 0 0 2 1

179

Table C12. Raw chironomid counts (whole head capsules) for R-2 core.

type

group

R-2

Depth (cm)

Corynoneura arctica Psectrocladius Hydrobaenus/Oliveridia Limnophyes/Paralimnophyes hygropetricus Metriocnemus Eukiefferiella/Tventenia Tanytarsus gracilentus spp. Tanytarsini Procladius Chironomus plumosus Indeterminable 0 1 13 17.5 0 20.5 0 3 33 1 5 1.5 0.5 6 12 13.5 0 14 0 3 19 1 4 0 1.5 4 20 11 2 18.5 3 1 24 0 2 0 2.5 5 5.5 2 0 12.5 5.5 0 4 0 2 0 3.5 2 8 3 1 11 4.5 0 0 0 1.5 1.5 4.5 0 2 0 0 2 3 0 0 0 1 0 5.5 0 2 0 0 3 0 0 0 0 3 0 6.5 0 2 0 0 0.5 1.5 0 1 0 3 0 7.5 0 6.5 1.5 0 1 0 0 1.5 1 1 0 8.5 0 0.5 0 0 0.5 0 0 0 0 0.5 0 9.5 2 2 0 0 2.5 2.5 0 0 1 1 0 10.5 0 3 0 0 2.5 1.5 0 0 1 1.5 0

180

Appendix D Chapter 3 (Lake Ontario bird pond) raw counts

Table D1. Raw chironomid counts (whole head capsules) for the East Brother (EB) core.

type

spp.

type

EB iella/Teventia

Depth (cm)

Procladius Chironomus anthracinus Chironomus plumosus Endochironomus Microtendipes Cryptochironomus Parachironomus spp. Tanytarsini Pseudochironomus Stempellina Paratanytarsus Micropsectra spp. Orthocladiinae Cricotopus/Orthocladius Corynoneura Corynoneura arctica Metriocnemus hygropetricus Parametriocnemus/Paraphaenocladius Parachaetocladius Psectrocladius (Mesopsectrocladius) Phaenopsectra Limnophyes/Paralimnophyes Eukieffer Smittia Hydrobaenus/Oliveridia Indeterminable 0.25 0 12.5 0 0 0 0 0 1 9 0 3 0 0 3.5 0 0 0 0.5 0 0 0 5.5 1 0 0 4.5 1.25 0 8.5 4.5 1 1 1 0 5.5 3.5 1 3 0 0 2.5 1 0 0.5 1 0 0 0 3.5 1 1 0 1 2.25 1 4 3.5 0 0 0.5 0 1 6 0 1.5 0 0 2.5 0 1 1 1 0.5 0 0 3.5 0 0 0 0.5 3.25 0 5.5 5 3 0 0 0 1 7 0 2 0 0 5 0 0 0.5 0.5 0 0.5 0 5.5 2 0 0 1 4.25 0 6.5 2 2 0 0 0 2 5 0 1 0 0 4.5 1 0 0 0.5 0 0 0 3.5 0 2.5 0 0.5 5.25 0 4 5 1 0 0 0 1 3 0 3.5 1 1.5 3 0 0 0 0.5 0 0.5 0 3 0 6 0 0.5 6.25 0 7 1 0 1 0 0.5 0 5 0 9 0 1 2.5 0 1 0 0.5 0 1 1 0 0 6.5 0 0 7.25 1 4.5 2 0 0 0 0 4.5 3.5 0 3.5 0 0 5 0 0 1 0 0 0 0 2 0 5.5 0.5 0 8.25 0 9 1 3.5 0 0 0 4 4 0 4 0 2.5 9.5 0 1 0 0 0 0 0 2.5 0 6 0.5 1

181

Table D2. Raw chironomid counts (whole head capsules) for the Little Galloo (LG) core.

type

spp.

group

cf.

ilocerus Depth (cm)

Chironomus anthracinus Chironomus plumosus Microtendipes Lauterborniella Cladopelma Glyptotendipes pallens Orthocladiinae Cricotopus/Orthocladius Hydrobaenus Psectrocladius Heterotrissocladius Limnophyes/Paralimnophyes Parakieferiella Metriocnemus Metriocnemus terrester Props Parametriocnemus/Paraphaenocladius Smittia Pseudosmittia pigerChaetocladius Pseudochironomus spp Tanytarsini Procladius Tanypodinae Indeterminable 0.25 4 0 0 0 0 0 1 11 0 3 0 1 3 2.5 3 3 0 6.5 0 0 0.5 0 1 0 0.5 0.75 1 1 1 0.5 1 0 0 4.5 0 1 0 1 0 2 1.5 2 3 19 0 0 6 0 0 0 1.5 1.25 3 2.5 0 0 0 0 2 3 0 4.5 0 6 0 1.5 1 0 0 11 1 0 0.5 1 0 0 0 1.75 1 1 0 0 0 0 0 3.5 0 2.5 0 0.5 3 0.5 2 0 4 4 0 0 1 0 0 0 0 2.25 2 0 0 0 0 0.5 2 13 0.5 6.5 1 5.5 4 7 1.5 0 5 11 0 0 1 0 0 0 0 3.25 1 3 0 0 0 1 1 7 0 2.5 0 1 2 1 0.5 0 4 14 0 0 0 1 0 0 0 4.25 3 3 0 0 0 0 0 5.5 0 3 0 2.5 1 5 0 3 3 21 0 0 0 1 1 0 0 5.25 3 5.5 0 0 0 0 1 8.5 0 8.5 0 0.5 0 5 0 1 1 18 0 0 1 0 0 1 0.5 6.25 3 4 0 0 0 0 1 6.5 0 2.5 0 3 1 1.5 0 0 3 27 0 1 1.5 0 1 1 0 7.25 10 4 0 0 0 0 0 14 0 6.5 0 3 1 2.5 1 0 2 25 0 2 1.5 0 0 0 0.5

