Assessment of an Adult Lake Sturgeon Translocation (Acipenser Fulvescens) Reintroduction Effort in a Fragmented River System

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Assessment of an adult lake sturgeon translocation (Acipenser fulvescens) reintroduction effort in a fragmented river system

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

TRENT UNIVERSITY Peterborough, Ontario, Canada © Copyright by Maggie L.E. Boothroyd 2017 Environmental and Life Sciences M.Sc. Graduate Program January 2018 Abstract Assessment of an adult lake sturgeon translocation (Acipenser fulvescens) reintroduction effort in a fragmented river system Maggie L.E. Boothroyd North American freshwater fishes are declining rapidly due to habitat fragmentation, degradation, and loss. In some cases, translocations can be used to reverse local extirpations by releasing species in suitable habitats that are no longer naturally accessible. Lake sturgeon (Acipenser fulvescens) experienced historical overharvest across their distribution, leading to endangered species listings and subsequent protection and recovery efforts. Despite harvest and habitat protections, many populations do not appear to be recovering, which has been attributed to habitat alteration and fragmentation by dams. In 2002, 51 adult lake sturgeon from the Mattagami River, Ontario, Canada were translocated 340 km upstream to a fragmented 35 km stretch of the river between two hydroelectric generating stations, where sturgeon were considered extirpated. This study assessed the translocation effort using telemetry (movement), demographics and genetic data. Within the first year, a portion of the radio-tagged translocated individuals dispersed out of the release area, and released radio-tagged individuals used different areas than individuals radio-tagged ten years later. Catches of juvenile lake sturgeon have increased over time, with 150 juveniles caught within the duration of this study. The reintroduced population had similar genetic diversity as the source population, with a marked reduction in effective population size (Ne). The results indicate that the reintroduction effort was successful, with evidence of successful spawning and the presence of juvenile lake sturgeon within the reintroduction site. Overall, the results

ii suggest adult translocations may be a useful tool for re-establishing other extirpated lake sturgeon populations.

Keywords: Lake sturgeon (Acipenser fulvescens), endangered species, translocation, reintroduction, conservation, telemetry, genetics.

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Dedication

This thesis is dedicated to my mom, Chris Cordy who has supported me in all my endeavours throughout my life. And who cared for me in the many months of recovery after meningitis. I wouldn’t have finished without you.

Acknowledgements

I am extremely grateful for the diverse, talented people I got to work with over the course of this thesis. I would like to thank everyone at the Fish Genetics Lab: the training

I have received in the lab over the years prepared me to efficiently complete the genetic component of this project. Over the years I have received so much training and help from

Kristyne Wozney, Caleigh Smith, Anne Kidd, and Cait Nemeczek, and it has been such a fun lab to work in. Additionally, the fish genetics lab provided me with information from the source population that allowed for the genetic diversity comparison. Also to my lab friend Natasha Serrao: although we had little overlap in our Masters degrees she is a great friend and was always a phone call away when I was having issues.

I have so many people to thank for aiding in the field work necessary to complete this project. There are so many individuals at the Ontario Ministry of Natural Resources and Forestry (OMNRF) that aided in the field work for this project. A special thanks to

Kevin Kilgore, Tarryn Adams, Adam Rupnik, Alyssa Dufresne, and Emma Harten from

MNRF for always accompanying me in the field and being so much fun to work with. I would also like to thank the many volunteers from Timmins Fur Council, Club

Navigateur-La Ronde, Glencore Canada Corporation, Kidd Operations, and Lake Shore

Gold, particularly Kevin Roy from the Timmins Fur Council for helping extensively during the 2015 netting effort. Also, thank you to Glen McFarlane and Heidi Etzel who both managed the Timmins District office during my project and provided me with the equipment necessary to complete my field work.

There were a number of very vital people working on this project for the duration of my involvement and before. A huge thank to Laurent Robichaud of Club Navigateur-

La Ronde for his guidance, knowledge, commitment to the project, and taking me to see other important sturgeon habitat on my days off. Thank you to Charlie Hendry of the

OMNRF for his early work that made this project happen, the tracking that allowed for analysis of post-release movement, and always being available for questions despite retiring. Thank you to Derrick Romaine of the OMNRF for being passionate about this project and acquiring the funding necessary to complete it, setting me up with the project, teaching me the various field techniques required to complete this project, and answering my questions whenever they arise. Thank you to Dan Gibson of Ontario Power

Generation (OPG) for reviewing my data chapters in a timely manner and returning comments that strengthened my thesis.

A huge thank you to my family, Chris Cordy, Ally Boothroyd, and Courtney

Boothroyd for their ongoing support during this project, and the encouragement while I was in the hospital, and the many vitamins and support provided during recovery. Thank you to Luna and Oliver, my senior dogs that always seem to give me perspective in times of stress. To my dad, who passed away before I started this degree but took a big interest in my career path in his last months. Thank you to my nephew, Deakin, who had me come speak to the grade 2-5s at his school - seeing the next generation’s interest in conservation is extremely motivating.

Lastly, I would like to thank my graduate committee for all their guidance and help. I thank Tim Haxton, my sturgeon expert for all the comments and resources he provided me. Especially his suggestions and guidance regarding Chapter 2, which made it a much more interesting and impactful chapter. I thank Tom Whillans for his guidance and suggestions, and for his involvement in ecological endeavours too numerous to

vi mention, it is an inspiration. A huge thank you to Chris Wilson, for taking me on as a shy undergrad and offering me so many opportunities. Your ongoing belief in my aptitude for research kept me motivated despite my memory issues post-meningitis. Thank you for providing me with your insight, guidance, patience, humour, unfailing support, and extensive knowledge.

Funding was provided by Ontario Power Generation (OPG), Lake Shore Gold, and Glencore Canada Corporation, Kidd Operations, the OMNRF, the Ministry of

Environment (MOE) Ontario Community Environment Fund, and managed by the

Wintergreen Fund for Conservation.

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Table of Contents Abstract ...... ii

Dedication ...... iv

Acknowledgements ...... v

List of Tables ...... x

List of Figures ...... xi

List of Appendices ...... xiii

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

Literature cited ...... 11

Chapter 2: Post-release dispersal, and spawning movements ...... 19

Introduction ...... 20

Methods ...... 26

Data analysis ...... 32

Results ...... 33

Discussion ...... 37

Conclusions ...... 46

Literature Cited ...... 49

Chapter 3: Demographic and genetic assessment ...... 75

Introduction ...... 77

Methods ...... 79

Results ...... 89

Discussion ...... 93

Conclusions ...... 99

Literature Cited ...... 101

Chapter 4: General Discussion ...... 116

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Literature cited ...... 132

Appendices ...... 139

List of Tables

Table 2.1: Fork length (cm), total length (cm), weight (g), and age estimates for adult lake sturgeon that underwent transmitter surgery for the Mattagami River Lake Sturgeon Restoration Project in 2002, 2011, and 2013. Information on year of surgery, and tags present and/or applied to fish ...... 60

Table 2.2: Annual comparison of spring water temperature (°C) and discharge (m3/s) within the Mattagami River, Ontario for 2014, 2015, and 2016 ...... 62

Table 2.3: The number of radio-tagged lake sturgeon tracked at each location along the spawning movements in the reintroduced Mattagami River, Ontario population for 2014, 2015, and 2016 ...... 62

Table 2.4: T-test results for inter-annual differences in the timing of lake sturgeon leaving wintering hole (A), arriving at Wawaitin (B), arriving at the Spillway (C), and descending past Rocheleau, with significant p values (p< 0.05) shown in bold ...... 63

Table 2.5: The GLMM model outputs for lake sturgeon movements in the Mattagami River, Ontario reintroduced population. Movement upstream of the wintering hole (A), arriving at the spillway in the Mattagami (B), and descending passed the Rocheleau station (C) Mixed effects of water temperature (°C) and discharge (m3/s), fixed effect of year. Tables include mean values (mean), standard deviation (sd), 95% confidence intervals (2.5% and 97.5%), the scale reduction factor (Rhat), and effective sample size (Neff)...... 64 Table 2.6: Reported spawning temperatures for lake sturgeon ...... 65 Table 3.1: Genetic diversity measures for the reintroduced Mattagami lake sturgeon population (n=104) sampled between 2011-2016, and the source (Adam’s Creek) population (n=43) provided by the OMNRF, based on 12 microsatellite loci. Values indicate mean and standard error (SE) for each metric. Statistical significance for differences between the source and reintroduced populations was tested using Student’s t- test ...... 110

Table 3.2: Estimated effective number of breeders (Nb) for the 2007, 2008, and 2009 cohorts of the Mattagami, Ontario reintroduced population (including below Sandy Falls Dam) sampled between 2011-2016 using the linkage disequilibrium Ne estimator (LDNE) and polymorphism thresholds (minimum allele frequencies of 0.05, and 0.02, with 95% confidence intervals) based on 14 microsatellite loci ...... 111 Table 3.3: Parent-offspring pairs identified within the reintroduced Mattagami River population using CERVUS and COLONY analysis based on 15 microsatellite loci, showing tag numbers of adults from the original translocation effort ...... 112

List of Figures

Figure 1.1: The total length (cm) of translocated individuals (n=49) released in the Mattagami River, Ontario. Information on two individuals not recorded ...... 18 Figure 2.1: Map of the reintroduction site (boxed area) within the Mattagami River watershed (James Bay drainage basin, Ontario). The inset map shows the location of the study system in Ontario, Canada. The source population for the translocated adults is starred. Black boxes represent hydroelectric facilities ...... 66 Figure 2.2: Locations of stationary monitors (blue) and temperature loggers (red) for the Mattagami River sturgeon restoration project ...... 67 Figure 2.3: The size frequencies for radio-tagged lake sturgeon translocated into the Mattagami River in 2002, acclimated12 post-release individuals radio-tagged in 2002 above, and eight acclimated individuals radio-tagged in 2011 and 2013 below, the dotted lines represent the means for each sampling period ...... 68 Figure 2.4: Post-release (31 August 2002-25 July 2003; green) and acclimated (24 June 2011-16 June 2016; red) kernel density maps for radio-tagged lake sturgeon translocated into the Mattagami River, Ontario. Kernel density based on tracked locations ...... 69 Figure 2.5: Post-release (2002-2003) and acclimated (2011-2016) kernel density maps for radio-tagged lake sturgeon translocated into the Mattagami River, Ontario. Kernel densities mapped in purple based on the summer, fall, and winter locations during the A. summer, winter, and fall seasons of the post-release period (2002-2003), B. summer, winter, and fall seasons of the acclimated period (2011-2016), C. spring of the post- release period (2002-2003), and D. spring of the acclimated period (2011-2016) ...... 70 Figure 2.6: Spring water discharge for April to September in 2014, 2015, and 2016 in the Mattagami River, Ontario ...... 71 Figure 2.7: Spring water temperatures for April to August in 2014, 2015, and 2016 in the Mattagami River, Ontario ...... 72 Figure 2.8: Dates that adults left the wintering hole (LW), arrived at Wawaitin station (AW), arrived at Spillway (AS), and descended past the Rocheleau station (DR) in 2014, 2015, and 2016 ...... 73 Figure 2.9: Boxplots of the mean daily water discharge and mean daily water temperature for lake sturgeon arrival dates at different locations along the spawning movements in the Mattagami River, Ontario in 2014, 2015, and 2016 ...... 74 Figure 3.1: Map of the site where lake sturgeon were reintroduced in 2002 (boxed area) within the Mattagami River watershed (James Bay drainage basin, Ontario). The inset map shows the location of the study system in Ontario, Canada. The source population for the translocated adults is starred. Black squares represent hydroelectric facilities ....113

Figure 3.2: Length frequencies of lake sturgeon caught in each sampling year (2011 (n=32), 2013 (n=15), 2015 (n=87), and 2016 (n=5)) within the reintroduction segment of the Mattagami River. Dotted line represents sampling year mean total length (mm). The 2011 year includes 19 juveniles caught below Sandy Falls dam ...... 114 Figure 3.3: Age structure of juvenile lake sturgeon caught within and immediately below the reintroduction area (n=115), based on back calculation pectoral fin ray growth annuli. Bars show total numbers of captured aged juveniles across all sampling years (2011- 2016) ...... 115

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List of Appendices

Table A3.1: Reintroduced Mattagami, Ontario lake sturgeon catches by year ...... 139 Table A3.2: The 15 loci genotyped for lake sturgeon from the reintroduced Mattagami, Ontario population (n=141) sampled in 2011-2016, with associated genotyping error rates ...... 140 Table A3.3: Allele frequencies for the reintroduced, downstream, source, and Groundhog River, Ontario lake sturgeon populations ...... 141 Table A3.4: Estimated allele frequencies accounting for potential null alleles for the 15 loci genotyped for lake sturgeon from the reintroduced Mattagami, Ontario population (n=141) sampled in 2011-2016 ...... 143 Figure A3.1: Lake sturgeon pectoral fin rays from an adult from the translocated Mattagami River, Ontario population (left), and an adult from the nearby natural Groundhog River, Ontario population (right). Scale bar indicates 1000µm in each photograph ...... 145 Figure A3.2: A power analysis of the number of observed and expected parental pairs based on potential genotyping error and mismatches between parent and offspring genotypes. Based on genotyped (15 loci) adults (n=13), juveniles (n=124), and larvae (n=4) from the reintroduced Mattagami River population sampled from 2011-2016 .....146 Figure A3.3: Length-weight relationship of unique juvenile lake sturgeon caught within the reintroduction segment of the Mattagami River (n=126) ...... 147

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Chapter 1: General Introduction Freshwater ecosystems support 10% of all known species and provide necessary resources for supporting humans (Carrizo et al., 2013). Heavy reliance on fresh water has led to widespread degradation of freshwater ecosystems through land use changes, introduction of non-natives species, and water course alterations (Dudgeon et al., 2006).

These combined impacts have resulted in freshwater ecosystems being a conservation priority, as one of the most endangered ecosystem-types in the world (Dudgeon et al.,

2006). It is estimated that dams and hydroelectric facilities regulate 77% of large rivers in the northern hemisphere (Dynesius and Nilsson, 1994).

Globally, riverine ecosystems are severely impacted by water resource developments that alter natural flow regimes and fragment riverine systems (Rosenberg et al., 2000). Dams are considered a major threat to freshwater biodiversity (Vörösmarty et al., 2010). Dams often obstruct dispersal and migration of riverine species, which can negatively impact population persistence and species’ recolonization ability (Fagan,

2002; Geist, 2011). Habitats required for the completion of species’ life cycles may be separated or reduced by habitat alterations and fragmentation (Auer, 1996a; Joy and

Death, 2001; Cumming, 2004). This reduction in connectivity can increase extirpation risk due to population decline, recruitment failure, and genetic diversity loss, especially in species that migrate long distances and rely on natural flow conditions (Freeman et al.,

2003; Faulks et al., 2011; Braulik et al., 2014). Reintroduction can allow species to recolonize extirpated areas that are no longer accessible naturally due to dam fragmentation (Fagan, 2002; Geist, 2011; Cochran-Biederman et al., 2014).

A variety of techniques have been used to mitigate dam fragmentation effects and restore habitat connectivity (Noonan et al., 2012). Fish passage structures such as fishways and lifts are often designed for economically important species such as salmonids (Calles and Greenburg, 2005; Parsley et al., 2007), but have generally been less effective for many other species (Moser et al., 2002; Noonan et al., 2012) including sturgeons (Thiem et al., 2011). Although stocking can be used as a management tool, this can also incur genetic risks which may negatively impact persistence of remnant populations (Hindar et al., 1991; Ford et al., 2002; Araki et al., 2007). Rearing and releasing fish can also be expensive (Brown and Day, 2002), and hatchery-reared fish often have high mortality rates following release into the wild (Brown and Laland, 2001); but mortality may be reduced by using best hatchery and release practices (Chebanov et al., 2011). Alternatively translocations, a form of reintroduction, can be used.

Translocations involve human-mediated movement of organisms from one site to release in another (IUCN/SSC, 2013) with the goal of re-establishing species in formerly occupied areas that are no longer accessible without active human intervention (Welsh et al., 2010; Cochran-Biederman, et al., 2014). Although translocations are an increasingly popular conservation tool (Armstrong and Seddon, 2008; Ewen et al., 2012), success has been low (Griffith et al., 1989; Germano and Bishop, 2009) and the causes of failures are often unknown because follow-up monitoring is rarely conducted (Fischer and

Lindenmayer, 2000). Long-term success of reintroduced populations requires populations to sustain themselves, which means reintroduction biologists need to consider ecological, environmental, disease, demographic, and genetic aspects that will increase the likelihood of success (Keller et al., 2012). The use of translocation will likely rise as it is applied in

more diverse ways, particularly as a tool for adaptation to climate change (Sansilverstri et al., 2015), and other human-driven impacts (McGill et al., 2015). It is therefore critical to research and refine translocation efforts to maximize success.

Although there are no broadly accepted criteria for successful reintroductions

(Robert et al., 2015), successful reintroductions require both the establishment and multi- generational persistence of the species in the release area (Chauvenet et al., 2013).

Establishment is the survival and reproduction of released individuals (Seddon et al.,

2012), and persistence is the increase in population and density in the release area

(Armstrong and Seddon, 2008). Historically, conservation translocations have had low success and were generally inadequately planned and monitored (Armstrong and Seddon,

2008). In 1998, the IUCN published guidelines for reintroduction to increase success, and publications emphasized the need for increased planning and monitoring of reintroduction efforts (Griffith et al., 1989). Despite improvements, the success rate of translocations was still generally low, and literature published was often based on retrospective evaluation (Armstrong and Seddon, 2008). Reintroduction success depends on a suite of factors, including release group size, release site location, release area habitat, movements after release, and genetic makeup of released individuals (Griffith et al., 1989; Wolf et al., 1998; Armstrong and Seddon, 2008; Ewen et al., 2012). To maximize the success of future reintroductions, we must research the various factors that influence the establishment or persistence of the population.

Movement

In early translocation failures, released individuals often disappear from the release area due to mortality or dispersal (Le Gouar et al., 2012). Translocation is 3

essentially a forced dispersal event making it inherently stressful (Teixeira et al, 2007;

Dickens et al., 2009), which is likely a factor in the mortality and dispersal frequently observed soon after release (Armstrong and Seddon, 2008; Swaisgood, 2010). Although mortality is the most common cause for the disappearance and is a critical determinant of establishment success (Armstrong and Seddon, 2008), post-release dispersal also occurs

(Le Gouar et al., 2012). Due to frequently inadequate post-release monitoring, the fate of missing individuals is rarely known (Wolf et al., 1996); it is therefore important to account for post-release dispersal while making population projections for reintroduced populations (Armstrong and Reynolds, 2012) and distinguish between mortality and dispersal to maximize future establishment success (Tweed et al., 2003). Post-release dispersal can affect translocation success in various ways that differ between the three different phases of population development, those being (i) establishment, (ii) growth, and (iii) regulation (Sarrazin, 2007; IUCN/SSC, 2013).

