Short-Term Effects of Lowhead Dam Removal on Emergent Aquatic Insect Communities in the Olentangy River, Ohio

THESIS

Presented in Partial Fulfillment of the Requirements for the Degree Master of Science in the Graduate School of The Ohio State University

Alexander Masheter

Graduate Program in Environment & Natural Resources

The Ohio State University

2018

Master's Examination Committee:

Dr. S. Mažeika P. Sullivan, Advisor

Dr. Lauren M. Pintor

Dr. Rachel S. Gabor

Copyrighted by

Alexander Masheter

2018

Abstract

Lowhead dams can significantly alter the geomorphology and ecology of rivers.

Removal of lowhead dams is becoming increasingly common, but the ecological impacts are not fully resolved. In this study, I investigated the short-term (9-21 months after dam removal) impact of a lowhead dam removal on emergent aquatic insect communities in an urbanized portion of the Olentangy River in Columbus, Ohio (USA). Seasonality was the strongest driver of emergent insect responses, in spite of dam-removal induced changes. Measures of diversity (Simpson’s Index) and community composition (% EPT) were most strongly affected by the dam removal. The interaction of time and season also emerged as important for some emergent insect responses including family evenness and richness. Water temperature, conductivity, and sediment grainsize distribution emerged as potential drivers of emergent-insect community shifts linked to dam removal. Overall,

I found limited evidence to support that dam removal had a strong effect on emergent aquatic insect communities within one year. Given the limited timeframe of this study, these results should be seen as preliminary. Ongoing research in the study system will provide additional insight into the effects of dam removal on emergent aquatic insects and other riverine communities and processes.

Acknowledgments

I would like to thank Danielle Vent, Danielle Cook, Alayna Doborek and Lars Meyer and other members of the Stream and River Ecology (STRIVE) Lab for their assistance in the field and laboratory. I would also like to thank Dr. David Manning for his statistical assistance. Finally, I would like to thank my committee members (Dr. Lauren Pintor and

Dr. Rachel Gabor) and my advisor, Dr. Mažeika Sullivan, for their expertise, patience, and help through this process. Funding support was provided by NSF DEB-1341215 (to

Dr. Mažeika Sullivan), the Ohio Department of Natural Resources, Division of Wildlife through the State Wildlife Grants Program and the Ohio Biodiversity Conservation

Partnership (to Dr. Mažeika Sullivan), the Ohio Water Development Authority (to Dr.

Mažeika Sullivan), and The Ohio State University (to Alex Masheter and Dr. Mažeika

Sullivan).

iii

Vita

June 2007 ...... Upper Arlington High School

June 2012 ...... B.S. Zoology, The Ohio State University

September 2012-Present ...... Graduate Research Assistant, The Ohio

State University

Fields of Study

Major Field: Environment and Natural Resources

Table of Contents

Abstract ...... ii

Acknowledgments...... iii

Vita ...... iv

List of Tables ...... vi

List of Figures ...... vii

Chapter 1: Background and Literature Review ...... 1

Literature Cited ...... 20

Chapter 2: Emergent Aquatic Insects Show Muted Responses to Lowhead Dam Removal

...... 22

References ...... 54

Appendix A: Summary statistics for emergent insect and water-quality variables by sampling period and reach, including mean and standard deviation (SD)...... 58

Appendix B: Relative abundance by study reach and year of emergent aquatic insects...61

List of Tables

Table 2.1 Summary statistics for emergent insect and water-quality variables by sampling period, including mimina (Min), medians, maxima (Max), mean, and standard deviation

(SD). Also see Appendix A for summary statistics by year and reach...... 32

Table 2.2 Results of paired t-tests comparing emergent aquatic insect metrics at OR1 and

OR3 before (June 2011-August 2012) and after (June 2013-June 2014) dam removal.

Statistical significance at p < 0.5 marked with *, indication of a trend at p < 0.10 marked with **...... 33

Table 2.3 Results of linear mixed-effects models for emergent aquatic insect family richness, density, Simpson’s Index, and evenness. Note that models include data from after dam removal only. Statistical significance at p < 0.5 marked with *, indication of a trend at p < 0.10 marked with **...... 34

List of Figures

Figure 1.1 Basic conceptual figures representing the presence of pulse and press disturbances over time from Lake (2000)...... 3

Figure 2.2 Map of the control reach (OR1), experimental reaches (OR2 – upstream unrestored, OR3 – upstream restored, OR4 - downstream), and previous 5th Avenue dam locations on the Olentangy River...... 27

Figure 2.3 Emergent insect metrics for experimental and control reaches including (a) mean emergent density, (b) mean family richness, (c) mean Simpson’s Index (1 – D), and

(d) mean evenness for Olentangy River reaches: Upstream control (OR1, black), upstream unrestored (OR2, red), upstream restored (OR3, green), and downstream (OR5, purple). Error bars are ± 1 SD. See Tables 2.2, 2.3 for statistical details...... 37

Figure 2.4 Percentage of Ephemeroptera, Plecoptera, and Trichoptera present in samples for experimental and control reaches for Olentangy River reaches: Upstream control

(OR1, black), upstream unrestored (OR2, dark gray), upstream restored (OR3, gray), and downstream (OR4, light gray). Error bars are ± 1 SD. See Tables 2.2, 2.3 for statistical details...... 39

Figure 2.5 NMS plots for family metrics grouped by (a) reach and (b) date. Years are represented by open circles (2013) and open triangles (2014). Reaches are indicated by different colors (black, gray, blue, and orange for OR1, OR2, OR3, and OR4 respectively). The stress value was 0.167 for the NMS ordination...... 40

vii

Figure 2.1 Simple linear regressions for relationships between emergent insect (a) density and temperature (R2 = 0.106, p = 0.031) (b) family richness and temperature (R2 = 0.103,