182

Table D3. Raw chironomid counts (whole head capsules) for the Main Duck Pond 2 (MD2) core.

type

spp.

type

MD2

Depth (cm)

Chironomini spp Chironomini Chironomus anthracinus Chironomus plumosus Microtendipes Cryptotendipes Cryptochironomus Cladopelma Glyptotendipes Endochironomus Polypedilum Lauterborniella/Zavriella Paratendipes Parachironomus spp. Tanytarsini Tanytarsus Tanytarus mendax Cladotanytarsus mancus Pseudochironomus Paratanytarsus Corynocera Micropsectra Stempellina spp. Orthocladiinae Cricotopus/Orthocladius Corynoneura Eukeifferiella/Tvetenia 0.25 0 4.5 1 2.5 1 0 0 0 2.5 0 0 0 1 7 2 0 2 9.5 17 0 0 0 0 4 0 0 1.25 0.5 4 1 2 6.5 0 0 0 1.5 0 0 0 1 2 0 0 2 3 15 0 0 0 0 13.5 0 1 2.25 0 2.5 0 1 2.5 1 0 1.5 0.5 1 0 0 0 11.5 0 0 8 8.5 15.5 0 3 0 1.5 9.5 2 0 3.25 0 2 1 3 4 1 1 2 2 0 0 0 0 19.5 0 0 4 7.5 32 0 0 0 0.5 20 0 0 4.25 0 3 1 7 4.5 0 0 2 0 0 0 0 0 14.5 0 0 1 8.5 20.5 0 2 0 1 24.5 1 0 5.25 0 1 1 1.5 0 0 0 5 1 0 0 0 1 9 0 9 0 17 13 0 4 0 0 12.5 0 0 6.25 0 1 1 0 0 0 0 0 2.5 2 1 0 0 28 0 0 1.5 16 0 2 0 0 0.5 13.5 0 0 7.25 0 4.5 5.5 0 2 0 0 5 1 3 3.5 1 0 36.5 0 0 2 24 0 0 0 0 3.5 24.5 0 0 8.25 0 1 2 0 0 0 0 8.5 1 2 0 0 0 16 1 0 0 8.5 0 2 0 1 2.5 22 1 0 9.25 0 1 0 0 0 0 0 2 1 1 0 0 0 24 0 0 1 9 0 2 0 0 0 11.5 2 0 10.25 0 4 1 0 0 0 0 0 0 4 1 0 0 51.5 0 0 2 27.5 0 0 0 0 0 17.5 4 0 15.25 0 1 7 0 0 0 0 0.5 0.5 1 0 0 0 20.5 0 0 5 15 0 0 0 0 0 21 0 0 20.25 0 0 2 0 0 0 0 4 3 4 0 0 0 19 0 0 3 8 0 1 0 0 2 26.5 3 0 25.25 0 0 2 0 0 0 0 1 1 1 0 0 0 11 0 0 1 9 0 0 0 0 0 11.5 1 0 30.25 0 0 2 0 0 0 0 1 3 2 0 0 0 6.5 0 0 0 10.5 0 0 0 0 0 7 1 0

(continued on next page)

183

group

MD2

(cont.)

Depth (cm)

Hydrobaenus/Oliveridia Metriocnemus Psectrocladius Parametriocnemus/Paraphaenocladius Parakeifferiella Smittia Limnophyes/Paralimnophyes Tanypodinae Procladius Type E Diamesinae Indeterminable 0.25 0 0 3 0 0 1 1 0 0 0 1 1.25 1 0 1.5 0 0 1 4 1 0 0 0 2.25 2.5 0 1.5 1 1 2 2 2 0 0 2 3.25 3 0 3 4 2 2 3 3 0 1 0 4.25 0.5 0 2.5 1.5 0 2 1 3 1 1 0 5.25 1.5 1 1 0.5 0 0 9 3.5 0 1 0 6.25 0 0 5.5 3.5 0 0 1 0 1 1 1 7.25 0 0 7.5 1 2 1 9 1 4 3 0 8.25 0 0 0.5 0.5 0 0 0 5 0 0 0.5 9.25 1 0 2 1.5 0 0 1 2 2 0 1 10.25 0 0 6.5 1.5 0 0 1 3 2 2 0 15.25 0 0 5 0 0 0 1 3 4 3 1 20.25 2 0 9 1 0 0 2 3 0 0 0 25.25 0 0 3 0.5 0 0 2 4 2 0 0 30.25 0 0 7 4 0 2 6 6 1 0 3

184

Table D4. Raw chironomid counts (whole head capsules) for the Calf (CF) core.