The establishment phase occurs from the first release to when the post-release effects on the translocated population are no longer apparent (IUCN/SSC, 2013). During this time the population is vulnerable to failure due to post-release dispersal and exploratory movements (Jachowski et al., 2016). Dispersal out of the release area during the establishment phase has been reported in many species (Griffith et al., 1989; Kleiman,

1989; Short et al., 1992; Yalden, 1993; Miller et al., 1999), and can result in early translocation failure (Ewen et al. 2012). Individuals that disperse out of the release area do not contribute demographically or to the genetic pool of the translocated population

(Ewen et al., 2012). Often, post-release there is a period of high movement intensity and frequency as individuals explore the novel habitat (Rittenhouse et al., 2007; Hester et al.,

2008). Exploratory movements are costly and may reduce time allocated to important behaviours such as foraging or reproduction (Ewen et al., 2012). Information regarding post-release movements can inform future translocation efforts, maximizing success of translocated populations.

Survival and reproduction are often considered early indicators of reintroduction success (Jachowski et al., 2016), although these criteria do not always lead to a self- sustaining population (Wolf et al., 1998). Monitoring movements of translocated individuals confirm survival and breeding, which are often reported as immediate measures of translocation success (Goossens et al., 2005; Strum, 2005; King et al., 2012;

Cochran-Biederman et al., 2014). Identifying reproductive habitat is important for both confirming reproduction (King et al., 2012) and the ongoing protection of the reintroduced population, especially in heavily human-impacted freshwater systems

(Schindler, 1998; Dudgeon, 2000; Malmqvist and Rundle, 2002).

Another issue to be considered when planning translocations, especially for migratory species, is accounting for the variables that cue reproductive behavior and movements, as well as ensuring that suitable habitat is available to support the requirements of each life stage (Jachowski et al., 2016). Spawning migrations are common in the life history of many fish species (Gross, 1987), and migration timing typically coincides with environmental conditions that facilitate adult movement and juvenile survival (Hodgson and Quinn, 2002). Spawning migration and timing of reproduction are initiated by important behavioral decisions stimulated by reliable environmental cues (e.g., temperature) which are associated with successful reproduction in fishes (Williams and Nichols, 1984). Environmental variables in the release habitat 5

may differ from the source habitat which may impact reproductive success of the translocated population. Monitoring seasonal movement and environmental cues can give insight into whether fish are reproducing during ideal environmental conditions.

Demographics

The primary objective of most reintroductions is to establish a self-sustaining population (Griffith et al., 1989; Fischer and Lindenmayer et al., 2000), which is broadly defined as a population that is stable or increasing without additional releases (Converse and Armstrong, 2016). Ultimately, reintroduction success is determined by demography and recruitment. Post-release monitoring of abundance, survival, dispersal, and reproduction of released individuals gives insight into whether the population is self- sustaining (Converse and Armstrong, 2016). Confirmation of juvenile recruitment into the population can be acquired by sampling reintroduced populations generations after release. For certain species, non-lethal ageing techniques can be employed to approximate the timing of the first successful reproduction and the number of subsequent reproductions.

Genetics

Translocation efforts should consider genetic makeup for both the selection of the source population, which should be adapted to similar conditions as the release habitat

(Welsh et al., 2010), and the ongoing genetic diversity monitoring of the translocated population. Reintroductions generally release small numbers of individuals, making them susceptible to genetic concerns associated with small populations such as loss of genetic diversity through founder effects, genetic drift, and inbreeding (Frankham et al., 2002;

Keller and Walker, 2002; Frankham, 2005; Allendorf et al., 2013). Loss of genetic diversity and increased inbreeding in a population can lead to inbreeding depression and a reduced ability to adapt, resulting in increased extinction risk (Saccheri et al., 1998;

Frankham, 2005). As these problems may contribute to the reported low success rates of translocation efforts (Fischer and Lindenmayer, 2000; Thévenon and Couvet, 2002;

Frankham et al., 2010; Allendorf et al., 2013), it is important that translocated populations receive sufficient genetic representation and diversity from their source population. Genetic variation also has significant consequences for the long-term ability of the population to adapt to change (Frankham et al., 1999; Pierson et al., 2015).

Determining the genetic diversity captured in translocation efforts can be accomplished by quantifying the number of individuals translocated, the proportion that contributes to the next generation (founders), and the allele frequencies in their genomes (Groombridge et al., 2012). These parameters define the founding genetic diversity of a new population, which is then subjected to the effects of genetic drift, selection, and mutation, as well as potential migration and gene flow (Groombridge et al., 2012). As translocations typically involve small numbers of released individuals, they are predisposed to lose diversity at an accelerated rate; accordingly, it is expected that translocated populations will have less genetic diversity than their source as a result of founder effects and genetic drift

(Groombridge et al., 2012). Although limited studies have monitored translocated populations for changes in genetic composition compared to the source (Seddon et al.,

2007), smaller release populations tend to experience more inbreeding and genetic diversity loss (Jameison and Lacy, 2012). Molecular markers are now more accessible than in the past, making the genetic monitoring of translocated populations easier.

Collecting long-term genetic monitoring data is therefore advisable for understanding the genetic effects of translocations and maximizing the long-term survival of translocated populations (Groombridge et al., 2012; Pierson et al., 2015).

Lake Sturgeon

Lake sturgeon (Acipenser fulvescens) are a large and long-lived freshwater fish species native to North America, with a widespread distribution spanning the Mississippi,

Great Lakes, and Hudson Bay drainages (Harkness and Dymond, 1961; Scott and

Crossman, 1998). Lake sturgeon populations underwent dramatic declines due to commercial harvest and riverine habitat loss, but still occupy much of their historical range at a fraction of the abundance (Houston, 1987; Bruch et al., 2016). Protections resulted in fisheries closures in the mid-1900s, but many populations do not appear to be recovering (Harkness and Dymond, 1961; Auer, 1996a; Haxton et al., 2014). Other factors contributing to lake sturgeon declines and impediments to recovery include habitat alteration, fragmentation, and pollution (Rochard et al., 1990; Birstein, 1993;

Haxton et al., 2015). Lake sturgeon have a complex life history, diverse habitat requirements, and are migratory, making them particularly sensitive to the impacts associated with habitat fragmentation by dams (Auer, 1996b; Haxton and Findlay, 2008;

Haxton et al., 2015). Dams in rivers, and their resultant effects on habitat alteration, fragmentation, isolation, and loss, are considered the obstacles most frequently impeding sturgeon recovery (Rochard et al., 1990; Haxton et al., 2014). The Ontario recovery strategy for lake sturgeon recommends re-introducing populations to historically occupied areas as one of the major goals (Golder Associates Ltd., 2011).

In 2002, lake sturgeon were released in a 35 km isolated river fragment of the

Mattagami River, between Wawaitin and Sandy Falls generating stations (GS) where the species was believed to be extirpated in an attempt to restore a population. The effort translocated 51 wild adults (≥1 m; Figure 1.1) from the Little Long Generation Station head pond by Adam’s Creek (~300km downstream; Figure 2.1). Lake sturgeon were believed to be extirpated after 40 years without a reported sighting or catch record (C.

Hendry, Ontario Ministry of Natural Resources and Forestry (OMNRF), pers comm).

Extirpation was attributed to a suite of factors including habitat loss, water quality issues from 50 years of log drives, fragmentation, and overexploitation (C. Hendry, OMNRF, pers comm).

Lake sturgeon were translocated and released in the late summer (August 27-30,

2002). Prior to release the length and weight of individuals was recorded, individuals were tagged externally using floy tags and released at the Timmins boat launch. Thirteen of the 51 translocated adults underwent surgery to implant radio-telemetry tags.

Study Objectives

This study aimed to assess the success of the lake sturgeon translocation effort using telemetry, demography, and genetic tools. In Chapter 2, I aimed to document post- release dispersal, compare post-release and acclimated movements of translocated individuals, and associated spawning movements with environmental cues using telemetry data from radio-tagged adults in the post-release period (2002-2003) and acclimated period (2011-2016). In Chapter 3, the goal was to assess the post-release demographic and genetic composition of a reintroduced population several years after the translocation effort, and use juvenile demographic and genetic data to assess the 9

effectiveness of the re-establishment effort. As the Ontario Lake Sturgeon Recovery

Strategy recommends re-introducing populations to historically occupied areas as a major goal (Golder Associates Ltd., 2011), it is hoped that the insights gained from the combined results from this thesis will be used to help inform future translocation efforts.

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Figure 1.1: The total length (cm) of translocated individuals (n=49) released in the Mattagami River, Ontario in 2002. Information on two individuals not recorded.

Chapter 2: Post-release dispersal, and spawning movements of a translocated lake sturgeon (Acipenser fulvescens, Refinesque 1817) population in the Mattagami River, Ontario.

Abstract: The continuing decline of North American freshwater fishes due to habitat degradation, fragmentation, and loss has led to translocation increasingly being used to mitigate these declines. Tracking fish movements after translocation can be used to assess success and inform future translocation efforts by identifying spawning events and aid in the continued management of translocated populations. This chapter assessed movement of translocated adult lake sturgeon (Acipenser fulvescens) at two different time periods: immediately post release and a decade later. Prior to their translocation in 2002, 13 adult lake sturgeon were surgically implanted with radio telemetry tags and tracked for one year. In 2011 and 2013, eight additional adults were captured within the reintroduction site and similarly implanted with radio tags. During early post-release movements six of the 13 sturgeon tagged in 2002 left the release area downstream, spilling over Sandy Falls

Generating Station (GS). In spring 2014, tagged lake sturgeon migrated to a spillway at the upper extent of the release area, and spawning was confirmed by the capture of larvae. Tagged sturgeon used different areas of the river during the post-release time compared to a decade later. Water velocity was correlated with lake sturgeon upstream and downstream spawning movements. Research into translocation movements and post- release dispersal gives insight into behavioural responses to translocation, and can inform future efforts to improve future translocation outcomes.

Introduction North American freshwater fishes have declined substantially over the past century, with 39% of species considered imperilled (Jelks et al., 2008). Habitat alteration, degradation and species range restrictions are primary threats to imperilled freshwater fishes (Jelks et al., 2008), with dams often compounding their impacts for freshwater biodiversity (Vörösmarty et al., 2010). Among their other effects, dams fragment freshwater habitats, blocking natural migration routes of freshwater fishes (Fausch et al.,

2002), which can negatively impact persistence and recolonization potential (Fagan,

2002; Geist, 2011). An estimated 77% of the large rivers in the northern hemisphere are regulated by dams and hydroelectric facilities (Dynesius and Nilsson, 1994).

Translocation can be used to mitigate the effects of fragmentation through restoring species to parts of their historical range that are no longer naturally accessible

(IUCN/SSC, 2013). Wild translocations involve capturing individuals at a site and releasing them in previously occupied areas (IUCN/SSC, 2013). As many translocation efforts have had low success (Griffith et al., 1989; Ewen et al., 2012), research on the successes and failures is critical to understanding how to establish populations (Griffith et al., 1989; Wolf et al., 1998; Germano and Bishop, 2009; Ewen et al., 2012).

Translocation outcome depends on many factors, including whether the agent of the initial decline of the population has been addressed, release group size, and the quality of the habitat at the release site (Griffith et al., 1989; Wolf et al., 1998; Ewen et al., 2012).

Before translocation, managers should also consider whether the environmental variables cueing fish migration at the release site are different from the source site and how fish will respond at the release site (Jachowski et al. 2016).

Translocation failures occur when released individuals disappear from the release area. Mortality was the most common documented evidence of failure, although post- release dispersal from the release site has also been shown to contribute (Ewen et al.,

2012). Post-release dispersal has the potential to affect translocation success in various ways that may differ between the three different phases of population development

(establishment, growth, and regulation phase; Sarrazin, 2007; IUCN/SSC, 2013). The establishment phase is characterized by slow growth which is limited by the low number of breeding individuals (Jachowski et al., 2016). If a population establishes, it undergoes a growth phase facilitated by the higher number of breeding individuals and high resource availability (Jachowski et al., 2016). The regulation phase occurs when the population begins approaching the carrying capacity (Jachowski et al., 2016). Early dispersal from the release area can be particularly detrimental to translocation success, which may be indicative the habitat was not suitable (Stamps and Swaisgood, 2007;

Ewen et al., 2012), and has been reported for many species (Griffith et al., 1989;

Kleiman, 1989; Short et al., 1992; Yalden, 1993; Miller et al., 1999). Species may disappear from the release site if they exhibit homing behavior, which drives individuals migrate back to their source habitat (Linnell et al., 1997; Oliver et al., 1998; Germano and Bishop, 2009; Mireles et al., 2012; Hinderle et al., 2015). Also, translocations tend to release small numbers of individuals, making the populations susceptible to Allee effects: post-release dispersal further reduces numbers, amplifying the Allee effects and impacting reproduction (Deredec and Courchamp, 2007).

After release, individuals may undertake an exploratory phase with increased movement intensity and frequency (Rittenhouse et al., 2007; Hester et al., 2008).

Exploration of the new environment exposes individuals to risk and is costly, potentially reducing energy allocation to foraging or reproduction, and is associated with low survival and reproductive success (Kurzejenski and Root, 1988; Letty et al., 2000;

Moehrenschager and Macdonald, 2003). Although there is no clear definition of reintroduction success (Seddon, 1999; Fischer and Lindermayer, 2000), immediate success is often measured by confirming survival and reproduction (Goossens et al.,

2005; Strum, 2005; King et al., 2012; Cochran-Biederman et al., 2014). Identifying reproductive grounds is important for both confirming successful reproduction (King et al., 2012) and the ongoing protection of the reintroduced population, notably in riverine systems which are impacted extensively by human activity (Schindler, 1998; Dudgeon,

2000; Malmqvist and Rundle, 2002).

Spawning migrations are common in the reproductive life history of many fish species (Gross, 1987). Migration timing typically coincides with environmental conditions suitable for adult movement and offspring survival (Hodgson and Quinn,

2002). The initiation of spawning migrations and timing of reproduction are important behavioral decisions stimulated by reliable cues related to the environment (e.g. temperature), which are associated with successful reproduction in fishes (Williams and

Nichols, 1984). Exogenous variables that have been associated with fish migrations include river discharge (Keefer et al., 2009), the lunar cycle (Kuparinen et al., 2009), and river water temperature (Binder and McDonald, 2008). Water resource developments are responsible for unprecedented worldwide impacts to riverine ecosystems, predominately from alterations to the natural flow regime (Rosenberg et al., 2000). Environmental flow assessment (EFA) establishes the extent to which the natural flow regime can be altered

for water resource development while maintaining biological integrity to allow important functions such as spawning (Poff et al., 1997). Lack of habitat that supports all life stages or variables cueing spawning movement and behaviours may also contribute to translocation failure (Jachowski et al., 2016).

Lake sturgeon

The historical range of lake sturgeon (Acipenser fulvescens) spans the Mississippi,

Great Lakes, and Hudson Bay drainages (Harkness and Dymond, 1961; Scott and

Crossman, 1998). Historically, lake sturgeon populations have experienced detrimental declines from commercial harvest; however, their distribution in Canada has remained the same (Houston, 1987; Bruch et al., 2016). The closure of fisheries in the mid-1900s does not yet appear to have allowed populations to recover (Harkness and Dymond,

1961; Auer, 1996a; Haxton et al., 2014). Habitat alteration, fragmentation, and pollution have also contributed to historical lake sturgeon declines and still impact recovery

(Rochard et al., 1990; Birstein, 1993; Haxton et al., 2015).

Lake sturgeon require diverse habitat for different life stages and are known to be migratory, making them particularly sensitive to the impacts associated with dams (Auer,

1996b; Haxton and Findlay, 2008; Haxton et al., 2015). Lake sturgeon populations commonly migrate long distances from their residential areas in rivers or lakes to spawning habitats characterized by rapid water currents (generally > 0.6m/s) and clean, coarse substrates (Seyler, 1997a; Auer, 1999; Johnson et al,. 2006). Environmental variables associated with lake sturgeon spawning migrations include photoperiod, river discharge, water temperature, ice off time, and the lunar cycle (Rusak and Mosindy,

1997; Scott and Crossman, 1998; McKinley et al., 1998; Auer, 1999b; Forshythe et al., 23

2012). Water temperature and flow in particular have been associated with lake sturgeon movements in many systems, although the timing of migration in each system varied

(Rusak and Mosindy, 1997; McKinley et al., 1998; Scott and Crossman, 1998; Auer,

1999).

Sturgeon recovery is often impeded by dams in rivers (Rochard et al., 1990;

Haxton et al., 2014) whereas fragmentation can reduce the ability to naturally recolonize when extirpation has occurred. For instance, it is suggested 250-300km of unfragmented river is required to support a self-sustaining lake sturgeon population (Auer, 1996a), which can be difficult in large rivers that are generally highly fragmented by dams

(Dynesius and Nilsson, 1994). This has recently been disputed, however, albeit under exceptional circumstances (McDougall et al., 2017). In many rivers, lake sturgeon spawning occurs just downstream of impassible barriers or dams (Auer, 1996b; Seyler,

1997a; Friday, 2005; Haxton 2006). Small river fragments may be able to support lake sturgeon, but they must contain connected habitat necessary for all life stages (Golder et al., 2011; McDougall et al., 2017). The Ontario Lake Sturgeon Recovery Strategy recommends populations should be reintroduced into historically occupied areas (Golder

Associates Ltd., 2011). Assessing translocations in small fragmented reaches gives insight into whether translocation can be an effective method for restoring lake sturgeon in previously occupied, fragmented habitats.

Study system and objectives

The Mattagami River is a part of the James Bay drainage basin, located in northern Ontario, Canada. The river begins at Mattagami Lake and flows northward

443km to its confluence with the Missinaibi River which forms the Moose River (OWA,

2004; Figure 2.1). The Moose River Basin is a complex ecosystem that covers an area of

109,000 km2 in northeastern Ontario (OWA, 2004). Historically, the Moose River Basin was interconnected, providing abundant and diverse habitats that supported lake sturgeon at all life history stages (OMNR, 2008). Hydroelectric development in the late 1930s fragmented the rivers into multiple sections, impacting the sturgeon populations (OMNR

2008). The Mattagami River compliance monitoring program manages water control structures on the river, which includes 10 hydroelectric generating stations, seven dams, and one diversion (Mattagami River Water Management Plan, 2002).

The 35km section of the Mattagami River in this study receives flows from the

Mountjoy, Grassy, and Tatachikapika Rivers (Figure 2.2). Lake sturgeon were believed to be extirpated from this 35 km segment of the upper Mattagami River (between the

Wawaitin Generating Station (GS) and the Sandy Falls GS) given that 40 years have passed without a reported sighting or catch record (Charles Hendry, OMNRF, pers comm). The Wawaitin Dam (<10m) was originally constructed in 1912, and was replaced by a larger 15W generating station in 2010. The Sandy Falls Dam (<10m) was constructed in 1911, and was replaced by a larger 6W generating station in 2010. Both generating stations operate as run-of-river flows, meaning they have limited water storage and are subject to seasonal river flows. Sturgeon extirpation from this river segment was hypothesized to have been caused by any of a suite of factors including habitat loss, water

quality issues due to log driving, fragmentation, and overexploitation (Charles Hendry,

OMNRF, pers comm).

In 2002, in an attempt to restore a lake sturgeon population in a 35km section of the Mattagami River, between Wawaitin and Sandy Falls dams, 51 wild adults (≥1 m) including 13 radio-tagged individuals were translocated from the Little Long Head Pond above Adam Creek (~300km downstream; Figure 2.1). The objectives of this chapter were to 1) compare movement and outmigration of individuals radio-tagged in the period just after release (2002-2003) and the individuals recaptured and radio-tagged a decade after release (2011-2016) when individuals had acclimated to the site; 2) identify spawning areas used by the lake sturgeon within the reintroduction site; 3) ascertain the environmental variables associated with the spawning migration in the translocated

Mattagami River lake sturgeon population to give insight into whether hydroelectric operation affects spawning success; and 4) determine whether the lake sturgeon population successfully established in this small section of river.

Methods

Field sampling

Sampling within the 35km river section occurred in 2002, 2011, and 2013 using a combination of extra-large gillnets (multifilament; 2.13m depth; total length of 24.8m; stretched-mesh panel sizes: 204, 230, 255, and 306mm) to target adult lake sturgeon, large-mesh nets (monofilament; 1.8 m depth; total length of 24.8 m stretched-mesh panel sizes: 38, 51, 64, 76, 89, 102, 114, and 127mm; North American Gillnet (NAG)), and

72m trotlines with 18 baited hooks consisting of three sizes (14/0, 13/0, and 12/0 Mustad circle hooks) spaced at 2m intervals. Hooks were baited with white sucker (Catostomus commersonii) meat and given enough rope to rest on the bottom. All gear types were set overnight as lake sturgeon catch rates tend to be higher at this time due to increased nocturnal activity (Chiasson et al., 1997). The 2002 sampling targeted only adult sturgeon using the extra-large gillnets and trotlines. The 2011 sampling effort targeted both reintroduction reach (between Wawaitin and Sandy Falls GS), and the river reach immediately downstream (between Sandy Falls and Lower Sturgeon GS; Figure 2.1).

Sampling efforts in subsequent years focused on the reintroduction release site. All sampling was conducted in accordance with Trent University ACC Protocol 23918.

Individual fish were measured for total length (cm) and fork length (cm), weighed

(g), tagged with an external floy tag, and released. Individuals captured from 2011 onwards were also tagged with a Passive Integrated Transponder (PIT; Model:

BIO12.B.01) on the left side under the third dorsal scute. Also beginning in 2011, the first fin ray of the left pectoral fin was extracted for aging and a tissue sample was collected for genetic analysis. For individuals smaller than 30cm, a small genetic sample

(clip of tail) was taken, however they were not tagged or fin clipped for aging. A subset of captured individuals underwent surgery to implant radio-tags, including all captured adults and larger juveniles.

Radio telemetry Captured lake sturgeon were placed in a 190L tub (Rubbermaid Roughneck

Jumbo Storage Tote) and transported to the surgery site located in a cleared area near the wintering hole (Figure 2.2). River water was regularly added to the tub to ensure the

water stayed cool and aerated. Upon arrival at the surgery site, lake sturgeon were moved to submerged pens located out of direct sunlight. A subset of adults (n=13) from the translocation effort in 2002, and all adults (n=8) captured in the 2011 and 2013 sampling efforts, underwent surgery to be implanted with internal coded radio transmitters (Model:

MCFT2-3A; 46mm long and weighed 6.7g in the water, frequencies ranged from

148.000-149.000 MHz.) that had an estimated battery life of 1374 days (Lotek,

Newmarket, ON; Table 2.1).

Surgical procedures were used to implant the internal radio transmitters into the ventral-side of the lake sturgeon just to the left of the center of the stomach. Individual selection followed the “two percent weight rule” of transmitter mass with respect to individual weight to minimize any harm or unnatural behavior due to the transmitter

(Adams et al., 1998; Brown et al., 1999; Jepsen et al., 2002). Lake sturgeon were placed into another 190L tub filled with river water to which tricaine methanesulfonate (MS-

222) was added as an anesthetizing agent and pH neutralised to ~7 with baking soda

(Ackerman et al., 2009). Individuals remained in the anesthetizing tub until they lost the ability to orient themselves in the water; once individuals were anesthesized they were moved to the operating table. Operating equipment and radio transmitters were sterilized using isopropyl alcohol prior to use. A 4 to 6 cm incision was made in along the mid- ventral line of the fish, shallow enough to ensure it did not damage internal organs, to expose the lake sturgeon’s body cavity. The antenna was coiled and placed inside the body cavity to prevent irritation and infection that can be caused by a trailing antenna.

The incision was sutured together with four to five stitches using Ethicon PDS

(polydioxanone) 36mm taper point sutures (http://www.ethicon.com/). The area was

rinsed with betadine (http://www.betadine.com), and the fish was placed in a recovery tank until fully revived.

Tracking involved a combination of manual tracking from a boat using an

SRX_400A WX5G (Lotek Inc., Newmarket, Ontario) receiver, and stationary monitors

(Model: SRX-DL) from 2011 on. A three-element folding antenna with a 150MHz range

(Model: F150-3FB) was used for manual boat and helicopter tracking. Manual tracking was conducted weekly for the spring, summer, and fall in 2002, 2003, and 2011 to 2013.

Manual tracking was conducted daily during the 2014 spawning season and weekly for the remainder of the summer and fall. In the spring of 2015 and 2016, manual tracking was conducted when possible during the netting effort. Lake sturgeon were tracked based on signal strength; a handheld GPS was used to obtain UTM coordinates when signal reached near maximum strength (Power=255). During the 2014 spawning season, adult lake sturgeon entered the Wawaitin GS spillway; as this required manual tracking from land, triangulation was used to pinpoint and record estimated locations.

Stations

Six Lotek SRX-DL stationary monitors were set up at base stations within the release reach to monitor lake sturgeon movements in the Mattagami River and its tributaries (Figure 2.2). Base stations were placed along the shoreline in areas easily accessible by boat, and stationary monitors were set to record data every five seconds.

Upon setup, the location of each base station (UTM) was recorded. Stationary monitors were powered by a 12V marine battery that was charged by a solar panel (Model: BSP-

2012, Power Up, Baltimore, Maryland). One four-element folding antenna was attached to each stationary monitor and directed at the river. In the 2015 season, a stationary 29

monitor was moved into the spillway, where spawning had occurred the year before, which provided four stations upstream of the wintering hole to monitor the spawning migration. The capacity of base stations was 272,000 detections; data were therefore downloaded monthly to avoid overwriting, which never occurred.

The base stations often experienced electrical interference while operating, which required that a method be developed to distinguish between valid and invalid detections.

Interference was often equated to a 999 code; in some cases, however, codes associated with fish were logged as interference. To be considered a valid detection of a radio- tagged lake sturgeon, the stationary monitor must have picked up the fish code during at least two scans occurring within 15 minutes or less.

Environmental conditions To determine inter-annual differences in spawning movement timing, the data collected from the wintering “hole”, a deep water hole where the tagged sturgeon spent the winter, was used to assess when spawning movement started and ended. Individuals were considered to have started spawning movement when they were tracked to a location upstream of the wintering hole, either manually or by a station upstream. If individuals were tracked only by the Wawaitin station (Figure 2.2), they were not given a spawning movement start date for the year. The spawning movement end date was when individuals migrated down past the Rocheleau station (Figure 2.2), either being tracked by the station or manually. The duration of the spawning movement was determined by subtracting the downstream spawning movement date from the upstream spawning movement date.

Environmental conditions that might be associated with spawning movements were monitored during the study period. Water temperature was recorded once every half hour using HOBO TidbiT v2 temperature data-loggers placed in five locations within the

Mattagami River and its tributaries (Figure 2.2). Water discharge data were obtained from Environment Canada’s Mattagami River water surveying station 04LA002 near

Timmins (Water Survey of Canada, 2016). The Mattagami surveying station recorded water level and calculated water discharge based on the relationship between water level and discharge.

Larval driftnetting

In June 9th to 13th 2014 and June 2nd to 5th 2015, six larval drift nets were set in the spillway overnight. As Ontario Power Generation (OPG) supervision was required while working within the spillway of Wawaitin GS larval driftnetting periods were limited to five sampling days each year. Drift nets were stainless steel, D-shaped frames

(0.76 m across the base, 0.53 m high, with 3.6 m tapered, knotless, 700 µm nylon mesh bag). Driftnetting dates were determined based on degree days following the protocol by

Friday (2014). The D-frame nets were deployed using a triple point bridle attached to anchors and ropes to assure that the net rested on the bottom of the river. Contents of driftnets were removed each morning, placed in 20 mL scintilliation vials filled with 95% ethanol and kept for larval sturgeon identification.

Data Analysis

Comparison of post-release and acclimatization locations

Post-release (2002-2003) and acclimated (2011-2016) telemetry data were compared by estimating the kernel density for all tagged individuals in each time period.

Kernel density maps were generated for all seasons combined, spring, and the remaining seasons using R (R Development Core Team, 2009).

Environmental cues to spawning movements

Fish that were recorded in the vicinity of Wawaitin Dam during the spawning season (late May- mid June) were considered for the migration analysis (2014: n=7;

2015: n=6). Telemetry data revealed that the adult sturgeon spent each winter largely confined to a deep wintering hole (~10.8m at deepest area) located in the Mattagami

River (Figure 2.2). Spawning movement distances were measured from the wintering location using ArcMap; only upstream direction movements from the wintering hole were considered for analysis. Julian dates for lake sturgeon started moving upstream, reached the Spillway station, and started moving downstream were compiled for each year and analyzed using R. The first location was where the lake sturgeon was tracked or the first day after moving upstream of the wintering hole (wintering hole; location=0), the second location was where the lake sturgeon entered the Spillway for spawning (12.8km upstream of wintering hole; location=0.5), the third location was where the lake sturgeon passed by the Rocheleau's station or was downstream (9.80km upstream of wintering hole, location=1). One-way ANOVAs were run in R to identify inter-annual differences in spawning movement time. Student’s t-tests were performed to identify significant

differences in years, and a sequential Bonferroni correction was applied to account for multiple tests (Rice, 1989).

The mean daily water discharge and mean daily surface water temperature were calculated for each day during the spawning season. This analysis determined whether the environmental cues were significant drivers of spawning movements, and whether water discharge or temperature was a better cue to predict the timing of lake sturgeon spawning movements. Relationships between potential environmental cues and spawning movements were plotted as box plots.

Three mixed effects general linear models (GLMM) were used to determine if environmental variables were significant drivers of spawning movements at each location. The mixed effects models were conducted using OpenBUGS (Lunn et al., 2009) through an R interface and the package ‘R2OpenBUGS’. The Julian date of arrival at each location was the dependent variable and the independent variables were mean daily water discharge and temperature. The water discharge and temperature were mixed effects, with the year as a fixed effect. Year was not a significant fixed effect for movement upstream of the wintering hole and was excluded from the GLMM. The 2016 year was excluded from the analysis as only one radio-tagged sturgeon reached the

Spillway.

Results

In 2002, 51 wild adults were translocated, of which 13 were fitted with internal radio tags (data from one individual was not recorded; C. Hendry, OMNRF pers. comm.).

Total length for radio-tagged individuals ranged from ~80-140cm with a mean of

119.3cm (SD=16.0) (Figure 2.3); the mean weight was 10.2kg (SD=3.4), with a range of

2650 to 15700 g (Table 2.1). The eight individuals tagged in 2011 (n=2) and 2013 (n=6) were substantially larger (Table 2.1, Figure 2.3); the mean total length for radio-tagged adults was 139.7cm (SD=12.0) with a mean weight of 23.1kg (SD=8.9).

Post-release movements

Six of the thirteen (46%) individuals that were radio-tagged in 2002 migrated downstream over Sandy Falls GS and out of the release site within days (C. Hendry,

OMNRF pers. comm.). Radio-tagged individuals spent more time in the area above

Sandy Falls GS during the early post-release years (Figure 2.4 and Figure 2.5).

Comparison of post-release and acclimatization locations

Post-release and acclimated kernel maps differed (Figure 2.4). During the initial

(2002-2003) post-release period, individuals were tracked mainly in the lower half of the section of the Mattagami River, including the location just above Sandy Falls GS (Figure

2.5). Kernel maps indicated that post-release fish remained in similar locations throughout the year, including spring (Figure 2.5). By contrast, acclimated fish that were radio-tagged in 2011 and 2013 were mainly tracked in the upper half of the section of the

Mattagami River (Figure 2.5). Spring kernel maps for the latter differed from the rest of the year, with a high density of locations occurring near the Wawaitin GS (Figure 2.5).

Assessment of spawning movements and conditions

In spring 2014 radio-tagged adults moved upstream to Wawaitin GS and moved into the Spillway (Figure 2.5). Four larval sturgeon were captured in the Wawaitin GS 34

Spillway in spring 2014 (refer to Chapter 3). It was not possible to locate or track all radio-tagged adults at all spawning movement points each sampling year (Table 2.3), and radio-tagged adults showed inter-annual differences in timing for spawning movements.

A one-way ANOVA found significant differences in timing between each year of the study (2014, 2015, 2016) and the timing of arrival at each spawning movement location

(left wintering hole, arrival at Wawaitin station, arrival at Spillway, and descended past

Rocheleau; F(1,55)=37.7, p<0.001). There were significant differences in the times of arrival and the various locations along the spawning movement stops between the 2014 year and the 2015/2016 years (Table 2.4). The 2014 year had a later start to the spawning movement (133 Julian days ±4.4 days SD), a later arrival at Wawaitin station (140 Julian days ±1.4 days SD), a later arrival at the Spillway (146 Julian days ±1.2 days), and a later descent past Rocheleau station (149 Julian days ±1.2 days SD). On average, lake sturgeon spawning movements occurred 10 days later in 2014 compared to the date individuals left the wintering hole in 2015 (124 Julian days ±2.3 days SD) and 2016 (123 Julian days ±0 days SD), arrived at Wawaitin station in 2015 (128 Julian days ±0.5 days SD) and 2016

(126 Julian days ±2.0 days SD), arrived at the Spillway 2015 (135 Julian days ±1.4 days

SD) and 2016 (137 Julian days, sample size too small for standard deviation), and descended past Rocheleau station in 2015 (139 Julian days ±0.8 days SD). The 2016 year did not appear to have many spawning movements of radio-tagged individuals because only one individual reached the Spillway station (Table 2.3). Student t-tests indicated that the timing of arrival at points along the spawning movement were significantly later for the 2014 season when compared with the 2015 and 2016 seasons. There were no significant differences between the 2015 and 2016 years (Table 2.4). For the 2014 and

2015 years, the duration lake sturgeon spent in spawning movement and spawning was calculated and compared. There was no significant difference in the duration between the two years (t=0.144, df=6, p=0.89).

Environmental cues

Spring water temperature was comparable in all years (Table 2.2; Figure 2.7). The water discharge was substantially greater in May 2014 compared to the 2015 and 2016 years (Figure 2.6). The 2016 year was excluded from the analysis because only one radio- tagged individual reached the spawning site (Spillway; Figure 2.9). Results from the

GLMM analysis indicated that mean daily water discharge was associated with movement upstream of the wintering hole and downstream of the spawning location

(Table 2.5). Water discharge was not significantly associated with arrival at the spillway

(Table 2.5). Water temperature, interaction and year were not significantly associated with arrival at the spillway or descent (Table 2.5). The environmental cues, particularly water discharge, were quite variable (Figure 2.9). Lake sturgeon started spawning movement at a mean daily water temperature of 5.82°C (±0.78°C SE), reached the spawning site (Spillway) at a mean daily water temperature of 8.61°C (±0.95°C SE), and descended downstream past the Rocheleau station at a mean daily water temperature of

12.89°C (±0.33°C SE). Lake sturgeon started spawning movement at a mean daily water discharge of 194.35m3/s (±36.38m3/s SE), reached the spawning site at a mean daily water discharge of 172.55m3/s (±32.91m3/s SE), and descended downstream past the

Rocheleau at a mean daily water discharge of 126.53m3/s (±21.36m3/s SE).

Discussion The telemetry data, in combination with evidence of wild reproduction by translocated adult sturgeon, indicate that this experimental translocation effort was successful at re-establishing sturgeon in the reach. Radio-tagged adults migrated to spawning locations in multiple years and larval sturgeon were captured within the small reintroduction reach. Early post-release monitoring of individuals did reveal dispersal out of the reintroduction area, but all evidence after that indicates long-term persistence.

Acclimated radio-tagged individuals tracked within the reintroduction reach (2011-2016) used different areas than radio-tagged individuals tracked in the first year (2002-2003) after the release. The environmental cue of water discharge correlated with upstream and downstream spawning movements.

Post-release dispersal

The downstream emigration out of the release area by six of thirteen radio-tagged translocated adults during early post-release monitoring indicated that a substantial proportion of the translocated individuals may have departed from the release area shortly after their introduction. Dispersal during the establishment phase can be detrimental to translocation success, as efforts usually release small numbers of founders (Ewen et al.,

2012). Early dispersal has been observed in many translocated populations with varying implications (Griffith et al., 1989; Armstrong et al., 1999; Le Gouar et al., 2008; Kermink and Kelser, 2013). In this study, post-release dispersal was not enough to cause translocation failure. However, high dispersal off-site increases the probability the translocated population will suffer from Allee effects (Fagan, 1999; South and Kenward,

2001) and can have negative effects on long-term persistence (Armstrong et al., 1999), as

these individuals do not contribute to the demographics or gene pool of the translocated population (Ewen et al., 2012).