type

pulsus

cf.

cf cf

(early instar)

Depth (cm)

Chironomus Chironomus anthracinu Chironomus plumosus Cladopelma Cryptochironomus Cryptotendipes Dicrotendipes Dicrotendipes Einfeldia Endochironomus Glyptotendipes Glyptotendipes/Dicrotendipes Glyptotendipes barbipes Glyptotendipes pallens Glyptotendipes severini Microchironomus Microtendipes pedellus Parachironomus Parachironomus varus Phaenopsectra Polypedilum Zavreliella spp. Orthocladiinae Chaetocladius Corynoneura Corynoneura arctica 0.25 0 3 2 1 0 5 0 0 0 0.5 1 0 1 3 0 0 0 0 1 0 0 0 0.5 0 0 0 1.25 0 2.5 2 0 0 0 0 0 1 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 1 2.25 0 2 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 1 1 2 0 0 0 3.25 0 2 1 0 0 0 0 1 2 0.5 2.5 0 0 0 0 0 1 0 1 0 0 0 2 0 0 0 4.25 2 2 4 0 0 0 0 0 1 0 2 0 0 0 0 0 0 0 1 0 1 0.5 4 1 0 0 5.25 0 1 0 0 0 0 1 0 3 0.5 5 0 0 0 0 0 1 0 0 0 0 1 1 0 1 0 6.25 0 0 5 1 0 0 1 0 1 1.5 2 0 0 0 0 0 0 0 0 0 2 4 3.5 1 0 3 7.25 0 1 4 0 0 0 1 0 3 1 3.5 0 0 0 0 0 0 1 0 0 0 1 0 0 0 1 8.75 1 0 3 0 0 0 1 0 3 2 2 0 0 0 0 0 0 0 0 1 1 1 0 0 0 2 9.25 0 2 3 1 0 0 1 0 0 1 5 0 0 0 0 0 0 0 0 0 0 0 0 5 0 2 10.25 0 3 0 0 0 0 0 0 1 1.5 2 0 0 0 0 1 0 1 0 0 1 3 0.5 0.5 0 2 15.25 2 6 4.5 0 0 0 0 0 2.5 1.5 1 0 0 0 0 0 0 1 0 0 0.5 0 0 0 0 2 20.25 1 5 4 0 1 0 0 0 1 1.5 5.5 1 0 0 0 0 0 0 1 0 0 1 0 0.5 0 5 25.25 0 3 7 0 0 0 1 0 0 0.5 3.5 0 0 0 0 0 0 0 2 0 0 1 0 0.5 0 4 30.25 0 2 3 0 0 0 1 0 0 1 0 0.5 0 0 0 0 0 0 0 0 0 0 0 0 0 0 35.25 0 1 2 0 0 0 1 0 0 1 4.5 0 0 0 0 0 0 0 0 0 1 0 1.5 2 0 0 40.25 0 0 2 0 0 0 0 0 1 0 2 0 0 0 0 0 0 0 0 0 4 0 0 2.5 0 0

(continued on next page)

185

type

cf.

group

(cont.) spp. ni

Depth (cm)

Cricotopus/Orthocladius Eukiefferiella/Tvetenia Limnophyses/Paralimnophyes Metriocnemus terrester branchiolus Nanocladius Paracricotopus Parakiefferiella Parametriocnemus/Paraphaenocladius Propsilocerus Psectrocladius (monopsectrocladius) Psectrocladius Pseudosmittia Psuedorothocladius Smittia/Parasmittia Tanytarsi Cladotanytarsus mancus Micropsectra Paratanytarsus Pseudochironomus Tanytarsus Tanytarsus/Micropsectra spp. Tanypodinae Ablabesmyia Labrundinia Macropelopia Procladius 0.25 7 0 4 0 0 0 0 0 0 1 1.5 0 0 0 5 0 0 1 8 0 2 6 0 0 0 1 1.25 22 0 3 0 0 0 0 2 0 0 0 1 0 1 3 2 0 5 4 0 0 2 0 0 0 0 2.25 17 0 2 0 0 0 0 0 0 0.5 0.5 0 0 0 3 2 0 2 9 0 1 1 0 2 1 0 3.25 15 0 0.5 0 0 0 0 0 0 0 0 0 0 0 3 2 0 1 5 1 0 4 0 1 0 0 4.25 19 1 1 0 1 0 0 3 1 0 1.5 0 1 2 6 1 0 1 6 1 1 5 0 0 1 1 5.25 24 1 1.5 0 0 2 1 4 0 0 3 0 1 1 21 0 0 0 12 0 2 1 0 0 0 0 6.25 49 0 2.5 0 0 2 1 2 0 0 2 0 0 1 12 3 1 1 13 0 0 4 0 2 1 0 7.25 22 0 0.5 0 0 3 0 0 0 0 1.5 0 1 3 8 0 1 0 13 0 0 4 0 0 0 0 8.75 23 0 2.5 0 0 2 0 1 1 0 1.5 0 0 1 8 0 0 3 13 0 0 1 1 1 0 1 9.25 38 0 2.5 0 0 2 0 1 0 0 3 0 1 0 5 0 0 1 6 0 1 0 0 0 0 0 10.25 27 0 4 0 1 1 0 0 0 0 0 0 0 0 7 1 0 11 15 1 0 2 0 0 1 0 15.25 34 0 3.5 1 0 2 1 0 0 0 0 0 1 2 11 1 0 8 13 0 0 2 0 0 0 0 20.25 34 0 6 0 0 3 0 1 0 0 5.5 0 0 2 27 4 0 7 8 0 0 3 0 2 0 0 25.25 46 0 5.5 0 2 2 0 0 0 0 0.5 0 1 3 17 1 0 5 7 0 0 3 0 0 0 1 30.25 25 0 1 0 0 2 0 0 0 0 1 0 0 2 12 0 0 5 2 0 0 0 0 1 0 0 35.25 9 0 9.5 0 0 1 0 1 0 0 0.5 0 0 3 10 0 0 10 8 0 1 1 0 0 0 0 40.25 10 0 15 0 1 0 0 0 0 0 0 0 0 0 8 1 0 0 6 0 0 5 0 0 0 0