Early post-release movements sometimes are the result of homing behaviour, which drives translocated individuals to migrate to their source habitat (Linnell et al.,

1997; Dickens et al., 2009). Many sturgeon species, including lake sturgeon, show genetic structure between spawning populations (McQuown et al., 2002; DeHaan et al.,

2006; Welsh et al., 2008; Marracca et al., 2015) and have been observed returning to the same spawning locations each year (Auer, 1999). This study translocated individuals upstream of the source population, which allows for downstream movement back to the source population. This effort translocated a small number of individuals (n=51), which may potentially have been halved by post-release off-site dispersal, ultimately impacting the demographics and potentially the genetics of the translocated population (Ewen et al.,

2012; Jachowski et al., 2016).

Methods to mitigate post-release dispersal include ensuring suitable release area, release site selection, and timing of release (Jachowski et al., 2016). Rapid movement away from release areas after translocation has been reported for many species (Griffith et al., 1989; Kleiman, 1989; Short et al., 1992; Yalden, 1993; Miller et al., 1999) indicating that the individual may perceive the release area to be unsuitable (Stamps and

Swaisgood, 2007). Release areas should meet all the species’ biotic and abiotic requirements, provide habitat for all life stages, be adequate for seasonal needs, be large enough to meet conservation goals, have adequate connectivity to suitable habitat in fragmented systems, and be isolated from potential non-habitat sink areas for the population (IUCN/SCC, 2012). Early post-release dispersal is less likely if the release 38

area habitat is similar to the source habitat, as this will mean individuals are well adapted to survival (Stamps and Swaisgood, 2007). As this study moved lake sturgeon upstream within the Mattagami River, the released individuals were likely well-suited to the habitat. Persistence of some adult fish and successful spawning substantiate this conclusion.

Post-release dispersal may also be mitigated through release site selection (where translocated individuals are released within the release area). Species should be released near areas that can be used as refugia post-release (Bodinif et al., 2012; Jachowski et al.,

2012; Jachowski et al., 2016). In this instance, translocated fish were released at the

Timmins boat launch, downstream of the radio-tagged lake sturgeon’s most used foraging and overwintering habitat from 2011-2016. In the first year after release, the radio-tagged lake sturgeon predominantly used the bottom third of the 35km section of the Mattagami

River and may have been unaware of the suitable habitat upstream. Releasing translocated individuals at the uppermost part of the section of the Mattagami River may have reduced post-release dispersal as the lake sturgeon would be exposed to suitable habitat as they moved downstream.

The timing of release should optimize seasonal resource availability and account for the species’ life history (Steury and Murray, 2004). Timing releases before or after major life history events such as reproduction can aid with establishment and overall translocation success (Llody and Powlesland, 1994; Bloxham and Tongue, 1995; Sarrazin et al., 1996; Jachowski et al., 2016). For instance, birds tend to move less during breeding season; releasing translocated individuals at this time may reduce post-release dispersal

(Sarrazin et al., 1996; Le Gouar et al., 2008). In this system, translocated individuals were 39

released in the late summer. By comparison, other studies that have translocated lake sturgeon above a dam during the spring had lower instances of post-release dispersal

(McDougall et al., 2013; Olach, 2015). In a study by Koenings et al. (2015) adult sturgeon translocated upstream in the spring reproduced before dispersing downstream over the dam. Further research on the timing of release should be considered to minimize post-release dispersal and maximize genetic contribution to the translocated population.

Comparison of post-release and acclimatization locations

Individuals that remained within the reintroduction reach (2011-2016) used different areas than the individuals tracked post-release (2002-2003). Acclimated individuals used mainly the upper half of the release reach, above the release site. As reintroduced animals are transported from their natal habitat to novel environments in a completely unfamiliar landscape (Berger-Tal and Saltz, 2014), obtaining local information is critical for their survival (Frair et al., 2007). Exploratory movements allow the individual to construct a spatial representation of their new environment to allow for efficient resource utilization (Russel et al., 2010; Berger-Tal and Avgar, 2012), but this comes at a costs (e.g. energetic demands; Eliassen et al., 2007). As the animal becomes more familiar with its new environment, its behaviour changes, movements of reintroduced animals should be indicative of the establishment process (Berger-Tal and

Saltz, 2014). The acclimated individuals spend much of their time in and around the wintering hole, indicating they perceive this area as valuable (MacArthur and Pianka,

1966; Bar-David et al., 2009). Other lake sturgeon populations spent their winter in deep holes that were a fraction of available habitats (Thayer et al., 2017). This indicates that they have passed the exploratory phase and are familiar enough to exploit resources in the 40

reintroduction reach. The heavy use of the wintering hole indicates it is important feature for lake sturgeon in the reach. This unique finding is knowledge that could aid in the selection of potential reaches for future reintroductions, as reaches with deep water habitats are likely more suitable for lake sturgeon reintroduction efforts.

Identification of spawning grounds

The Wawaitin GS Spillway was identified as a spawning location for the radio- tagged adults in the translocated lake sturgeon population, and was confirmed in 2014 by collecting larvae. The main initial objectives of translocations are the survival of translocated individuals, the settlement of individuals into the release area, and successful reproduction in the release area (Festa-Bianchet and Apollonio, 2003; Letty et al., 2002).

Identifying critical habitat such as spawning areas is important for the ongoing management and protection of species (McKinley et al., 1998;Species at Risk Act, 2002

Rosenfield and Hatfield, 2006) and the provincial Lake Sturgeon Recovery Strategy suggests maintaining habitat that supports sturgeon as one of the six major goals (Golder

Associates Ltd., 2011).

No radio-tagged lake sturgeon moved upstream towards the spawning area during early post-release telemetry, indicating that spawning may not have occurred in 2003.

Just after release individuals may have high movement intensity during the exploratory phase (Rittenhouse et al., 2007; Hester et al., 2008). Energy may be allocated towards exploratory movements rather than reproduction; this phase is generally associated with low survival and reproductive success (Kurzejenki and Root, 1988; Letty et al., 2000;

Moehrenschlager and Macdonald, 2003). Additionally, translocation causes stress, which may cause individuals to divert energy for normal biological functions (e.g. reproduction) 41

to cope with the stressor (Teixeira et al., 2007). The apparent lack of spawning in 2003 may be the result of individuals allocating energy to exploratory movements or reduced reproductive viability resulting from the stress of translocation. However, juvenile catches in 2011-2016 (see Chapter 3) indicate reproductive failure was not an ongoing problem for the reintroduced population.

Spawning movement patterns

Radio-tagged lake sturgeon showed strong site fidelity to the overwintering hole.

This site fidelity to deep pools in riverine habitats has been observed in other lake sturgeon populations (Borkenholder et al., 2002; Knights et al., 2002; Haxton et al., 2003;

Trested et al., 2011; Ecclestone et al., 2013; Thayer et al., 2017) with other riverine systems having high densities of lake sturgeon juveniles at >13.7m depth (Barth et al.,

2009). This study showed strong fidelity and capture rates in three deep water pools (see

Chapter 3), with most radio-tagged adults spending summer, late fall, and winter within the wintering hole.

The spawning movements of the translocated Mattagami population varied across years, specifically in 2014, which was characterized by high water velocity and low water temperature. Consistent with other lake sturgeon studies where unusual conditions (e.g. early onset spring) resulted in different spawning times, in these cases early onset spring resulted in early spawning movement and spawning times (Rusak and Mosindy, 1997;

Eccelstone, 2011). There was no difference in the spawning duration between years indicating that the lake sturgeon shifted the timing of spawning movement rather than the total amount of time spent in spawning movement and spawning. In this study, 2014 had high flows and low water temperature and resulted in later spawning movement and 42

spawning dates. Conversely, the 2016 year had low water velocity and a majority of radio-tagged individuals did not enter the spillway. Lake sturgeon have long intervals between spawning, with females generally spawning once every 4-9 years and males spawning once every 1-3 years (Roussow, 1957; Harkness and Dymond, 1961). Lake sturgeon’s large body size and high fecundity improves spawning success in years with suitable conditions and reduces the need to spawn in years with unsuitable conditions

(Beamesderfer and Farr, 1997).

Water temperature and flow have been associated with lake sturgeon movements in many systems (Rusak and Mosindy, 1997; McKinley et al., 1998; Auer, 1999; Bruch and Binkowski, 2002). Although movements upstream of the wintering hole and downstream post-spawning movements within the reintroduction site were associated with water velocity, no pattern was evident between movements and water temperature.

The lack of association of water temperature with movements may have been a result of the small sample size (eight radio-tagged adults). Although ice-off dates were not recorded in this study, ice conditions have also correlated with spawning time in other studies (Rusak and Mosindy, 1997; Eccelstone, 2011). Rusak and Mosindy (1997) found movement of Kettle River lake sturgeon correlated highly with water velocity and less so with temperature. The translocated population started spawning around 8°C, comparable to other lake sturgeon populations, especially the nearby Groundhog River population, indicating water temperature is a key environmental signal (Bruch and Binkowski, 2002;

Table 2.6). Lake sturgeon responding and reproducing at ideal water temperatures increases fitness, as reproduction at water temperatures above 20°C causes total egg mortality (Wang et al., 1985). However, the sample size was small and results should not

be over-interpreted. Additionally, during spring conditions water temperature and velocity increase simultaneously; parsing out which environmental cue has more effect on spawning movements can be problematic and given their synchrony may not be necessary.

Spawning movements are critical to long-term survival and reproduction. It is important to consider the extent to which migratory behaviour is innate or learned and the environmental cues that drive spawning movement before translocating a species

(Jachowski et al., 2016). For instance, Neilson et al. (2001) observed different strains of reintroduced Atlantic salmon (Salmo salar L.) exhibited a 15-19 day difference in migratory timing. In the Great Lakes, studies have documented two lake sturgeon spawning events during the spring at a single location (Auer and Baker, 2002; Bruch and

Binkowski, 2002).

Translocation success

Translocated individuals remained within the reintroduction reach and reproduced, indicating lake sturgeon can be successful within a small reach, providing that suitable habitat exists for all life stages. Lake sturgeon are known to move extensively between habitats to fulfill life history requirements (Golder Associates Ltd.,

2011). Non-fragmented habitat is considered critical to lake sturgeon survival with adults in many populations migrating over 130 km to spawn (Lyons and Kempinger, 1992;

Fortin et al., 1993). It has been estimated that 250-300 km of unimpounded river is necessary to support a self-sustaining lake sturgeon population (Auer, 1996a), and sturgeons typically do poorly in fragmented rivers (Limburg and Waldman, 2009).

Although this translocation effort released individuals into a small reach it appears to 44

have been successful in both reproduction and juvenile recruitment (see Chapter 3). Other studies have reported lake sturgeon populations thriving in isolated river segments (49 km; Barth et al. 2011), leading to an alternative hypothesis that in some rivers as little as

10 km is required to support a self-sustaining population (McDougall et al., 2017). Small river reaches may therefore be viable for supporting lake sturgeon populations if there is suitable and sufficient habitat present to support all life stages, including reproduction, feeding and growth, refugia and movement, and all habitats are accessible through unobstructed corridors (Golder Associates Ltd., 2011). Although this study did not assess habitat directly, spawning was confirmed within the release reach, in the Wawaitin GS spillway with habitat characterized by high flows, shallow depth, and coarse substrate associated with spawning habitat in other lake sturgeon populations (Auer, 1996b; Lane et al., 1996; McKinley et al., 1998; Peterson et al., 2007). Habitat requirements of early developmental life stages of lake sturgeon is largely unknown (Golder Associates Ltd.,

2011); young-of-year lake sturgeon have been caught in a variety of substrates and depths

(Kempinger, 1996; Peake, 1999; Holtgren and Auer, 2004; Golder Associates Ltd.,

2011). The recruitment of juveniles into the translocated population suggests that the habitat necessary for early life stages is present within the reach. Lake sturgeon habitat use changes seasonally, with individuals often occupying deep pools in periods of high flow (Seyler, 1997b). In the Winnipeg River, juvenile lake sturgeon preferred deep water habitats (Barth et al., 2009). The bathymetry data (provided by C. Hendry, OMNRF) revealed three deep pools within the reintroduction reach: the telemetry data indicated that both juvenile and adult lake sturgeon preferentially used these deep holes especially in the winter and peak summer season, suggesting these deep habitats are used during

periods of high flows or sub-optimal temperatures. This may also be a result of the manual tracking which occurred in the daytime, with lake sturgeon being primarily nocturnal. This stresses the importance of confirming deep water habitat in reaches being considered for future reintroductions. This lake sturgeon translocation into a small reach was successful, indicating this could be a useful tool for reintroducing lake sturgeon in other areas where they have been extirpated providing that the release reach contains necessary habitat.

Conclusions

Monitoring movements after a translocation event can provide important information for the ongoing management of the translocated population and can inform future translocation events. High rates of post-release dispersal can lead to translocation failure. Making informed decisions regarding species’ biology, release area, release site, and release time can reduce post-release dispersal. The translocated population may establish with moderate rates of post-release dispersal, but this may affect the genetics of the population and may decrease the probability of long-term persistence. The identification of reproductive areas is important to confirm the initial success of the translocation effort as well as ongoing protection and management of the population.

Spawning movement behaviours should be considered prior to translocating and when picking a source population. Lake sturgeon reintroduction into a small reach can be successful if the reach contains all habitat necessary for different life stages. As ecosystems continue to be degraded and fragmented globally, translocation may become increasingly important as a mitigation tool. Ongoing research into the movement of translocated species could inform future efforts and increase translocation success. 46

Acknowledgements Funding was provided by Ontario Power Generation (OPG), Lake Shore Gold, and Glencore Canada Corporation, Kidd Operations, the OMNRF, the Ministry of

Environment (MOE) Ontario Community Environment Fund, and managed by the

Wintergreen Fund for Conservation. Field sampling was conducted by Derrick Romain,

Charlie Hendry, Adam Rupnik, Emma Harten, Tarryn Adams, and Matt Werner

(OMNRF), along with countless volunteers from the Timmins Fur Council, Club

Navigateur-La Ronde, Glencore Canada Corporation, Kidd Operations, Lake Shore Gold, and OMNRF. Laurent Robichaud provided invaluable guidance, knowledge, and assistance tracking radio-tagged sturgeon. Tim Haxton provided substantial support with suggestions, guidance, and help troubleshooting the telemetry analysis.

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Fork Total Transm Weight Year Date External Tag Sex Length Length Age # (g) (cm) (cm) Hydro Orange- 49.443 2002 U 99.4 110 27-Aug-02 2283 9100 N/A 49.449 2002 27-Aug-02 Not Recorded U 0 N/A OPG 2002- 49.436 2002 U 71.4 81.7 27-Aug-02 1450 2650 N/A Hydro Red- 49.456 2002 F 106.9 117.9 27-Aug-02 1080 10250 N/A Hydro Orange- 49.469 2002 U 104.5 115.3 27-Aug-02 1615 10000 N/A OPG 2002- 49.463 2002 U 102.2 114.7 27-Aug-02 1226 9500 N/A OPG 2002- 49.483 2002 U 103.5 116 27-Aug-02 1445 7800 N/A Hydro Red- 49.516 2002 U 131.5 141.5 27-Aug-02 0502 15250 N/A MNR Red- 49.51 2002 U 129.5 140.1 28-Aug-02 02630 12600 N/A OPG 2002- 49.429 2002 1466 (OPG U 101.8 112.5 28-Aug-02 2002, yellow) 9400 N/A OPG 2002- 49.477 2002 U 107.5 121.5 28-Aug-02 1464 9800 N/A Hydro Red- 49.49 2002 F 119.6 134.8 28-Aug-02 0766 15700 N/A OPG 2002- 49.496 2002 U 114.5 125.5 28-Aug-02 1474 10000 N/A OPG 2002- 23 2011 U 128 142 19000 23 6-Sept-11 1446 OPG 2002- 24 2011 U 117 128 13000 27 6-Sept-11 1448 Orange 11-Sept-13 OMNR-11278; 27 2013 F 129 140.5 24050 35 Hydro Red 2610 11-Sept-13 Orange 28 2013 F? 130 145.5 24550 39 OMNR- 11279 Orange 29 2013 F? 143 155 38550 36 11-Sept-13 OMNR-11280; 60

OPG 2002- 1457 6- Sept-13 Orange 31 2013 U 106 117 10750 17 OMNR-11297 6-Sept-13 Orange 32 2013 U 128 140.5 27300 29 OMNR- 11291 Orange 11-Sept-13 OMNR-11281; 36 2013 U 136 149 27550 34 OPG 2002- 1455

Table 2.2: Annual comparison of spring water temperature (°C) and discharge (m3/s) within the Mattagami River, Ontario for 2014, 2015, and 2016.

Date May mean June mean April May mean June mean when water water mean monthly monthly 20°C was temperature temperature monthly discharge discharge reached (°C) (°C) discharge (m3/s) (m3/s) (m3/s) 2014 June 9th 7.6 18.4 48 277 11 2015 June 10th 11.0 17.4 59 108 104 2016 June 14th 11.5 17.0 107 105 55

Table 2.3: The number of radio-tagged lake sturgeon tracked at each location along the spawning movements in the reintroduced Mattagami River, Ontario population for 2014, 2015, and 2016.