(continued on next page)

186

CF (cont.)

Depth (cm)

Tanypus Sp. 1 Podonominae Indeterminable 0.25 0 1.5 0.5 1.25 0 1 0 2.25 0 0 0.5 3.25 1 3.5 0 4.25 0 2.5 1.5 5.25 3 3 0 6.25 2 4.5 2 7.25 0 3.5 1.5 8.75 3 1.5 0.5 9.25 3 1.5 1 10.25 1 2.5 1 15.25 5 7.5 0 20.25 3 5.5 0 25.25 2 8.5 1 30.25 2 1.5 0 35.25 0 0 0.5 40.25 1 0 0

187

Table D5. Raw chironomid counts (whole head capsules) for the Pigeon (PGN) core.

type

hygropetricus

PGN

Depth (cm)

Microtendipes Pseudochironomus spp. Orthocladiinae Cricotopus/Orthocladius Metriocnemus Parametriocnemus/Paraphaenocladius Psectrocladius Limnophyes/Paralimnophyes Smittia Indeterminable 0.25 0 1 1 0.5 1.5 0.5 0 3 9.5 0 1.25 1 0 0.5 1.5 0 0 0 0 19 0.5 2.25 0 0.5 0.5 0 0 0 0.5 0 27 0 3.25 0 0.5 0.5 0 0 0 0 0 19 0 4.25 0 0.5 0.5 0.5 0 0 0 0 17 0 5.25 0 0.5 0 0 0 0 0 0 5.5 0.5 6.25 0 0 1.5 0 0 0 0 0 1 0

188

Appendix E Chapter 4 (Niven Lake) raw counts

Table E1. Raw concentration data of sterols in dry sediment (mg/g) for the Niven Lake core.

Depth (cm)

Coprostanol Epicoprostanol Coprostanone Cholesterol Cholestanol Cholestanone Sitosterol Stigmastanol 0.25 4.51 1.18 2.63 29.84 9.35 0.94 24.32 18.75 1.25 3.83 0.91 1.83 13.91 6.44 0.60 12.70 14.31 2.25 3.91 0.94 1.54 6.30 5.96 0.63 9.18 13.61 3.25 3.37 0.74 1.02 6.83 9.21 0.50 9.19 16.53 4.25 2.75 0.87 0.74 3.16 3.55 0.88 5.64 11.00 5.25 5.23 0.78 0.92 4.14 4.89 0.38 6.10 13.64 6.25 6.53 1.25 1.45 2.95 5.10 0.44 4.47 12.34 7.25 3.90 0.72 0.79 3.08 4.77 0.33 4.94 15.66 8.25 3.60 0.56 0.84 2.80 4.26 0.29 3.33 11.86 9.25 4.74 0.73 0.91 2.65 5.59 0.35 4.35 16.17 10.25 3.99 0.63 0.93 2.53 4.01 0.31 6.67 11.45 11.25 7.25 1.05 1.36 2.61 7.46 0.44 8.49 15.30 12.25 3.55 0.54 0.78 1.44 4.25 0.30 4.54 9.96 14.25 2.10 0.32 0.49 1.00 2.52 0.20 2.76 7.59 16.25 1.23 0.17 0.27 0.58 1.25 0.13 1.22 5.56 18.25 0.56 0.09 0.14 0.42 0.55 0.07 0.75 3.08 20.25 1.45 0.19 0.33 0.48 1.39 0.15 1.05 5.68

189

Table E2. Raw counts for diatoms (valves) from the Niven Lake core.

inaequalis

pusillum

minutissimum

descripta

silesiacum

cf. cf.

delicatula

Depth (cm)