Year Left Wintering Arrived at Arrived at Decended Past Hole Wawaitin Spillway Rocheleau 2014 7 6 7 3 2015 5 6 6 6 2016 1 6 1 0

A. Left Wintering Hole

2014 2015 2016 2014 0.002 0.002 2015 0.002 0.623 2016 0.002 0.623

B. Arrival at Wawaitin

2014 2015 2016 2014 <0.001 <0.001 2015 <0.001 0.192 2016 <0.001 0.192

C. Arrival at Spillway

2014 2015 2016 2014 <0.001 <0.001 2015 <0.001 0.310 2016 <0.001 0.310

D. Descended past Rocheleau

2014 2015 2016 2014 <0.001 NA 2015 <0.001 NA 2016 NA NA

mean sd 2.5% 97.5% Rhat Neff b0 -5.62 1.49 -8.93 -5.389 1.00 2100 Temperature (°C) -0.26 0.36 -1.10 0.26 1.00 1900 Discharge (m3/s) 0.03 0.01 0.01 0.06 1.00 15000 Deviance 15.06 4.09 8.57 24.63 1.00 10000

B. Arrive at Spillway

mean sd 2.5% 97.5% Rhat Neff b0 6.28 33.81 -55.07 74.69 1.03 88 Temperature (°C) 0.04 0.034 -0.53 0.80 1.02 420 Discharge (m3/s) 0.14 0.06 0.00 0.25 1.00 1800 Interaction 0.02 0.01 0.00 0.04 1.00 1500 Year -0.01 0.02 -0.04 0.02 1.03 87 Deviance 16.97 4.55 13.60 28.39 1.00 4300

C. Descended past Rocheleau

mean sd 2.5% 97.5% Rhat Neff b0 16.89 34.48 -40.88 89.56 1.49 8 Temperature (°C) 0.00 0.25 -0.51 0.52 1.01 430 Discharge (m3/s) 0.08 0.02 0.05 0.11 1.01 720 Interaction 0.01 0.00 0.00 0.01 1.01 600 Year -0.01 0.02 -0.05 0.01 1.49 8 Deviance 4.34 4.73 -1.46 16.43 1.00 8400

Table 2.6: Reported spawning temperatures for lake sturgeon. Population Reported Spawning Citation Temperature Range (°C) Groundhog River, Ontario, 10.0-11.0 Seyler (1997a) Canada Kaministiqua River, Ontario, 13.4-17.2 Friday (2004; 2005; Canada 2006a) Gull River, Lake Nipigon, 13.0-18.8 Harkness and Dymond Ontario, Canada (1961) Ottawa River, Ontario, Canada 16-20 Haxton (2008) St. Clair River, Ontario, Canada 12.0-13.0 Thomas and Haas (2004) Des Prairies River, Quebec, 11.6-15.4 LaHaye et al. (1992) Canada L`Assomption River, Quebec, 11.0-21.5 LaHaye et al. (1992) Canada Sturgeon River, Michigan, USA 9.5-19.0 Auer (1996) Wolf and upper Fox rivers, 8.8-21.1 Bruch and Binkowski Wisconsin, USA (2000)

Figure 2.1: Map of the reintroduction site (boxed area) within the Mattagami River watershed (James Bay drainage basin, Ontario). The inset map shows the location of the study system in Ontario, Canada. The source population for the translocated adults is starred. Black boxes represent hydroelectric facilities. 66

² Sandy Falls Dam

Timmins

Mountjoy Station

Wintering Hole Potvin Chutes Station Tatachikapika Grassy Station Station Spillway Station Wawaitin Station

00.5 1 2 3 4 Kilometers

Figure 2.2: Locations of stationary monitors (blue) and temperature loggers (red) for the Mattagami River sturgeon restoration project.

Esri, HERE, DeLorme, MapmyIndia, © OpenStreetMap 67 contributors, and the GIS user community

Figure 2.3: The size frequencies for radio-tagged lake sturgeon translocated into the Mattagami River in 2002, acclimated12 post-release individuals radio-tagged in 2002 above, and eight acclimated individuals radio-tagged in 2011 and 2013 below, the dotted lines represent the means for each sampling period.

Figure 2.4: Post-release (31 August 2002-25 July 2003; green) and acclimated (24 June 2011-16 June 2016; red) kernel density maps for radio-tagged lake sturgeon translocated into the Mattagami River, Ontario. Kernel density based on tracked locations.

Figure 2.5: Post-release (2002-2003) and acclimated (2011-2016) kernel density maps for radio-tagged lake sturgeon translocated into the Mattagami River, Ontario. Kernel densities mapped in purple based on the summer, fall, and winter locations during the A. summer, winter, and fall seasons of the post-release period (2002-2003), B. summer, winter, and fall seasons of the acclimated period (2011-2016), C. spring of the post- release period (2002-2003), and D. spring of the acclimated period (2011-2016).

Figure 2.6: Spring water discharge for April to September in 2014, 2015, and 2016 in the Mattagami River, Ontario.

Figure 2.7: Spring water temperatures for April to August in 2014, 2015, and 2016 in the Mattagami River, Ontario.

Figure 2.8: Dates that adults left the wintering hole (LW), arrived at Wawaitin station (AW), arrived at Spillway (AS), and descended past the Rocheleau station (DR) in 2014, 2015, and 2016.

Figure 2.9: Boxplots of the mean daily water discharge and mean daily water temperature for lake sturgeon arrival dates at different locations along the spawning movements in the Mattagami River, Ontario in 2014, 2015, and 2016.

Chapter 3: Translocation as a mitigation tool: demographic and genetic analysis of a reintroduced lake sturgeon (Acipenser fulvescens Rafinesque 1817) population

Abstract: This study assessed the establishment success of a translocation of adult lake sturgeon upstream of a hydroelectric dam in northern Ontario, Canada, using demographic and genetic data from juveniles and adults. The objectives of this study were to (1) assess the size and demographic structure of the reintroduced population; (2) determine if juveniles are present; (3) assess the genetic diversity of the reintroduced and source populations; and (4) determine whether translocated adults are related to juveniles within the population. Gillnet and trotline sampling in multiple years (2002-2003; 2011-

2016) resulted in the capture of many juveniles (n=126) and some adults (n=13) at the release site and downstream. The first fin ray of the left pectoral fin was collected for ageing and genetic analysis, and individuals were genotyped at 15 microsatellite loci.

Age interpretations from the juvenile samples showed consistent cohorts starting in 2007

(2006-2012). Successful reproduction and recruitment by translocated adults was confirmed through genetic parentage analysis of microsatellite data which linked juveniles to parents that had retained tags from the original translocation. Based on the microsatellite data, the genetic diversity of the reintroduced population (HO) was comparable to its source (HO = 0.57 ±0.07 and 0.53 ±0.06 respectively), although its estimated effective population size (Ne) was lower (Mattagami = 20.4 (13.5-30.5);

Adam’s Creek = ∞ (72.0-∞)). These results suggest that the experimental translocation of

* This chapter is in press: Boothroyd, M.; Whillans, T.; Wilson, C., 2017: Translocation as a mitigation tool: demographic and genetic analysis of a reintroduced lake sturgeon (Acipenser fulvescens Rafinesque 1817) population. J. Appl. Ichthyol. (in press). 75

wild adult lake sturgeon was successful, and highlight the value of treating translocation efforts as experimental reintroductions.

Introduction

Human construction of dams has led to substantial alteration and fragmentation of aquatic habitats (Minckley, 1995; Sala et al., 2000; Jackson et al., 2001; Nilsson et al.,

2005; Dudgeon et al., 2006). By fragmenting lotic habitats, dams obstruct the migration and dispersal of resident and migratory species, which can reduce population persistence and recolonization potential (Fagan, 2002; Geist, 2011). This may also hamper the completion of species’ life cycles (Auer, 1996a; Joy and Death, 2001; Cumming, 2004), resulting in reduced genetic diversity and increased extirpations, particularly for migratory species that rely on natural flow conditions and extensive migration routes

(Freeman et al., 2003; Faulks et al. 2011; Braulik et al., 2014). As dams also limit species’ ability to naturally recolonize extirpated areas (Fagan, 2002; Geist, 2011), reintroduction is an effective option for restoring species presence in formerly occupied upstream habitats (Cochran-Biederman et al., 2014).

Translocation is a form of reintroduction, which involves human-mediated movement of organisms from one site to release in another (IUCN/SSC, 2013) with the goal of re-establishing species in formerly occupied areas that are no longer accessible without active human intervention (Welsh et al., 2010; Cochran-Biederman, 2014).

Although translocations are an increasingly used conservation tool (Armstrong and

Seddon, 2008; Ewen et al., 2012), success rates have been low (Griffith et al., 1989;

Germano and Bishop, 2009) and the causes of failures are often unknown because follow-up monitoring is rarely conducted (Fischer and Lindenmayer, 2000).

The lake sturgeon (Acipenser fulvescens) is one of the largest and longest-lived freshwater fish in North America (Scott and Crossman, 1998), with a widespread 77

distribution of North American sturgeons, spanning the Mississippi, Great Lakes, and

Hudson Bay drainages (Harkness and Dymond, 1961; Scott and Crossman, 1998). Lake sturgeon still occupy much of their historical distribution in Canada, although many populations underwent dramatic declines as a result of commercial harvest and are currently present at a fraction of their original abundance (Houston, 1987; Bruch et al.,

2016). Despite fisheries closures in the mid-1900s, many lake sturgeon populations do not appear to be recovering (Harkness and Dymond, 1961; Auer, 1996a; Haxton et al.,

2014). In addition to exploitation, other factors which contributed to lake sturgeon declines and are still impediments to their recovery include habitat alteration, fragmentation, and pollution (Rochard et al., 1990; Birstein, 1993; Haxton et al., 2015).

The complex life history, diverse habitat requirements, and migratory nature of lake sturgeon make them particularly sensitive to the impacts associated with habitat fragmentation by dams (Auer, 1996b; Haxton and Findlay, 2008; Haxton et al., 2015), and dams in rivers are considered the obstacles most frequently impeding sturgeon recovery (Rochard et al., 1990; Haxton et al., 2014). In Ontario, Canada, the recovery strategy for lake sturgeon recommends re-introducing populations to historically occupied areas as one of the major goals (Golder Associates Ltd., 2011).

In 2002, 51 wild lake sturgeon were released in an effort to restore the species in an isolated river fragment where it was believed to be extirpated. This study assessed the effectiveness of the 2002 transfer for re-establishing a self-sustaining population. More specifically, this study assessed the post-release demographic and genetic composition of a reintroduced population several years after an adult translocation effort, and used juvenile demographic and genetic data to assess the effectiveness of the re-establishment

effort. The objectives of this study were to 1) assess the size and demographic structure by targeted sampling of the reintroduced population; 2) determine if juveniles were present (cohorts and year class contributions); 3) assess the genetic diversity of the reintroduced and source populations; and 4) determine whether juveniles in the population were offspring of translocated adults using genetic parentage analysis.

Methods

Study Site

The Mattagami River is located in northern Ontario, Canada and is a part of the

James Bay drainage basin (Figure 3.1). The river begins at Mattagami Lake and flows

443km to its confluence with the Missinaibi River which then forms the Moose River

(Figure 3.1). The Mattagami River compliance monitoring program manages water control structures on the river, which includes 10 hydroelectric generating stations, 7 dams, and one diversion (Mattagami River Water Management Plan, 2002).

Lake sturgeon were believed to be extirpated from a 35km segment of the upper

Mattagami River, between the Wawaitin Generating Station (GS) and the Sandy Falls GS

(Figure 3.1), after 40 years without a reported sighting or catch record (C. Hendry,

Ontario Ministry of Natural Resources and Forestry (OMNRF), pers comm). Extirpation was attributed to a suite of factors including habitat loss, water quality issues, fragmentation, and overfishing (C. Hendry, OMNRF, pers comm). In 2002, in an experimental attempt to restore lake sturgeon to this stretch of the Mattagami River, 51 adults (individuals ≥1 m) were translocated from the Adam’s Creek population located

below the confluence of the Mattagami, Kapuskasing, and Groundhog Rivers to the 35km section of river between Wawaitin and Sandy Falls generating stations (Figure 3.1). All translocated individuals were marked with external Floy tags, and radio-telemetry transmitters were implanted within 13 of the translocated adults (C. Hendry, OMNRF, pers. comm.), 6 of which emigrated downstream over the Sandy Falls generating station shortly after their reintroduction (see Chapter 2). Since the release, there have been catch reports of juvenile lake sturgeon attributed to the reintroduction effort (OMNRF, unpubl. data), but no quantitative studies were conducted to assess the effort’s success.

Field Sampling

Sturgeon were targeted in the reintroduction area in 2011, 2013, 2015, and 2016 using a variety of sampling gears. Netting efforts used a combination of extra-large gillnets (multifilament; 2.13 m depth; total length of 24.8 m; stretched-mesh panel sizes:

204, 230, 255, and 306 mm) to target adult lake sturgeon, large-mesh nets

(monofilament; 1.8 m depth; total length of 24.8 m stretched-mesh panel sizes: 38, 51,

64, 76, 89, 102, 114, and 127 mm; North American Gillnet (NAG); Lester et al. 2009) to target juvenile lake sturgeon (Haxton et al. 2014), and 72 m trotlines with 18 hooks baited with white sucker (Catostomus commersonii) consisting of three sizes (14/0, 13/0, and 12/0 Mustad circle hooks) spaced at 2 m intervals. All gear types were set overnight as lake sturgeon catch rates tend to be higher at this time due to increased nocturnal activity (Chiasson et al., 1997). Additional samples including fin clips for genetic analysis were provided by incidental catches by local anglers, which increased our sample size.

Samples from the source population (Adam’s Creek; n=43), and an undisturbed population (Groundhog River; n=54) were provided for comparison by the OMNRF.

The 2011 sampling effort targeted both the reintroduction release site (between

Wawaitin and Sandy Falls dams), and the river stretch immediately downstream (between

Sandy Falls and Lower Sturgeon dams). Sampling efforts in subsequent years focused on the reintroduction release site between the Wawaitin and Sandy Falls Dams including the spawning reach, although downstream sampling for adults was also conducted in 2016.

The 2015 effort consisted of targeted adult sampling during the spawning season

(Haxton, 2006), targeted juvenile sampling in July, and a depth-stratified randomized sampling period in September following a protocol used throughout Ontario (Haxton et al., 2014) which enabled comparison with other populations.

The adult sampling consisted of net sets just below the spawning area to target individuals undergoing a spawning migration. The effort consisted of only extra-large gillnet sets as it aimed to target adult individuals. There were 10 nets deployed on the first five nights; on the last night only one net was deployed as the remaining nets were filled with debris and in need of repair due to an overnight increase in flow. The adult sampling effort was conducted May 7th to May 12th 2015, and consisted of total of 51 extra-large gillnet sets, with a total fished time of 1173 hours.

Juvenile sampling targeted deep areas (>8m) of the river as previous studies in similar rivers found high densities of juvenile lake sturgeon occupied these depths (Barth et al., 2009). The targeted juvenile sample effort was conducted July 13th and July 23rd

2015, and consisted of 32 trotline sets, 32 NAG sets, and five extra-large gillnet sets.

Extra-large gillnets were added near the end of the sampling period to increase the sample of adults. The total number of hours fished for all gears combined was 1617

hours, NAG fishing a total of 743 hours, trotlines fishing a total of 757 hours, and extra- large gillnets fishing a total of 116 hours.

As a high density of juvenile lake sturgeon was expected within deep areas (Barth et al., 2009), the randomized recapture effort was depth-stratified to ensure allocated deep water sets. A raster was generated using bathymetry point data collected for the river in

2011 using ArcGIS 10.2.2 (Environmental Systems Research Institute (ESRI)). The river was separated into three depth stratum polygons, shallow (0-4m), moderate (4-8m), and deep (>8m). The three depth stratums contained a third of the sets for each gear type despite the deep stratum being composed of comparatively less area. Although the shallow depth stratum was designated as 0-4m, the sets were in areas >1.5m of depth to allow the gear to fish correctly. There were a total of 32 sets for each gear type, which consisted of four of each gear types being deployed for eight netting nights. Sampling locations were generated in each depth stratum using the random point generation tool in

ArcGIS. Points for each gear type were generated separately. The total number of hours fished for all gears combined was 2084 hours, NAG fishing a total of 701 hours, trotlines fishing a total of 694 hours, and extra-large gillnets fishing a total of 669 hours.

The 2016 sampling consisted of targeted sampling during spawning season in an attempt to collect additional adult samples for the genetic parentage analysis. The sampling took place May 5th - 8th 2016, and May 10th - 12th 2016 and consisted of all extra-large gillnet sets. This sampling consisted of 20 overnight net sets, with 457 total hours fished. This sampling focused on the Wawaitin and Tatachicapika areas (Figure

3.1), two corridors known be used by sturgeon during their spawning movements

(OMNRF, unpubl. data).

Individual sampling consisted of measuring total length (cm), fork length (cm), and mass (g). The first fin ray of the left pectoral fin was taken for ageing and genetic analysis. Individuals were tagged using an external Floy tag to the left of their dorsal fin, and a Passive Integrated Transponder (PIT; Biomark: Model: BIO12.B.01) inserted under the third dorsal scute on the left side. When possible, the sex of the individual was determined based on the presence of gametes (Bruch and Binkowski, 2002). Individuals less than 30 cm in length were not tagged or fin clipped for ageing, and a small genetic sample (<0.25cm2 tail clip) was taken before they were released.

The length-weight equation for all size classes was calculated using the log-round weight and log-total length (Schneider et al., 2000), and the length-weight equation was used to estimate form factor (Bruch et al., 2011). The 2015 population size was estimated from the mark-recapture data using the Lincoln-Peterson modified method assuming a closed population and negligible post-marking mortality (Chapman, 1951). Catch-per- unit effort (CPUE) was estimated for the randomized netting in 2015 for comparison to other populations and was calculated as the number of lake sturgeon caught-per-net-night

(Haxton et al., 2014); other sampling efforts were targeted and not comparable.

Ageing

Pectoral rays were dried for a minimum of 30 days, and cut into 60, 65, and 70μm sections using a Buehler low speed isomet saw (Wilson, 1987). Each section was briefly immersed in 70% ethanol to illuminate growth annuli. Pectoral ray sections were examined using a Nikon SMZ1500 stereomicroscope with a 0.5x objective at 3x magnification. The microscope was calibrated prior to each ageing session using a stage micrometer (1 µm scale). The annuli were counted, with translucent bands interpreted as 83

representing summer growth and opaque bands representing winter growth (Noakes et al., 1999). Ageing error was expected beyond the age of fourteen given lake sturgeon growth slows beyond this age, making the annuli very difficult to count (Rossiter et al.,

1995; Bruch et al., 2009). Digital photographs of the fin ray sections were taken with a

Nikon DS-U2 camera controller (Nikon Instruments Inc.). The distances between the annuli were measured using NIS-Elements (Nikon) imaging software (Nikon Instruments

Inc.), which allowed back calculation to estimate length at age (Francis, 1990). The imaging software was used to attempt to distinguish differences between tagged and untagged adult lake sturgeon in an attempt to identify remnant individuals, assuming that translocated individuals would either show reduced growth due to increased stress or increased energy allocated to movement (Rittenhouse et al., 2007; Hester et al., 2008), or increased growth due to increased access to resources. The translocated individuals fin ray sections were compared to individuals from the Groundhog River natural population provided by OMNRF in an attempt to distinguish differences in annuli between translocated and natural adults.

DNA Extraction and Genotyping

DNA was extracted from small (<0.25cm2) pectoral fin ray sections collected for the ageing analysis using a simple isopropanol precipitation (Wozney et al., 2011) and was eluted in 150μL TE buffer (10mM Tris pH8, 1mM EDTA). Samples from 2016 were extracted using an alkali extraction protocol (Dissing et al., 1996) due to the small number of captured individuals. Successful DNA extraction was confirmed via visualization in a 1.5% agarose gel stained with SybrGreen alongside a molecular mass ladder (Bioshop Canada). Samples were amplified and genotyped at 14 microsatellite loci

that were previously identified for lake sturgeon, 10 of which have been confirmed to exhibit disomic inheritance (McQuown et al., 2002; Welsh et al., 2003; Welsh and May,

2006). The multiplex reactions consisted of a total volume of 15μL containing 1x PCR

Buffer (Qiagen, Mississauga, Ontario), 0.23 to 0.38μM of each primer pair, 2mM MgCl2

(Qiagen, Mississauga, Ontario), 2mM of each dNTP (Bioshop, Burlington, Ontario), 0.2 mg·ml-1 BSA (Bioshop, Burlington, Ontario), 0.025U Taq DNA polymerase (Qiagen,

Mississauga, Ontario), and approximately 10 ng of template DNA (Wozney et al., 2011).

Polymerase Chain Reaction (PCR) assay conditions consisted of an initial denaturation at

95°C for 11 minutes, followed by 35 cycles of 94°C denaturation for 1 minute, 55°C annealing for 1 minute and 72°C extension for 1 minute, and a final extension of 60°C for

45 minutes. Amplified products were run on a Life Technologies 3730 automated sequencer with a ROX-350 size standard (Life Technologies, Foster City, California).

Microsatellite genotypes were scored using GeneMapper version 4.0 (Life Technologies,

Foster City, California). Scoring of the 14 microsatellite loci followed established criteria for data standardization (Welsh and May, 2006).

An additional microsatellite locus was also used for the parentage analysis

[Spl120 (McQuown et al., 2000)], using the same PCR conditions but an annealing temperature of 62°C (Duong, 2010). A random subset of 10% of individuals were genotyped a second time for estimating the rate of genotyping error (Hoffman and Amos,

2005; Pompanon et al., 2005). The program CREATE (Coombes et al., 2008) was used to format genotyping data for analysis in several genetic software programs.

Genetic Analysis

All loci were tested for deviations from Hardy-Weinberg equilibrium (HWE) and linkage disequilibrium using GENEPOP v.4.3 (Raymond and Rosset, 1995). Fisher’s exact tests for heterozygote deficiency were performed for all loci and linkage disequilibrium was tested for all loci pairs, and by population (10,000 each of dememorization steps, batches and iterations per batch). Sequential Bonferroni corrections were used to adjust all significance values to minimize the inflation in Type I error associated with multiple comparisons (Rice, 1989). Tests for heterozygote deficiency showed that two loci (AfuG122 and AfuG68) deviated from HWE in the reintroduced population; AfuG68 was also found to be out of HWE in the source population. As deviations from HWE and potential null alleles have been reported for these loci (AfuG122: Welsh and McClain, 2004; McDermid et al., 2014; AfuG68:

Pyatskowit et al., 2001; McQuown et al., 2002), ML-Relate was used to calculate the allele frequencies with null alleles included (Appendix, Table 3.3). Both loci were removed from the genetic diversity analyses, but retained for the parentage analysis.

Genetic diversity of the reintroduced population was compared to the source population (Adam’s Creek), using juveniles captured between the Wawaitin and Sandy

Falls Dams (excluding juveniles captured below the Sandy Falls Dam). Observed and expected heterozygosity (HO and HE) for each population were estimated in GenAlEx v6.5 (Peakall and Smouse, 2006; Peakall and Smouse, 2012); gene diversity (Nei, 1973) and inbreeding coefficients (FIS) were calculated in FSTAT v2.9.3 (Goudet, 1995).

Standardized allelic richness (Ar) and the number of private alleles were calculated using

HP-Rare (Kalinowski, 2004). These metrics were compared between the reintroduced

and wild populations using Student’s t tests calculated using R v3.1.2 statistical software

(R Development Core Team, 2009). All data were tested for normality (Lilliefors, 1967), and the assumption of homogeneity of variance was tested using Fisher’s exact test. As the number of private alleles deviated from the assumptions of normal distribution and homogeneity of variance, a non-parametric Wilcoxon Sign Rank test was used to test for differences among the populations.

Genetic effective population size (Ne) of the reintroduced and source populations as well the undisturbed Groundhog River population was estimated in NeEstimator v2

(Do et al., 2014) using linkage disequilibrium (LDNe; Waples and Do, 2008). The linkage disequilibrium method is commonly used for single-sample estimates, and is known to provide greater precision with microsatellite data than the temporal method

(England et al., 2006; Luikart et al., 2010; Waples, 2010). Ne estimates for the Adam’s

Creek (source) and Groundhog River populations used samples from adult sturgeon

(n=43 and 54, respectively); generational Ne of the reintroduced population was calculated based on all juvenile samples, excluding larvae (pooled n=124). The effective number of breeders (Nb) for the sampled juvenile cohorts [2007 (n=40), 2008 (n=29), and

2009 (n=24)] were similarly estimated using LDNe in NeEstimator v2. Confidence intervals for Ne and Nb estimates were calculated using the jackknife method, which has been shown to perform better than parametric methods (Waples and Do, 2008).

Parentage Analysis

The parentage analysis included both the juveniles and larvae caught between

Wawaitin and Sandy Falls Dams, and the juveniles caught below the Sandy Falls Dam, to give insight into whether juveniles from the reintroduced population were being entrained 87

through the dam (n=128; McDougall et al., 2013). The genetic probability of identity

(PID) and the probability of identity between siblings (PSIB) were calculated PID and PSIB using Genalex v.6.5.1 (Peakall and Smouse 2006, 2012). The probability of identity (PID), the probability that two individuals within the population having identical multilocus genotypes, was used to quantify the power of the loci to distinguish individuals (Waits et al., 2001). As it is probable that the reintroduced population would contain many siblings, the probability that two full siblings within the population will share identical multilocus genotypes was also calculated (Waits et al., 2001).

Genotype data from 13 adults, 124 juveniles, and 4 larvae were used to estimate parentage of juveniles and larvae using three complementary software packages:

SOLOMON (Christie et al., 2013), COLONY 2.0 (Jones and Wang, 2010), and CERVUS

3.0 (Kalinoski et al., 2007). SOLOMON uses Bayes’ theorem to determine the posterior probability of a parent-offspring pair being false given the frequencies of shared alleles

(Christie et al., 2013), and was used to estimate exclusion-based parentage. As using

SOLOMON for parentage analysis is not ideal for populations with many siblings or missing parents (Christie et al., 2013), the parent and offspring data were also analyzed using COLONY (Jones and Wang, 2010) and CERVUS (Kalinoski et al., 2007).

COLONY assigns offspring into sibling relationships before identifying the most likely candidate parents (Jones and Wang, 2010). Accounting for parent-offspring and sibling relationships concurrently increases the accuracy for assignments in COLONY (Wang,

2007; Walling et al., 2010). The run time was set to long, using a full-likelihood method and assumed a polygamous mating system and 2% genotyping error rate. CERVUS assigns the most likely parents as the individuals with the highest natural log of the

likelihood-odds ratio (LOD) score for specific adult-juvenile pairs (Kalinoski et al.,

2007). The CERVUS analysis consisted of a simulation of 10,000 offspring with a strict level of 95% and a relaxed level of 90% to determine confidence based on allelic frequencies. Adults with the highest positive LOD score in CERVUS and the highest maximum likelihood score in COLONY were considered to be probable parents. Only parentage was inferred (i.e. not paternity or maternity), as sexing data were not available for captured adults.

Results

Field Season

Netting yielded a total of 169 lake sturgeon captures, with 139 unique individuals caught (30 recapture events), including 21 individuals captured below Sandy Falls Dam

(Figure 3.2). Of the total number of captures, 32 lake sturgeon were caught in 2011, including 19 juveniles captured below the Sandy Falls Dam; 15 lake sturgeon (6 adults and 9 juveniles) were captured in 2013; 114 sturgeon were captured in 2015 including 27 previously captured individuals (described below). Sampling efforts in 2016 resulted in the capture of 6 sturgeon (5 adults and 1 juvenile), of which 3 adults had been previously captured (Appendix: Table: A3.1). Of the 13 adults captured between 2011 and 2016, 8 had tags from the 2002 translocation event. Size data and ageing and genetic samples from 1 juvenile and 1 adult captured by OMNRF personnel below the Sandy Falls in the fall of 2016 were also included. Across all gear types, juvenile length-weight relationship was calculated (Figure A3.3), and the form factor was 5.98.

The targeted adult sampling in 2015 captured 2 adult males, one of which still carried a tag from the translocation effort in 2002. The July 2015 targeted juvenile sampling captured 78 sturgeon, with 57 being new individuals, and 21 being previously captured individuals. The September 2015 stratified random sampling effort yielded 34 catches (all juveniles), 6 of which had previously been caught, with 5 individuals recaptured from the 2015 year.

The fall 2015 randomized recapture resulted in 19 large mesh and 2 extra-large mesh gillnet captures with a mean total length of 665 cm (SD=184 cm); the catch per unit effort was 0.59 (SE=0.16) and 0.06 (SE=0.02) for the large and extra-large mesh nets, respectively. The modified population estimate was 413 (134-692), with 68 individuals marked or captured in the 2015 season, 36 individuals caught during the recapture period, only 5 of which were recaptures leading to large confidence intervals.

Ageing

A total of 128 lake sturgeon from the reintroduced population were aged, with 77 being captured in the 2015 sampling. Thirteen adult sturgeon, including 8 with translocation tags, were aged and determined to be from the 1967 to 1997 cohorts. The adult reintroduced lake sturgeon pectoral fin rays had thick dark bands, suggesting slow growth that generally coincided with the translocation year. This contrasted with pectoral fin rays from the Groundhog River (natural population) sturgeon, which did not have the thick dark band (Appendix, Figure A3.1). A total of 115 juveniles were aged, including

20 individuals captured below the Sandy Falls Dam; age estimates were verified from back-calculating length at age based on distance between the annuli. The 2007, 2008, and

2009 cohorts appeared to have strong year classes (Figure 3.3). The capture of 4 larvae in

2014 indicated the presence of a 2014 cohort as well.

Genetics

In total, 124 juveniles, 4 larvae and 13 adult lake sturgeon from the reintroduced

Mattagami population were genotyped. Repeated genotyping of a random subsample of

10% of these individuals resulted in a mean genotyping error estimation of 0.2% across loci (Appendix: Table 3.2). The total number of alleles per locus ranged from two

(AfuG195, AfuG204) to seven (AfuG68b); allele frequencies for each locus are provided in Appendix: Table A3.3. All genetic diversity measures for the reintroduced Mattagami population were comparable to the source population, with no significant differences

(Table 3.1). The FIS value for the reintroduced population (-0.23) was significantly different from zero due to pooling genetic data from the multiple juvenile cohorts (Table

3.1).

Effective population size (Ne) estimates for the reintroduced Mattagami population (20.4 (13.5-30.5)) was significantly lower than the Adam’s Creek (source) population (∞ (72.0-∞)). The estimated effective number of breeders (Nb) for the 2007,

2008, and 2009 Mattagami cohorts were very similar (Table 3.2), and comparable to the population Ne estimate. The Nb estimates indicated multiple adults contributed to each cohort; although the sizeable confidence intervals were likely a result of limited sample sizes, the lower confidence limits suggest that more than ten adults contributed to each cohort (Table 3.2).

All parentage analysis revealed parent-offspring relationships between juveniles and adults that retained tags from the translocation (Table 3.3). The combined PID for the 91

-8 -4 15 loci used was 1.1 x 10 and the PSIB was 2.0 x 10 . The SOLOMON power output indicated that even with relying on zero mismatches (full matching/exclusion), there was still a high number of expected false pairs (Appendix: Figure 3.2). The SOLOMON analysis identified eleven juveniles that had no mismatches from more than two adults, highlighting the limitation of the set of loci to resolve parent-offspring relationships. The

Bayesian analysis accounting for siblings resulted in two juveniles matching parents with low probability of the pair being false, adult 35 and Matt16 (probability of pair being false= 0.013), and adult 32 (tagged: OPG 1448) and 13 (below Sandy Falls Dam; probability of pair being false= 0.036). The output from COLONY identified 9 maternal parents and 17 paternal parents based on a probability of greater than 0.90. As the

CERVUS analysis resulted in 11 of the same parent-offspring relationships (73% agreement; Table 3.3), the parents selected by both methods were therefore considered as actual parents. Ten of the eleven parents considered the most likely parent by both

CERVUS and COLONY had retained tags from the original transfer. Three of the juveniles located downstream (below Sandy Falls Dam) were offspring of translocated individual (OPG 1448) located at the release site. It should be noted that as sexing data were not available for some individuals, gender could not be inferred for all putative parents.

Discussion

All objectives of our study were successfully achieved. The experimental translocation of wild adult lake sturgeon appeared to be successful, as evidenced by the recapture of translocated adults and capture of several year classes of wild recruits at the release site (between Wawaitin and Sandy Falls Dams) and below Sandy Falls Dam.

Ageing of the juvenile samples showed recruitment starting in 2006. Successful reproduction by translocated adults and juvenile recruitment was confirmed by parentage analysis linking juveniles to parents that had retained tags from the original translocation event. Genetic diversity measures for the reintroduced and source populations were comparable, other than reduced Ne in the reintroduced population.

Adult and Juvenile Sampling

Sampling resulted in multiple adult catches including adults that retained tags from the original translocation. The catch-per-unit effort of adults with extra-large mesh nets (0.06, SE=0.02) was much lower than the other Ontario lake sturgeon populations

(mean=0.48; 0-1.73; Haxton et al., 2014), reflecting the small numbers and lower abundance of the reintroduced adults relative to natural lake sturgeon populations.

Although it is possible that untagged individuals were originally present within the area, as no netting effort was conducted prior to the translocation effort, evidence from the adult ageing data suggests that this is unlikely. The thick, dark bands observed in pectoral rays of aged adults from the reintroduced population were not present in the undisturbed

Groundhog River population, and coincided with the translocation event independently of individual fish ages. This may be a result of reduced growth due to increased stress or increased energy allocated to movement (Rittenhouse et al., 2007; Hester et al., 2008). 93

The consistent temporal appearance of this band in adults within the reintroduction site suggests that untagged individuals may have experienced tag loss in the ten years after translocation, rather than representing a remnant cryptic population.

The number of juveniles captured within and below the reintroduction site suggests that the translocation effort was successful for wild reproduction and juvenile recruitment. The lack of juvenile cohorts before 2006 suggests that a reproducing population was likely not present at the site prior to the translocation, and that translocated adults experienced a multi-year delay before reproducing. Post-release telemetry and larval driftnetting data (2002-2003) support the lack of reproduction in the early years (OMNRF, unpubl. data), as well as confirmation of spawning in later years

(2014-2015; refer to Chapter 2). Other species have exhibited similar delays in reproduction after release (Comly, 1993; Vandel et al., 2006); potentially due to Allee effects (Deredec and Courchamp, 2007). The combination of stress and unfamiliarity with new habitat may also reduce or restrict short-term reproductive success (Busch and

Hayward, 2009). Elevated stress can impede reproduction (Teixeira et al., 2007; Busch and Hayward, 2009; Scarlata et al., 2012), and unfamiliarity with a new location and resources may affect site choices, energy budgets and reproduction of translocated individuals (Armstrong and Perrot, 2000; Kermink and Kelser, 2013).

All three genetic parentage methods detected relationships between tagged translocated adults and juveniles within the population. The limited resolution from the

SOLOMON exclusion results may have been due to the high genetic similarity among translocated adults due to limited effective population size of the source population, as well as sampling a small population with many siblings, and the limited number of

genotyped loci. By contrast, COLONY is useful for populations with many siblings as expected in the translocated population (Harrison et al., 2013). The COLONY and

CERVUS analysis resulted relationships between 10 juveniles and 4 parents that retained tags from the translocation, and one juvenile related to an untagged parent. The inferred parentage for three juveniles caught below Sandy Falls Dam to adults above the dam was not surprising, as other studies have shown larval drift occurring up to 45 river kilometers

(rkm) downstream (Auer and Baker, 2002).

The catch-per-unit effort of juveniles based on large mesh nets (0.59, SE=0.16) was comparable to the higher end of other Ontario lake sturgeon populations (mean=0.3;

0-2.29; Haxton et al., 2014), suggesting the reintroduced juvenile population has a similar relative abundance to other lake sturgeon populations. Although the presence of the first generation of juveniles is an encouraging indicator of short-term success (Serfass et al.,

1993; Germano and Bishop, 2009), monitoring for a sufficient amount of time to determine long-term persistence may take several decades for long-lived species such as the lake sturgeon (Dodd and Seigel, 1991).

Genetic Analysis

Despite the small number of founders being released into the fragmented section of the Mattagami River, there were no significant differences in genetic diversity between the reintroduced and source populations, indicating that the genetic diversity of the population was captured in the reintroduction effort. Similar to most translocations, this effort reintroduced a small number of adults, which is often associated with increased genetic drift and inbreeding in future generations, resulting in loss of diversity (Leberg,

1993; Wolf et al., 1996; Frankham et al., 2010). Many lake sturgeon populations have 95

undergone significant historical declines in abundance without significant decreases in genetic diversity (DeHaan et al., 2006; Drauch and Rhodes, 2007; Welsh et al., 2008;

McDermid et al., 2011; Wozney et al., 2011; McDermid et al., 2014). The unique life history characteristics of sturgeon, such as longevity, long generation time, and overlapping generations, may buffer against loss of genetic diversity (McDermid et al.,

2011; Wozney et al., 2011; McDermid et al., 2014). Potential losses of genetic diversity may be lessened by rapid population growth after reintroduction (Allendorf and Luikart,

2007), which may be the case for the reintroduced Mattagami population.

The marked reduction in Ne of the reintroduced population compared to its source was expected due to the limited number of translocated adults. Ne has been shown to be a sensitive measure of genetic loss compared to other diversity metrics in lake sturgeon populations (McDermid et al., 2014) and other reintroduced species (Miller et al., 2011;

Ottewell et al., 2014). This result indicates that the limited number of founding adults has left a genetic signature, and that without rapid growth or additional reintroductions, the population may eventually be susceptible to loss of genetic diversity over time (Ottewell et al., 2014). The Ne of the reintroduced Mattagami population is well below conservation recommendations of Ne ≥100 over several generations and Ne ≥1000 to maintain long- term evolutionary potential (Frankham et al., 2014). Although Welsh et al. (2010) recommended translocating 250 females and 250-1,250 males of diverse age groups over

25 years to rehabilitate lake sturgeon populations in the Great Lakes, this would not be feasible for the finite habitat available in the 35km river segment for the reintroduced population. Translocation of additional fish into the reintroduced Mattagami population could increase local Ne and reduce future genetic threats to the population (Frankham et

al., 2014), although adult transfers should not exceed 5% of the source population (Welsh et al. 2010).