Achnanthidium Achnanthidium Platessa conspicua Planothidium lanceolata Psammothidium curtissimum Eucocconeis flexella Rossithidium pusillum Rossithidium Achnanthes saccula Amphora holsatica Amphora ovalis Amphora pediculus Amphora veneta Amphora libyca Brachysira neoexilis Brachysira procera Encyonopsis microcephala Encyonopsis descripta Encyonopsis Encyonema silesiacum Encyonema Encyonopsis cesatii Cymbopleura tumida Cymbella pusilla Navicymbula Cymbella Cymbopleura lapponica 0.25 12 22 9 2 0 0 0 0 2 0 0 0 0 1 0 1 0 0 0 0 0 0 0 0 0 0 1.25 6 33 40 2 0 0 0 0 0 1 0 0 0 4 0 2 0 0 0 0 0 0 0 0 0 0 2.25 3 22 68 2 0 0 0 0 0 0 5 0 0 8 0 5 0 0 0 0 0 1 0 0 0 0 3.25 30 21 50 9 0 0 0 0 0 0 0 10 2 8 0 4 0 0 1 0 0 0 0 0 0 0 4.25 60 2 23 10 0 0 0 0 0 0 0 3 0 57 0 59 0 0 2 0 2 0 0 0 0 0 5.25 55 0 19 4 0 0 0 0 0 0 0 4 0 45 5 44 7 0 5 0 6 0 0 0 0 0 6.25 99 4 19 12 2 0 0 0 0 0 0 4 0 89 7 52 0 0 2 0 7 0 0 0 0 0 7.25 64 4 3 0 0 0 0 0 0 0 0 2 0 41 4 43 0 0 2 0 0 0 1 1 0 0 8.25 93 5 13 8 0 0 0 2 0 0 0 5 0 50 3 51 3 0 4 0 2 0 0 2 1 0 9.25 74 2 3 0 0 0 0 0 0 0 0 2 0 52 5 74 4 0 2 0 4 0 0 0 0 0 10.25 93 0 1 0 0 1 0 0 0 0 0 4 0 85 3 78 2 0 0 0 3 0 0 0 1 0 11.25 86 3 3 4 0 0 0 0 0 0 0 1 0 62 6 62 0 0 2 0 7 0 0 0 4 0 12.25 106 0 3 5 0 4 0 0 0 0 0 0 0 76 6 68 2 0 0 0 4 0 2 0 0 5 13.25 86 0 2 3 0 2 0 0 0 0 0 0 0 62 5 65 5 0 0 0 11 0 0 0 0 0 14.25 112 0 8 8 1 4 0 0 0 0 0 0 0 50 5 49 8 0 1 2 5 0 0 0 2 0 15.25 119 0 0 10 0 0 2 0 0 0 1 0 0 68 15 30 0 8 1 0 12 0 0 0 0 0

(continued on next page)

190

11.5

8.5

venter

lemanica

var.

sensu lato

var.

vaucheriae

rossii michiganiana

silicula

stelligera

bodanica

cf. cf.

cf. cf. cf. (Cont.) cf.

unknown broken valve broken 8 unknown valve broken 11 unknown

Depth (cm)

Encyonema hebridicum Cyclotella Lindavia Discostella Lindavia michiganiana Lindavia ocellata Cyclotella bacillum Caloneis Caloneis silicula Caloneis schumanniana Caloneis Cocconeis placentula Denticula kuetzingii Eunotia Eunotia Eunotia bilunaris Eunotia pectinalis Epithemia argus Epithemia sorex rhomboides Frustulia mesolepta Fragilaria Staurosira construens brevistriata Pseudostaurosira pinnataStaurosirella vaucheriae Fragilaria Fragilaria 0.25 0 0 0 1 0 0 0 2 0 0 0 73 0 0 0 0 0 9 0 1 149 2 0 0 0 0 1.25 0 0 0 0 0 0 0 0 0 0 0 78 0 0 0 0 0 10 0 0 84 0 0 0 0 0 2.25 0 1 0 0 0 0 0 2 0 0 0 77 0 0 0 0 0 7 0 0 50 0 0 0 0 0 3.25 0 0 0 0 0 0 0 0 0 0 0 118 0 0 0 0 0 0 8 0 60 0 3 5 0 0 4.25 0 0 1 18 3 0 0 0 0 0 0 87 0 0 0 1 0 0 2 0 12 0 1 0 2 0 5.25 0 1 0 6 5 0 0 0 0 0 0 74 0 0 0 0 0 0 2 0 16 0 4 1 0 0 6.25 0 3 0 2 7 0 0 0 0 0 0 92 2 0 0 2 2 0 6 0 13 0 3 2 0 0 7.25 0 0 0 0 7 0 0 0 0 0 0 94 0 0 0 0 0 2 0 0 8 7 7 0 0 0 8.25 0 1 0 2 2 1 0 0 0 0 0 80 0 1 0 0 0 1 0 0 2 0 7 2 12 0 9.25 0 0 0 14 9 1 0 0 2 0 0 71 0 0 0 0 0 2 0 0 4 0 1 0 7 0 10.25 0 3 0 7 10 0 0 0 0 0 0 49 0 0 0 0 0 0 0 0 1 0 4 1 0 0 11.25 0 0 0 3 6 0 0 0 0 0 1 61 0 0 2 0 0 0 0 0 0 0 6 5 0 4 12.25 0 1 0 8 19 0 0 0 0 0 0 38 0 0 0 0 0 0 0 0 2 0 6 0 0 0 13.25 3 0 0 7 8 0 2 0 0 2 0 19 0 0 0 0 0 2 0 0 4 0 3 0 0 0 14.25 3 0 0 20 8 0 0 0 0 0 0 18 0 0 0 0 0 0 0 0 2 0 0 4 0 0 15.25 11 1 0 2 10 0 0 0 0 0 0 10 0 0 0 0 0 0 0 0 1 0 3 0 0 0

(continued on next page)

191

(Ruhland 2003) et al. (Ruhland 2015) al. et (Wetzel

(Ruhland et al. et al. 2003) (Ruhland 2015) et(Wetzel al.

olivaceum

atomoides

cf. cf.

halophila modica

minima

cf. cf.