The effective number of breeders (Nb) for each juvenile cohort provides insight into the effective number of contributing reproductive adults for a single reproductive season. Cohort values of Nb appeared to be comparable to the population Ne, although these estimates are based on small sample sizes with large confidence intervals. It should be noted that when sample sizes are smaller than Ne, the linkage disequilibrium method may be biased (England et al., 2006; Waples, 2006; Waples and Do, 2008). In a multi- year study of a natural lake sturgeon population, Duong et al. (2013) found Nb to be temporally consistent and comparable to overall Ne, similar to what was observed in this study. As lake sturgeon life history characteristics include intermittent spawning

(Harkness and Dymond, 1961; Billard and Lecointre, 2001; Duong et al., 2013), the Nb estimates for the juvenile cohorts, together with the demographic and ageing data, should be interpreted as evidence of successful spawning by translocated adults but not overinterpreted. In other long-lived interoparous species, values of Nb close to Ne are thought to reflect breeders from a mixture of age classes and generations (Gautschi et al.,

2003; Beebee, 2009). As the reintroduced population does not yet have interbreeding of fish from multiple generations, and lake sturgeon life history characteristics include intermittent spawning (Boreman, 1997; Billard and Lecointre, 2001), the Nb estimates should be interpreted with caution.

Demographics and Persistence

The number of juveniles and juvenile cohorts observed within the formerly unoccupied river segment is encouraging. The somatic growth rate of juveniles within the 97

reintroduced Mattagami population was comparable to other lake sturgeon populations

(Power and McKinley, 1997; Bruch et al., 2011). The form factor (mFF=5.98) was in the midrange of other lake sturgeon populations (Bruch et al., 2011), and slightly lower than form factors reported for other Mattagami populations (mFF=6.34; Power and McKinley,

1997; Bruch et al., 2011). This suggests that habitat and food resources for the reintroduced population are not currently limiting. As well as the number of juveniles captured, it is encouraging that the size estimate for the reintroduced population based on data from 2015 was substantially greater than the number of adults (n=51) that were translocated in 2002.

Successful reproduction and the presence of multiple yearclasses are critical to the primary objective of reintroduction, which is to achieve a low extinction risk in the population of interest (Burgman et al., 1993). Translocation efforts tend to release small numbers of individuals, which are subject to extinction risks due to demographic stochasticity and demographic disequilibrium (Leigh, 1981; Lande, 1988; Robert et al.,

2007). Populations with more than 100 individuals are more likely to avoid threats related to demographic stochasticity (Mace et al., 2008). Long-lived, iteroparous species such as lake sturgeon, however, experience reduced impacts to demographic stochasticity, and releases of less than 100 individuals are typically considered adequate to ensure their short-term persistence (Legendre et al., 1999, but see Welsh et al., 2010). As the reintroduced Mattagami population now has been estimated to comprise of more than 100 adult and sub-adult individuals from several year classes, the risk of extinction due to demographic stochasticity could be considered low.

Conclusions

Dam fragmentation continues to be a major factor affecting lake sturgeon populations and their recovery (Rochard et al., 1990; Haxton et al., 2014). This study demonstrates that translocations of adult lake sturgeon can mitigate fragmentation effects caused by dams, providing that sufficient habitat is present to support all life stages in sufficient numbers to establish a self-sustaining population. In this instance, reintroduction via wild translocation appears to have been successful based on natural recruitment as well as adult survival. Assessment of future translocation efforts would be strengthened by pre-introduction surveys of candidate introduction sites to confirm that the species of interest is no longer present. Other recommendations include marking of translocated individuals with both internal and external tags to guard against tag loss and facilitate identification if follow-up assessment is done. Recording length, weight, gender, and age data of released individuals would also be valuable. Taking genetic samples from all translocated individuals would also facilitate future assessment of survival, reproduction and recruitment. Post-release monitoring is especially important for assessing success, and this study affirms that timing of sampling after release should be based on species biology to maximize the probability of detecting successful reproduction. As biodiversity challenges in fragmented aquatic systems continue to increase (Sala et al., 2000), translocation and assisted reintroductions may become increasingly necessary approaches to reverse local extirpations. Treating translocations and reintroductions as experiments in reintroduction biology should improve both our understanding of required elements for re-establishment and the success of these efforts.

Acknowledgements

Field sampling was conducted by Derrick Romain, Charlie Hendry, Adam

Rupnik, Emma Harten, Tarryn Adams, and Matt Werner (OMNRF), along with countless volunteers from Timmins Fur Council, Club Navigateur-La Ronde, Glencore Canada

Corporation, Kidd Operations, and Lake Shore Gold. Laurent Robichaud for his guidance and knowledge. Staff of the OMNRF Fish Genetics Lab provided invaluable guidance and troubleshooting with lab work and genetic analysis. Funding was provided by

Ontario Power Generation (OPG), Lake Shore Gold, and Glencore Canada Corporation,

Kidd Operations, the Ontario Ministry of Natural Resources and Forestry (OMNRF), the

Ministry of Environment and Climate Change (MOECC) Ontario Community

Environment Fund, and managed by the Wintergreen Fund for Conservation. None of the authors have any conflicts of interest to declare.

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Table 3.1: Genetic diversity measures for the reintroduced Mattagami lake sturgeon population (n=104) sampled between 2011-2016, and the source (Adam’s Creek) population (n=43) provided by the OMNRF, based on 12 microsatellite loci. Values indicate mean and standard error (SE) for each metric. Statistical significance for differences between the source and reintroduced populations was tested using Student’s t- test.

Mattagami Adam’s Creek Significance

Observed heterozygosity (HO) 0.57 (±0.07) 0.53 (±0.06) t22=-0.47, p=0.64

Expected heterozygosity (HE) 0.47 (±0.05) 0.49 (±0.06) t22=-0.44, p=0.66

Gene diversity (Nei) 0.47 (±0.05) 0.50 (±0.06) t22=-0.34, p=0.74

Inbreeding coefficient (FIS) -0.23 -0.07 t20=-1.50, p=0.15

Allelic richness (Ar) 2.80 (±0.28) 3.16 (±0.40) t22=-0.76, p=0.46

Private alleles (Ap) 0.07 (±0.05) 0.30 (±0.14) W=52, p=0.20

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Nb Cohort Year n 0.05 0.02 2007 40 24.4 (13.3-49.5) 26.5 (14.4-55.8) 2008 29 22.7 (11.6-65.4) 21.6 (12.2-48.4) 2009 24 24.0 (10.8-121.6) 39.7 (16.0-∞)

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Table 3.3: Parent-offspring pairs identified within the reintroduced Mattagami River population using CERVUS and COLONY analysis based on 15 microsatellite loci, showing tag numbers of adults from the original translocation effort.

Offspring Parent Tag Numbers Matt37 31 OPG 1446 13 32 OPG 1448 Matt40 32 OPG 1448

9 32 OPG 1448 Matt11 43 OPG 1455 Matt17 43 OPG 1455 OPG 1455 Matt46 43 Fall2 43 OPG 1455 Fall4 43 OPG 1455 42 45 OPG 1457 Fall2016_2 16Matt_2_MP1

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Figure 3.1: Map of the site where lake sturgeon were reintroduced in 2002 (boxed area) within the Mattagami River watershed (James Bay drainage basin, Ontario). The inset map shows the location of the study system in Ontario, Canada. The source population for the translocated adults is starred. Black squares represent hydroelectric facilities.

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Figure 3.3: Age structure of juvenile lake sturgeon caught within and immediately below the reintroduction area (n=115), based on back calculation pectoral fin ray growth annuli. Bars show total numbers of captured aged juveniles across all sampling years (2011-2016).

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Chapter 4: General Discussion

The results of both chapters clearly demonstrated that the translocation effort was successful in re-establishing lake sturgeon in the 35 km fragment of the Mattagami River, and highlighted the importance of research to inform species reintroductions and maximize translocation success. Chapter 2 confirmed that a portion of the translocated adults stayed within the release area and moved upstream to spawning habitat during the spring freshet. Chapter 2 also demonstrated the importance of quantifying the post- release and exploratory movements that can occur soon after a translocation effort versus post-establishment. The results of Chapter 3 indicated that the experimental translocation was successful with the capture of multiple juvenile yearclasses within the release site, and that the translocated population largely retained the genetic composition of its source population. Furthermore, the population estimate ten years after release showed the presence of substantially more sturgeon than the original 51 released individuals, despite early losses from post-release dispersal (Chapter 2). Genetic analysis also confirmed that adults that retained tags from the translocation were related to juveniles within the population. Both the telemetry and catch data indicated the importance of deep water pools for juvenile and adult lake sturgeon. The combined findings from this study support two goals of the Ontario recovery strategy for lake sturgeon by re-establishing a population in an extirpated site, and addressing knowledge gaps to inform recovery efforts. The results also contribute to the conservation biology of lake sturgeon in general, and may have implications in the broader field of reintroduction biology.

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Reintroductions as a conservation tool

Reintroduction biology is the research undertaken to improve the outcomes of conservation translocations, and is a relatively new area of conservation biology

(Armstrong and Seddon, 2008; Ewen et al., 2012; Jachowski et al., 2016). Over the past few decades, reintroduction biology has accumulated hundreds of peer-reviewed publications (Fischer and Lindenmayer, 2000; Armstrong and Seddon, 2008). However, long-term studies with experimental design are often not the objective of reintroduction managers (Caughley, 1994). As a result, many studies are opportunistic and retrospective evaluations of translocation efforts, and often lack components of robust experiments including clearly defined hypotheses, replicates, and controls (Armstrong and Seddon,

2008). Although this study resulted in important data that can inform future reintroduction efforts, it was largely a post-hoc study. Reintroductions should strive to set goals and questions before release to maximize the knowledge gained from the effort.

Many factors influence reintroduction success, including sex and age composition of released individuals, habitat quality, and post-release movement (Moseby et al., 2011;

Ewen et al., 2012). Post-release monitoring allows assessment of habitat utilization and the influence of environmental factors on the survival and population dynamics of individuals (Boyd and Bandi, 2002; Nichols and Armstrong, 2012). The information collected is critical to informing management decisions, enhancing post-release survival, and planning future reintroduction efforts (McCarthy and Possingham, 2007; Bernardo et al., 2011; McCarthy et al., 2012).

Monitoring the movement of translocated lake sturgeon post-release and a decade later in Chapter 2 gave insight into the post-release dispersal, as well as exploratory, post-

117 acclimation, and spawning movements. Understanding the factors influencing post- release dispersal and exploratory movements is important as these movements affect translocation success (Armstrong and Seddon, 2008; Armstrong and Reynolds, 2012; Le

Gouar et al., 2012), and the affects may differ between the phases of population establishment (Sarrazin, 2007; IUSN/SCC, 2013).

The establishment phase is sensitive to post-release dispersal, particularly for small release groups which is often the case in translocation efforts (Le Gouar et al.,

2012). This effort released only 51 lake sturgeon and a proportion of the radio-tagged individuals dispersed downstream out of the release site (6 of the 13 translocated individuals with radio-tags); the total number of introduced individuals that remained in the area is unknown. If the proportion of radio-tagged lake sturgeon that departed from the release area was representative of the founding population as a whole, it is possible that approximately only half of the originally translocated individuals remained, substantially reducing the founding population size, which could also have reduced the probability of translocation success (Armstrong and Seddon, 2008). Twelve adults were captured in the sampling of the release reach between 2011 and 2016, none of which were individuals radio-tagged in 2002, although eight had retained external tags from the translocation. During early telemetry (2002-2003), seven translocated individuals remained within the release fragment, and the capture of adults that retained external tags a decade later proves that at least a subset of released individuals remained long-term.

The other six radio-tagged individuals dispersed downstream over Sandy Falls GS, and one untagged adult was caught downstream between Sandy Falls and Lower Sturgeon GS

118 in 2016. The early downstream dispersal of a subset of the translocated lake sturgeon may be indicative of homing behaviour.

Homing is the process where released individuals migrate back to their source habitat and can result in translocation failure (Linnell et al., 1997). Homing behaviour occurs often in a variety of translocated species (Linnell et al., 1997; Oliver et al., 1998;

Germano and Bishop, 2009; Mireles et al., 2012; Hinderle et al., 2015). As lake sturgeon tend to show site fidelity to natal spawning grounds (DeHaan et al., 2006), this effort released translocated individuals upstream of their source habitat, making homing behavior possible. In fragmented riverine habitats selecting release sites below the source habitat could potentially eliminate the potential for homing behaviour.

Early monitoring (2002-2003) showed translocated individuals using the habitat downstream of the release site above the Sandy Falls GS. Early dispersal away from the release site may indicate the release habitat is unsuitable or may be a result of chronic stress (Stamps and Swaisgood, 2007; Dickens et al., 2009; Ewen et al., 2012). Radio- tagged individuals that remained in the release site long-term (2011-2016) used mostly areas upstream of the release site, indicating the value of this habitat (MacArthur and

Pianka, 1966; Bar-David et al., 2009). The careful selection of release sites may reduce the amount of post-release dispersal, for example choosing a release site near refuge habitat (Bodinif et al., 2012; Jachowski et al., 2012). In this translocation, releasing lake sturgeon above the areas frequented after the exploratory phase may have decreased post- release dispersal as individuals would need to pass valued habitat while dispersing downstream. The timing of release may also decrease post-release dispersal: in another lake sturgeon translocation, adults translocated in spring spawned prior to dispersing

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(Koenings et al., 2015), potentially contributing to the next generation. This result may merit more research on the timing of release to maximize genetic contribution to translocated populations.

Monitoring movements of translocated sturgeon during post-release and acclimated periods showed that habitat use in the two periods differed. In the establishment phase, reintroduced individuals tend to spend time exploring their novel environment in order to locate habitats necessary to support their needs (Frair et al., 2007;

Berger-Tal and Saltz, 2014). The process of learning and adapting to the release environment is reflected in the movement, behaviour, and habitat selection of translocated individuals (Vickery and Mason, 2003; Bose and Sarrazin, 2007; Berger-Tal and Saltz, 2014). The difference in habitat use and movement patterns between the two time periods suggests that acclimation took over a year and up to five years for the translocated sturgeon.

The movements in the reproductive season also differed between the post-release and acclimated periods. Lake sturgeon tend to spawn below the furthest upstream natural or artificial barrier (Auer, 1996b; Golder Associates Ltd., 2011), and acclimated radio- tagged sturgeon used habitat by the Wawaitin GS during the spring freshet. Larval lake sturgeon were caught in driftnets in the Wawaitin GS spillway in 2014, which confirmed these movements were for reproduction. Conversely, none of the radio-tagged lake sturgeon moved upstream to Wawaitin GS in spring 2003. During the establishment period, individuals allocate energy to exploration of their habitat (Russel et al., 2010;

Berger-Tal and Avgar, 2012) which can have costs (e.g. energetic demands; Eliassen et al., 2007), and reduces energy available for reproduction. Translocations are also

120 inherently stressful as they release individuals into a habitat that is unfamiliar (Parker et al., 2012), chronic stress can cause impacts such as decreased reproductive output

(Herzog et al., 2009). Radio-tagged lake sturgeon not moving to reproductive habitat in the post-release period likely indicates reproduction delay resulting from stress or allocating energy to explorative movements. It remains unknown whether sturgeon without radio-tags moved upstream to reproductive habitat. The ageing of juvenile samples in Chapter 3 indicates that 2007 was the first strong cohort, and this evidence supports delayed reproduction. Important habitat for various life stages such as reproduction can be determined from monitoring the movements of translocated individuals, which aids filling knowledge gaps for species’ protection.

Monitoring the movements of translocated sturgeon identified areas of high habitat use, which has implications for the ongoing management and protection of the population (Rosenfield and Hatfield, 2006). The radio-tagged adults used different habitat seasonally indicating they need to fulfil different requirements through the year. Radio- tagged lake sturgeon used deep holes within the release reach heavily, indicating the importance of this habitat. During the winter season all radio-tagged individuals were tracked to deep holes, this is consistent with other lake sturgeon populations that have been shown to aggregate in a fraction of available habitat during winter (Rusak and

Mosindy, 1997; Thayler et al., 2017). The duration lake sturgeon spend aggregated in overwintering habitat indicates that this habitat is critical to their survival. Radio-tagged adults were most often tracked in deep holes in summer and fall seasons as well, emphasizing the importance of deep water habitats for adult lake sturgeon.

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Chapter 2 focused on the movements of radio-tagged adults, but the high number of juvenile catches in deep pools during the sampling for Chapter 3 indicates deep water habitat is also important for juvenile lake sturgeon. Juvenile and subadult lake sturgeon in other populations have been shown to utilize pool habitats during periods of high flow

(Seyler, 1997), and in the Winnipeg River juvenile lake sturgeon preferred deep water habitats (Barth et al., 2009). The Mattagami River is large and fast-flowing, which may in part explain the high number of juvenile catches in the deep pools. The results of Chapter

2 and 3 indicate the importance of deep pools for both juvenile and adult lake sturgeon, meaning that deep habitat is likely an important feature throughout lake sturgeon life history, which should be considered when selecting potential reaches for lake sturgeon reintroduction.

Spring movements of the radio-tagged adults identified potential reproductive habitat by Wawaitin GS for the translocated population. The radio-tagged sturgeon used the deep habitat within the Wawaitin GS tailrace prior to spawning in the spillway, which was confirmed as reproductive habitat by the capture of larval lake sturgeon in the spillway. Similar to other populations, the identified reproductive habitat was located underneath of the furthest upstream barrier (Auer, 1996b; Golder Associates Ltd., 2011).

The presence of deep habitats can be a primary factor in spawning site selection (Bruch and Binkowski, 2002) as deep habitat is often used for staging. Whether lake sturgeon use the GS tailrace or spillway may depend on the amount of water released during spawning season (Murray and MacDonald, 2010). In high spill years the Pointe du Bois population spawns within the spillway of the GS and in low spill years they use the tailrace (Murray and MacDonald, 2010). The larval lake sturgeon catches from 2014

122 confirm the spillway as reproductive habitat, however the Wawaitin GS tailrace was not targeted due to dangerous flows. There were no larval sturgeon caught in the spillway in

2015, which was a low spill year, potentially making conditions in the GS tailrace more ideal. Spawning habitat is critical to the survival and is important to identify to confirm reproduction and for the ongoing management and protection of the species (McKinley et al., 1998; Species at Risk Act, 2002; Rosenfield and Hatfield, 2006).The identification of critical habitat is one of the six major goals for lake sturgeon recovery in the provincial recovery strategy (Golder Associates Ltd., 2011), and therefore is of importance for the ongoing management and conservation of the species.

Releasing lake sturgeon into short fragments of river has the potential for failure if habitat to support all life stages is not available within the reach. Previous studies have estimated that 250-300 km unfragmented habitat is required to support a self-sustaining lake sturgeon population (Auer, 1996a). However, recent studies focusing on the Nelson and Winnipeg Rivers in Manitoba have contested this, stating this is not the case in all rivers, and that some river fragments as little as 10 km may contain the habitat necessary to support a lake sturgeon population (McDougall et al., 2017a; McDougall et al., 2017b).