(Cont.)

cf. cf.

Depth (cm)

Fragilaria capucina Fragilaria leptostauron Staurosirella Gomphonemaparvulum Gomphonemapumilum olivaceumGomphoneis Gomphoneis Gomphonemaangustatum Gomphonema truncatum Sellaphora cryptocephalaNavicula cryptotenella Navicula radiosaNavicula pupula Sellaphora seminulum Sellaphora seminulumSellaphora seminulumSellaphora Eolimna minima Eolimna minima Eolimna minima Eolimna Navicula halophila Craticula Craticula submuralis Navicula angusta Navicula incertataNavicula 0.25 0 0 0 0 2 0 0 0 2 5 14 0 5 7 4 0 29 11 0 0 7 0 0 0 0 0 1.25 0 0 0 0 0 0 9 0 2 11 10 1 3 17 5 0 45 0 0 0 0 1 0 0 0 0 2.25 0 0 2 0 0 0 0 0 0 8 7 3 2 21 2 0 39 0 0 0 0 0 0 1 7 7 3.25 0 0 4 0 0 0 0 0 0 7 9 0 0 0 0 8 26 0 0 0 0 2 0 11 0 2 4.25 0 0 0 0 0 0 0 1 0 7 14 2 4 0 0 19 0 0 19 0 0 0 0 9 0 0 5.25 0 0 0 6 0 0 0 0 0 5 10 5 3 0 0 17 0 0 21 0 0 0 0 9 0 0 6.25 0 0 7 0 0 0 0 1 0 7 12 8 1 0 0 13 0 0 18 0 0 0 0 0 0 0 7.25 8 0 0 2 0 0 0 0 0 1 12 8 4 0 0 18 0 0 8 0 0 0 0 1 0 0 8.25 6 0 0 0 0 0 0 3 0 2 6 6 2 0 0 7 0 0 11 0 0 0 0 0 0 0 9.25 29 0 2 0 0 0 0 4 0 9 9 6 1 0 0 3 0 0 1 0 0 0 0 0 0 0 10.25 3 0 0 0 0 2 0 5 0 1 17 9 2 0 0 5 0 0 0 0 0 0 0 0 0 0 11.25 3 0 2 0 0 0 0 1 0 8 10 4 2 0 0 6 0 0 2 0 0 0 0 0 0 0 12.25 4 0 0 0 0 0 0 2 0 10 24 5 5 0 0 9 0 0 2 0 0 0 0 0 0 0 13.25 0 0 0 0 0 0 0 1 0 4 14 14 1 0 0 3 0 0 3 0 0 0 0 0 0 0 14.25 0 1 5 0 0 0 0 2 0 4 9 14 2 0 0 0 0 0 0 6 0 0 0 0 0 0 15.25 0 0 1 0 0 0 8 0 0 2 8 7 9 0 0 0 0 0 0 4 0 0 2 0 0 0

(continued on next page)

192

)

faintones

(

hantzschii hantzschii parvus

jaagii

cf. cf. cf. cf.

cf. cf.

perminuta fonticola

incertata erifuga cincta constricta

pseudoscutiformis

cf. cf. cf.

cf. cf. cf. cf. cf.

cf. cf. (Cont.)

Depth (cm)

Navicula subminuscula Craticula venetaNavicula subrotundata Navicula kuelbsiiNavicula jaagiiKobayasiella Kobayasiella indifferens Navicula erifugaNavicula Navicula rhynchocephala Navicula Navicula Navicula Cavinula Nitzschia Nitzschia perminuta Nitzschia frustulum Nitzschia fonticola Nitzschia Nitzschia amphibia Nitzschia palea Stephanodiscus Stephanodiscus Stephanodiscus hantzschii Stephanodiscus alpinus Stephanodiscus 0.25 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 26 0 0 0 0 2 1 1.25 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 53 3 4 0 0 0 0 2.25 0 15 0 0 0 0 0 0 0 0 0 0 0 0 0 0 2 0 0 40 1 7 0 0 0 0 3.25 0 9 12 1 2 0 0 0 0 0 0 0 0 0 0 0 1 0 0 12 0 9 13 0 0 0 4.25 0 5 5 0 0 0 0 2 0 0 0 0 0 0 4 0 4 16 0 12 0 6 3 0 0 1 5.25 0 11 4 0 0 2 0 1 0 0 0 0 0 0 0 0 1 5 0 4 5 6 3 0 0 0 6.25 0 8 12 0 0 1 0 0 0 0 0 0 0 0 0 0 2 11 0 13 3 0 0 8 0 0 7.25 0 11 7 0 0 0 4 0 1 0 0 0 0 0 0 0 6 13 0 3 5 7 5 0 0 0 8.25 0 8 2 0 0 0 0 0 0 0 0 0 0 0 0 0 4 13 0 2 1 10 0 0 0 0 9.25 0 12 6 0 0 1 0 2 2 0 0 2 0 0 0 0 4 13 0 2 7 4 0 0 0 0 10.25 0 4 6 0 0 4 0 0 7 0 0 0 0 0 0 0 6 18 0 5 0 1 0 0 0 0 11.25 2 8 9 0 0 0 0 0 0 0 0 0 0 0 0 0 2 9 0 0 5 1 0 0 0 0 12.25 0 3 0 0 0 1 0 0 3 0 0 0 1 1 0 3 8 15 0 7 0 2 0 0 0 0 13.25 0 2 0 0 0 3 0 0 5 0 0 0 0 0 0 0 13 39 5 2 0 1 0 0 0 0 14.25 0 4 0 0 0 0 4 0 13 0 0 0 0 0 0 0 2 19 0 8 2 2 0 0 0 0 15.25 0 0 0 0 0 0 0 0 0 5 2 0 0 0 0 0 9 35 0 2 0 1 0 0 0 0