The genetic structure in these areas predates hydroelectric development, indicating the populations have thrived in these small river fragments for many generations (McDougall et al., 2017b). The provincial recovery strategy states that small river fragments may be able to support lake sturgeon, but they must contain connected habitat necessary for all life stages (Golder Associates Ltd., 2011). This study demonstrated that lake sturgeon population can spawn, establish, and grow within a 35 km river fragment. A river fragment has the potential to support a self-sustaining lake sturgeon population if it

123 contains habitat sufficient for spawning, larval drift, and habitat required for settling and establishment (including wintering) of early life stages in downstream areas (McDougall et al., 2017a). However, low-gradient rivers with minimal habitat diversity may require hundreds of kilometers to support a self-sustaining population (McDougall et al., 2017a).

In short fragments that lack settling habitat, a majority of larval lake sturgeon would likely drift downstream of the reach, as larvae can drift up to 61 km downstream of the spawning area (Auer and Baker, 2002). When selecting potential release reaches, it is important to identify habitat to support all life stages.

The habitat within the reintroduction reach used during lake sturgeon in early life stages is unknown and a limitation of the study. The juvenile captures a decade after release indicate that lake sturgeon are surviving and growing within the release reach, but lack of information on the habitats used by young sturgeon may hinder protection of habitat important for early life stages. The habitat preferences of early life stage lake sturgeon are identified as a knowledge gap in the provincial Lake Sturgeon Recovery

Strategy (Golder Associates Ltd., 2011). Future research should focus on identifying habitats important for early life stage lake sturgeon.

The genetic diversity of the translocated population was comparable to that of its source, apart from a reduction in the effective population size (Ne). Ne has been shown to be more sensitive to contemporary genetic loss compared with other diversity metrics

(Miller et al., 2011; McDermid et al., 2014; Ottewell et al., 2014). This indicates that the number of founders in the translocated population was small; without rapid growth or additional releases, the population will be susceptible to genetic diversity loss over time

(Ottewell et al., 2014). Currently the Ne of the translocated population is well below the

124 recommendation for conservation of Ne ≥100 over several generations and Ne ≥1000 to maintain long-term evolutionary potential (Frankham et al., 2014). However, a high number of juvenile catches indicates the population is growing rapidly, which can reduce genetic diversity loss (Allendorf and Luikart, 2007). Genetic diversity within a reintroduced population is important for long-term persistence as it has consequences for its` ability to adapt to change (Frankham et al., 1999; Keller et al., 2012; Pierson et al.,

2015). As reintroduction efforts tend to release small numbers of individuals making them vulnerable to genetic diversity loss through inbreeding and genetic drift, it is important to monitor genetic structure and diversity at the various stages of reintroduction

(Keller et al., 2012). This is especially true for translocations where released individuals emigrate out of the release site as seen in a subset of the released radio-tagged lake sturgeon in Chapter 2. Multiple releases may be considered to increase genetic diversity in situations where the population is losing diversity through emigration or growing too slowly to buffer genetic diversity loss (Tracy et al., 2011; Weiser et al., 2013).

The effective number of breeders (Nb) for each juvenile cohort was comparable to

Ne, although Nb calculations were based on small sample sizes which could potentially bias the estimates (England et al., 2006; Waples, 2006; Waples and Do, 2008). Another study on a natural lake sturgeon population found Nb to be comparable to overall Ne

(Duong et al., 2013). This suggests that much of the genetic diversity is captured in each lake sturgeon spawning event, meaning that multiple families contributed to recruitment.

Due to unique life history characteristics such as intermittent spawning (Harkness and

Dymond, 1961; Billard and Lecointre, 2001; Duong et al., 2013), Nb estimates for lake sturgeon should be interpreted with caution. Similar Nb and Ne values are thought to

125 reflect breeders from a mixture of generations in long-lived interoparous species

(Gautschi et al., 2003; Beebee, 2009). Although Nb and Ne are similar for the translocated population, only one generation is of reproductive age so this is not a reflection of the breeding between multiple generations. Nonetheless, the Nb estimates, coupled with the demographic and ageing data, are indicators that translocated adults are spawning successfully.

Role of reintroductions in lake sturgeon conservation

The knowledge gained from assessing movement, habitat use, demographics, and genetics of this translocated lake sturgeon population can be used inform future reintroduction efforts. Restoring lake sturgeon to areas they once occupied is outlined as a suggested goal in the Lake Sturgeon Recovery Strategy for Ontario (Golder Associates

Ltd., 2011). As sturgeon recovery is often impeded by dams in rivers (Rochard et al.

1990; Haxton et al., 2014), and fragmentation can prevent natural recolonization of areas historically occupied by lake sturgeon, translocation of wild fish can be used as a human- mediated tool to mitigate at least some effects of dams. Although this case study outlines a successful lake sturgeon translocation, it also raises many questions regarding maximizing translocation success and selecting release habitat. There are many potential opportunities for reintroducing lake sturgeon into isolated river segments across their historical range, but managers must consider habitat required for all life stages when selecting fragments for restoration. After identifying suitable reintroduction sites, managers must identify healthy source populations that can provide the individuals necessary for re-establishing a population.

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Within the release area of this translocation, the historical factors that led to the local extirpation of lake sturgeon were likely a combination of overexploitation, fragmentation, and water quality issues due to mining and log drives (Charles Hendry,

OMNRF, pers. comm.), similar to other declines in the Moose River Basin (Golder

Associates Ltd., 2011). Water quality issues impact lake sturgeon by altering water chemistry (Haxton and Findlay, 2008), contributing to egg mortality (Kempinger, 1988), and reducing macroinvertebrate populations (Beamish et al., 1998). Other lake sturgeon populations have shown signs of recovery after water quality improvements (Mosindy and Rusak, 1991). The establishment of fishing protections and water quality standards allowed reintroduction to be a viable option to re-establish lake sturgeon within this isolated river segment.

Although fragmentation prevented the species from naturally recolonizing the study area, the translocation of wild adults was successful despite the relatively small size

(35 km) of the reintroduction site. This suggests translocation is a viable tool to meet the goal of reintroducing populations in historically occupied areas (Golder Associates Ltd.,

2011), even in the heavily fragmented rivers of the Anthropocene. The results suggest that fragmented systems can support lake sturgeon populations providing that they have the habitats necessary to support the diverse needs of the species (Golder Associates Ltd.,

2011), as has been shown in other river fragments (Barth et al., 2009, McDougall et al.,

2017a). It is important to identify habitat for spawning, drifting, settling, establishing, and overwintering are present within the fragment prior to releasing translocated individuals

(McDougall et al., 2017a). Prior to selecting a reach for lake sturgeon reintroduction, managers should confirm that the factors that contributed to the initial decline or loss

127 have been mitigated, and that the site includes diverse habitat with deep sections, especially in small reaches.

The movement, demographic, and genetic monitoring to assess the success of this translocated population provided information regarding the ecology of lake sturgeon.

Addressing knowledge gaps to enhance the protection, conservation, and recovery is a major goal of the provincial Lake Sturgeon Recovery Strategy (Golder Associates Ltd.,

2011). Movement monitoring yielded important information regarding staging, spawning, and overwintering areas for the translocated population. Improving knowledge on lake sturgeon will aid in the conservation and management of the species, including the selection of ideal release sites for reintroduction.

As the main objective of most reintroductions is to establish a self-sustaining population (Griffith et al., 1989; Fischer and Lindenmayer, 2000), it is important to monitor demographics in reintroduction efforts to confirm success. In my thesis, juvenile captures confirmed reproductive success, and combined with ageing data, provided evidence of multiple reproductive events. As well as documenting multiple juvenile yearclasses within the reintroduced population a decade after the translocation event, the sampling design (targeted sampling followed by stratified randomized recapture effort) allowed for a population estimate. Although the confidence intervals were large due to the small number of recaptures, the estimate showed evidence of a population increase.

Study shortcomings and future directions

Despite the successful translocation and re-establishment of lake sturgeon in this river segment, this case study suffered from several shortcomings. The study was

128 necessarily limited to being a post-hoc analysis of potential establishment by the transferred adults, as the goal of the original translocation was simply to re-establish a lake sturgeon population in this fragment of the Mattagami River, and was not framed as an experiment (C. Hendry, OMNRF, pers. comm.). The information value of the original translocation would have been strengthened by conducting pre-introduction sampling to confirm that no remnant individuals were present within the reach. Assessment of similar future translocation or reintroduction efforts would be strengthened by surveying the area prior to release to confirm that the species of interest is not present. Another shortcoming of this study was that released individuals were only tagged externally. Tag loss may confound individual identification, especially in cases such as this where sampling was conducted a decade after release. Internal and external tagging would facilitate individual identification in follow up monitoring. Additionally, robust sampling of the translocated individuals prior to release would be valuable, including length, weight, gender, age, and genetics. Genetic information on all released individuals would aid future assessments of survival, reproduction, and recruitment (Allendorf and Luikart, 2007), particularly in light of analytical methods that enable genetic recognition or reconstruction of kin groups

(Chapter 3). Here, the genetic analysis was necessarily limited to adults that were recaptured a decade after release, as well as captured juveniles. The Chapter 2 telemetry monitoring was limited to adult lake sturgeon. Although netting yielded high juvenile catches in deep holes, the habitat used by early life stages is unknown and is a knowledge gap outlined in the provincial Lake Sturgeon Recovery Strategy (Golder Associates Ltd.,

2011). Additionally, this study affirmed that the timing of sampling after release should

129 be based on species biology to maximize the probability of detection success and information return on sampling efforts.

Translocation success has historically been low (Griffith et al., 1989; Germano and Bishop, 2009); selecting appropriate source populations or releasing multiple groups generally, from a single source may increase success rates. Source populations should be adapted to the current environmental conditions of the release site (Houde et al., 2015).

Using multiple releases have the potential to reduce post-release movement and genetic diversity loss (IUCN/SSC, 2013; Weiser et al., 2013). Multiple releases can increase also the number of individuals in the translocated population, enhancing the probability of translocation success, while minimizing impacts on the source population (Griffith et al.,

1989; Armstrong and Seddon, 2008; IUCN/SSC, 2013). The presence of conspecifics can increase fidelity of translocated individuals to the release site (Dolev et al., 2002; Mihoub et al., 2011; Richardson and Ewen, 2016), minimizing post-release dispersal.

Conspecifics adapted to the area are likely using quality habitat, so their presence guides released individuals to optimal habitat (Serrano et al., 2003; Harrison et al., 2009).

Alternatively, streamside rearing is proving effective for increasing fidelity of released lake sturgeon (Crossman, 2008; Welsh et al., 2010).

Multiple releases can also increase the genetic diversity of translocated populations (Tracy et al., 2011; Weiser et al., 2013). For instance, the genetic stocking guidelines for lake sturgeon in the Great Lakes suggests releasing 250 females and 250-

1250 males of diverse ages over an extended time period to maximize the effective population size of the translocated population (Welsh et al., 2010). Although these numbers are not feasible in smaller reaches such as the 35 km fragment of the Mattagami,

130 incorporating multiple releases into reintroduction plans may improve success. It should be noted that although multiple releases have potential advantages, releasing individuals from multiple sources should be avoided unless no source populations are adapted to the environmental conditions in the release site (Houde et al., 2015), as using multiple sources may potentially result in outbreeding depression from interbreeding between genetically divergent populations (Huff et al., 2011; IUCN/SSC, 2013; Cochran-

Beiderman et al., 2014).

As reintroduction biology gains more and more applications for conservation and mitigating human impacts on biodiversity, it is increasingly important that we identify and assess the relevant variables to maximize success. In recent years, the number and diversity of species reintroduced has increased vastly; initially reintroduction biology focused on restoring endangered species, but has expanded into a wide variety of conservation translocations, each with their own research needs (Jachowski et al., 2016).

Translocations and reintroductions have potential applications in many novel and controversial conservation tools including assisted colonisations to mitigate climate change, ecological replacements to restore functions in impacted ecosystems, rewilding ecosystems with wide-ranging, large animals, and resurrecting species through de- extinction (Jachowski et al., 2016). For example, assisted colonisations move species from native areas into areas with suitable climatic conditions, with the goal of mitigating climate change (Aitken and Whitlock, 2013). Post-release monitoring is important for assessing the failure or success of introductions, managing translocated populations, and informing future efforts.

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Appendix Table A3.1: Reintroduced Mattagami, Ontario lake sturgeon catches by year. Sampling Year Captures Recaptures 2011 32 0 2013 15 0 2015 114 27 2016 8 3 Overall 169 30

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Table A3.2: The 15 loci genotyped for lake sturgeon from the reintroduced Mattagami, Ontario population (n=141) sampled in 2011-2016, with associated genotyping error rates. Locus Error Rate AfuG122 0 AfuG195 0 AfuG63 0 AfuG67 0 AfuG68 0 AfuG74 0 AfuG160 0 AfuG204 0 AfuG61 0.025 AfuG68b 0 AfuG71 0 AfuG112 0 AfuG9 0 AfuG56 0 Spl120 0 Mean 0.002

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Table A3.3: Allele frequencies for the reintroduced, downstream, source, and Groundhog River, Ontario lake sturgeon populations.

Locus Allele Reintroduced Source (43) Downstream Groundhog (121) (21) (54) AfuG195 161 0.318 0.395 0.381 0.426 AfuG195 165 0.682 0.605 0.619 0.574 AfuG63 127 0 0.013 0 0 AfuG63 131 0 0 0.050 0 AfuG63 135 0.269 0.244 0.475 0.540 AfuG63 139 0239 0.359 0.150 0.090 AfuG63 143 0.492 0.372 0.325 0.370 AfuG63 147 0 0.013 0 0 AfuG67 193 0.456 0.744 0.467 0.606 AfuG67 241 0.029 0.047 0.100 0.011 AfuG67 289 0.515 0.209 0.433 0.383 AfuG74 218 0.079 0.154 0.048 0.310 AfuG74 222 0.217 0.231 0.143 0.240 AfuG74 226 0.704 0.615 0.810 0.450 AfuG160 131 0 0.012 0 0 AfuG160 135 0.568 0.616 0.474 0.433 AfuG160 143 0.009 0 0 0 AfuG160 147 0.265 0.326 0.132 0.529 AfuG160 151 0.158 0.047 0.395 0.038 AfuG200 141 1 1 0.974 1 AfuG200 145 0 0 0026 0 AfuG61 196 0.496 0.476 0.452 0.314 AfuG61 200 0.496 0.524 0.524 0.686 AfuG61 204 0.008 0 0.024 0 AfuG68 153 0.004 0 0 0 AfuG68 157 0 0.034 0 0 AfuG68 161 0.009 0 0 0 AfuG68 165 0.056 0.103 0.190 0.177 AfuG68 173 0.116 0.103 0.048 0.260 AfuG68 177 0 0.017 0 0.125 AfuG68 181 0.168 0.345 0.286 0.250 AfuG68 189 0.629 0.310 0.452 0.188 AfuG68 193 0.017 0.086 0.024 0 AfuG71 231 0.267 0.186 0.250 0.115 AfuG71 235 0.733 0.814 0.750 0.783 AfuG112 231 0 0.091 0 0 AfuG112 235 0 0 0.024 0 AfuG112 251 0.175 0.182 0.190 0.125 AfuG112 259 0.633 0.500 0.429 0.596

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AfuG112 263 0.192 0.212 0.357 0.230 AfuG112 267 0 0.015 0 0.002 AfuG9 124 0.144 0.071 0.167 0.112 AfuG9 128 0.059 0.024 0 0 AfuG9 140 0 0.060 0 0.367 AfuG9 144 0.458 0.595 0.524 0.408 AfuG9 148 0 0 0 0.020 AfuG9 152 0.314 0.250 0.214 0.092 AfuG9 156 0.025 0 0.095 0 AfuG56 262 0.068 0.114 0.143 0.030 AfuG56 266 0.712 0.629 0.595 0.770 AfuG56 274 0.220 0.257 0.262 0.200

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Table A3.4: Estimated allele frequencies accounting for potential null alleles for the 15 loci genotyped for lake sturgeon from the reintroduced Mattagami, Ontario population (n=141) sampled in 2011-2016. Adjusted allele Locus Allele frequencies AfuG122 NULL 0.4687 159 0.0036 163 0.3883 167 0.1282 171 0.0111 AfuG195 161 0.3619 AfuG195 165 0.6381 AfuG63 NULL 0.0284 AfuG63 131 0.0038 AfuG63 135 0.2919 AfuG63 139 0.2198 AfuG63 143 0.4561 AfuG67 193 0.4573 AfuG67 241 0.0385 AfuG67 289 0.5043 AfuG74 218 0.0745 222 0.2057 AfuG74 226 0.7199 AfuG160 NULL 0.0359 AfuG160 135 0.5371 AfuG160 143 0.0039 AfuG160 147 0.2383 AfuG160 151 0.1848 AfuG204 141 0.9979 AfuG204 145 0.0037 AfuG61 196 0.4493 AfuG61 200 0.5000 AfuG61 204 0.0107 AfuG68b NULL 0.0182 AfuG68b 153 0.0036 AfuG68b 161 0.0073 AfuG68b 165 0.0766 AfuG68b 173 0.1039 AfuG68b 181 0.1837 AfuG68b 189 0.5910 AfuG68b 193 0.0157 AfuG71 231 0.2647 AfuG71 235 0.7353

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AfuG112 235 0.0035 AfuG112 251 0.1773 AfuG112 259 0.6028 AfuG112 263 0.2163 AfuG9 124 0.1475 128 0.0504 AfuG9 144 0.4388 AfuG9 152 0.2986 AfuG9 156 0.0360 AfuG56 262 0.0791 AfuG9 266 0.6942 274 0.2266 Spl120 NULL 0.0506 Spl120fuG56 254 0.0545 Spl120f 258 0.0215 Spl120f 262 0.2544 Spl120f 266 0.0040 Spl120f 282 0.2413 Spl120f 286 0.3736

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Figure A3.1: Lake sturgeon pectoral fin rays from an adult from the translocated Mattagami River, Ontario population (left), and an adult from the nearby natural Groundhog River, Ontario population (right). Scale bar indicates 1000µm in each photograph.

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Figure A3.2: A power analysis of the number of observed and expected parental pairs based on potential genotyping error and mismatches between parent and offspring genotypes. Based on genotyped (15 loci) adults (n=13), juveniles (n=124), and larvae (n=4) from the reintroduced Mattagami River population sampled from 2011-2016.

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Figure A3.3: Length-weight relationship of unique juvenile lake sturgeon caught within the reintroduction segment of the Mattagami River (n=126).

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