(continued on next page)

193

str. IIIp

(Cont.)

Depth (cm)

Stephanodiscus minutulus Stephanodiscus medius angustataTryblionella flocculosa Tabellaria Stauroneis phoenicenteron Chrysophyte Cyst 0.25 0 0 0 0 0 9 1.25 0 0 0 0 0 22 2.25 0 0 0 0 0 41 3.25 9 0 0 0 0 41 4.25 0 0 0 0 0 38 5.25 2 0 0 0 1 47 6.25 0 0 0 0 0 56 7.25 0 3 0 0 0 47 8.25 0 0 0 0 0 51 9.25 0 0 0 1 0 35 10.25 0 0 2 0 0 43 11.25 0 0 0 0 0 31 12.25 0 0 0 0 0 44 13.25 0 0 0 0 2 28 14.25 0 0 0 0 2 43 15.25 0 0 0 0 2 42

194

Table E3. Raw counts of chironomids (whole head capsules) for the Niven Lake core.

type

- type

type

- type

type

cf.

omus varus

cf.

Depth (cm)

Chironomini indeterminable Chironomini early instar Chironomini Chironomus anthracinus Chironomus plumosus lateralis Cladopelma Cryptochironomus Dicrotendipes Einfeldia Einfeldia Einfeldia pagana Endochironomus albipennis Endochironomus Endochironomus impar Endochironomus tendens Glyptotendipes Glyptotendipes barbipes Lauterborniella Microtendipes pedellus Pagastiella Parachiron Polypedilum nubeculosumPolypedilum nubiferPolypedilum sordensPolypedilum Sergentia indeterminable Orthocladiinae Chaetocladius 0.25 0 0 0 6.5 0 0 2 0 0 2 0 2 0 1 0 0 0 1 1 3 0 0 0 0 1 0.5 0 1.25 0 1 2 9 0.5 0 5.5 5 0 1 0 0 0 0 0 0 1 0 0 1 4 0 0 1 0 0 0 2.25 0 0 1 10 0 0 0 1 0 2 0.5 1 0 0 0 0 1 0 0 0 0 1 0 0 0 0 0 3.25 0 2 1 13 0 0 4.5 3 0 1 0 2 0 0 0 0 1 2 0 2 0 2 0 0 1 0 0 4.25 0 8 0 6 2 0 4.5 0.5 0 1 0 1 0 0 0 0 1 2 2 0 0 4 1 0 0 1 0 5.25 0 4 1 5 2.5 0 4 1 0 3 0.5 1 0 1 0 0 1 4 3 2 1 3 2 0 2 0 0 6.25 0 2 0 2 0 0 2 1 0 1 0 0 0 0 1 0 0 1 2 0 0 4 3 0 1 0 0 7.25 1 2 0 3 0 0 2 1 0 0 0 1 0 0 0 0 0 0 0 0 0 4 0 0 0 2 1 8.25 0 2 2 7.5 5 1 3 2 0 0 0 0 0 0 0 0 2 2 1 0 1 3 0.5 0 1.5 0 0 9.25 0 5 2 5 2 0 3 1 0 0 0 0 0.5 0 2 0 1 1 2 0 0 8 0 0 1.5 0 0 10.25 1 2 3 3 3 0 6 1 0 1 0 0 0 0 0 0 1 3 5 0 0 4 0 0 2 0 0 11.25 0 2 6 3.5 1 0 3.5 1 0 0 0 0 0 0 0.5 3 0 5 2 0 0 11 0 0 0 0 0 12.25 0 1 1 3 3 0 4 3 0 0 0 0 0 0 0 0 1 4 1 0 0 13 0 0 0 0 0 13.25 0 0 2 6 2 0 6 1 0 0 0 0 0.5 0 1 0 3 2 5 1 0 9 0 0 1.5 0 0 14.25 0 0 1 4 2 0 8 1 0 0 0 2 0 0 1 0 0 2 0 0 0 10 0 0 0.5 0 0 15.25 0 0 1.5 4 4 0 7.5 3.5 0 0 0 3 0 0 1 0 0 3 1 0 0 8 0 0 0 0 0

(continued on next page)

195

type

psilopterus

type

type/

type type

- -

type

type cf.

type

group

cf.

early instar early

(Cont.) cf.

Depth (cm)

Corynoneura arctica Cricotopus/Orthocladius Hydrobaenus Limnophyes/Paralimnophyes Metriocnemus terrester Nanocladius nigra Parakiefferiella Psectrocladius (monopsectrocladius) Psectrocladius (psectrocladius) Psectrocladius barbatipes Psectrocladius barbimanus Psectrocladius Psectrocladius sordidellus indeterminable Tanypodinae Arctopelopia Djalmabatista Guttopelopia Natarsia Procladius Procladius Tanypus indeterminable Tanytarsini Cladotanytarsus mancus Micropsectra Micropsectra contracta Micropsectra insignobilus Micropsectra junci 0.25 6 0.5 0 0 0 0 0 0 0 0 0 1 8 0 1 3 0 0 3 0 0 3 0 0 0 0 0 1.25 6 0 0 0 0 0 0 1 1.5 0 0 0 8.5 1 5 4 0 0 2 0 0 3 1 0 0 1 0 2.25 5 2.5 0 0 0 0 0 0 6.5 0 0 0 10 3 7 3 0 0 5 0 0 2 1 1 0 0 0 3.25 7 2.5 0 0 0 0 0 0 0 1 0 0 10 0 4 3 1 0 2 0 1 3 1 0 0 0 0 4.25 4.5 3 0 0 1 0.5 0 0 1.5 3.5 0 0 8.5 6.5 2 0 0 0 6 2 1 11 0 1 0 0 0 5.25 6 3 0 0 0 0 0 0 0 1.5 0 0 7.5 2 7 1 0 0 5 1 3 13 0 2 0 0 0 6.25 7 3.5 0 0 0 0 0 0 1.5 1.5 0 1.5 6.5 3 1 1 0 0 1 2 0 2 2 0 0 3 0 7.25 3 3 0 0 0 0 0 0 0 1.5 0 2 5 0 3 0 0 0 0 0 0 10 2 0 1 0 0 8.25 3 0.5 0 0 0 0 0 1 0 0.5 0 0.5 5.5 2 5 2 0 0 4 1 0 4 3 0 0 1 0 9.25 2 1.5 0 0.5 0 0 0 1 0 0 0 0 9 1 4 0 0 0 5 0 0 11 1 0 0 6 0 10.25 0 2 0 0 0 0 0 0 0 1 0 0 5 2 3 3 0 0 4 0 0 4 0 2 0 1 0 11.25 3 3 0 0 0 0 0 0 0 2 0 0 8 1 4 3 0 0 4 4 0 12 5 0 0 4.5 0.5 12.25 1 1 0 0 0 0 1 0 0 4 0 0.5 6 2 5 0 0 0 5 1 0 7 2 0 1 2 0 13.25 3 8.5 1 0 0 0 0 0 0 1.5 0 2 6.5 2 2 2 1 1 3 1 0 5 4 0 0 1 0 14.25 1 3.5 0 0 0 0 0 0 0 1 0 1 5 0 5 0 0 0 4 1 0 5 2 0 1 1 0 15.25 1 5 0 0 0 0 0 0 0 1 0 1 5 2 4 0 0 0 5 0 0 12 0 0 0 1 0

(continued on next page)

196

type cf. type 2

type

- -

type

type cf.

type

type cf. type -

type

cf.

(Cont.)

Depth (cm)

Micropsectra pallidula Micropsectra Micropsectra radialis Neozavrelia Paratanytarsus Pseudochironomus Stempellinella/Zavrelia Tanytarsus/Micropsectra Tanytarsus Tanytarsus chinyensis Tanytarsus glabrescens Tanytarsus lactescen Tanytarsus lugens Tanytarsus mendax Tanytarsus nemorosus Tanytarsus pallidicornis Tanytarsus pallidicornis Indeterminable 0.25 0 0 0 0 0 0.5 0 1 0 0 0 0 1 0 0 1 0 1.25 0 0 0 1 0.5 0 0 2 0 1 1 1 0 0 0 0 0 2.25 0 1 0 0 0.5 0 0 0 0 0 0 0 0 0 0 0 0 3.25 0 3 0 0 0 0 0 0 0 4 1 0 3 0 0 3 0 4.25 0 3.5 0 0 1 0 0 2.5 0 0 0 0 2 0 0 0 0 5.25 2 11 0 1 1 0 0 2 0 1 0 3 8 0 0 2 0 6.25 0 1 0.5 1 0 0 0 0 0 5 0 0 0 0 0 0 0 7.25 0 3 0 0 0 1 1 0 1 0 1 0 4 0 0 0 0 8.25 0 2.5 0 1 0.5 0 0 1 0 2 0 1 2.5 0 0 1 0 9.25 1 2 0 2 0.5 0 2 1 0 1 1 1 0 0 0 1 0 10.25 0 0 0 0 0 0 2 0 0 3.5 0 0 1 0 1 0 0 11.25 0 1 0 0 2 2 3.5 0.5 0 2 0 1.5 8 0 0 9.5 0.5 12.25 0 1 0 0 0 1 2 1 0 2 2 1 5.5 1 3 4 0 13.25 0 6 0 0 1 1 5.5 0 0 2 4 3 4 0 0 5 1 14.25 0 3 0 0 0 0 2 1 0 2 1 1 2 0 1 0 0 15.25 0 1 0 0 0 0 4 0 0 2 1 1 6 0 0 3 0

197