2 p = 0.032) (c) density and conductivity (R = 0.238, p = 0.001), and (d) evenness and D50

(R2 = 0.073, p = 0.084). Colored regions represent confidence curves at α = 0.05...... 41

viii

Chapter 1: Background and Literature Review

Introduction

Aquatic insects, throughout their life cycles, form a crucial component in linked river-riparian food webs. In larval form, insects convert basal resources such as leaf litter into secondary production through shredding (Gonzales and Graça 2003). Insect larvae are also key components of the diets of many fish, amphibians, and other aquatic macroinvertebrates (Klecka and Boudal 2012). In their adult form, aquatic insects that emerge from the stream and disperse into the terrestrial environment represent a critical energy source for riparian consumers, including riparian arthropods, bats, birds, and lizards (Baxter et al 2005, Fukui et al 2006, Wesner 2012, Kraus and Vonesh 2012). For riparian insectivores, the importance of this export of aquatic prey (hereafter “emergent aquatic insects”) is directly linked to the abundance, accessibility, and quality of aquatic prey subsidies (Baxter et al 2005). For example, studies in Japanese streams showed that depressed aquatic insect flux lowered the abundance of riparian spiders of the family

Tetragnathidae during months when emergence typically occurred (Kato 2003).

In recent years, the removal of lowhead dams has been advocated as a way to restore natural flow regimes and habitat conditions in impounded rivers. The removal of these dams, however, represents a pulse disturbance that can lead to appreciable

1 hydrogeomorphic changes associated with changes in streamflow dynamics (Cantelli et al

2004) and a mass release of sediments previously trapped behind the dam (Baldigo and

Smith 2012) leading to increased turbidity, remobilization of previously bound nutrients, modification of bottom composition, and mechanical damage from debris (Ahearn and

Dahlgren 2005, Cheng et al 2008). Collectively, these hydrogeomorphic changes can cause significant changes in aquatic invertebrate larval communities (Orr et al 2008,

Hansen and Hayes 2012).

The long-term ecological impacts of dam removal are still heavily debated

(Bushaw-Newton et al 2002, Doyle et al 2005, Ahearn and Dahlgreen 2005). Prior studies of river recovery following dam removal have indicated that aquatic invertebrate richness and diversity recovers within a few years, but may take decades to return to their former densities (Hansen and Hayes 2012). An in-depth study accounting for changes before and after the dam removal will greatly enhance the collective knowledge of how these restoration efforts affect the ecosystem. Additionally, by considering emergent insects, this work addresses the impacts of dam removal on integrated aquatic-riparian ecosystem function. The study analyzes the effects of lowhead dam removal on riparian flux by gauging the impact on emergent aquatic insect communities.

Pulse and Press Disturbances

Ecological disturbances have two major classifications, based on the duration of their existence. A pulse disturbance is a short-term, clearly delineated event, while a press disturbance has no clear ending, and tend to exert their influence over long periods of

2 time (Lake 2000, Figure 1.1). By this definition, the installation of a lowhead dam is a press disturbance, as it generates a long-term disruption in flow patterns and sediment deposition with no well-established date of removal. The removal of a dam, on the other hand, is a pulse disturbance, creating a flooding event that abates after all labor has been completed (Tullos et al 2014).

Figure 1.2 Basic conceptual figures representing the presence of pulse and press disturbances over time from Lake (2000).

Geomorphic Consequences of Dam Removal

There are currently over 87,000 dams in the continental United States; over

43,000 of them are considered lowhead, or run-of-river dams (i.e. < 7.5 m in height,

Dorobek et al 2015). Lowhead dam removal has been shown to be a significant pulse disturbance with ecological consequences at both short-term (one to five years) and long- term (30 years) temporal scales (Doyle et al 2005).

Multiple ecosystem shifts following dam removals have been documented (Hart et al 2002). Due to the disruption of natural flow patterns, sediments slowly accumulate behind dams. This tends to occur more rapidly in impoundments downstream of 3 agricultural areas (Doyle et al 2005). When the dams are breached, the pulse disturbance caused by the sudden increase in water flow also can disturb the sediments and resuspend them in the water column (Hart et al 2002). This, in turn, causes a pulsed increase in turbidity. In addition, the rapid decrease in water levels above the removed dam can expose previously anoxic sediment to the air, which results in a pulse increase in biological activity as previously unavailable nutrients can be cycled by aerobic processes

(Ahearn and Dahlgren 2005). The sudden increase in river flow can also lead to changes in erosion patterns, both immediately upstream and downstream of the dam (Doyle et al

2005). Following the initial effects of the dam removal, occurring in the first few years, press disturbance is expected to continue, due to the channelization effect caused by sudden flows on the downstream reaches, and biotic adjustments to the restored flow regime. In a stream system with a homogenous bed, channel incision will occur at a sharp change in channel slope, known as a knickpoint, which moves back as it erodes, while the channel widens downstream (Gartner et al 2015). Sometimes, the adjustments to geomorphology can be extremely drastic, with one flow event documented in Wales increasing channel width by 10 m (Thomas et al 2015). However, recent studies have promoted different models. One emphasizes the effects of preexisting bed conditions on the spatial extent and intensity of the erosion that occurs (Doyle et al 2005). As an example, an analysis of the regular construction and removal of a recreational run-of-the- river dam showed that sand-bed rivers recover rapidly from dam removal, showing a return to pre-removal geomorphic attributes in as little as seven months (Costigan et al

2016). Another model claims the most important variables for geomorphology are hydrology and the establishment of stabilizing vegetation (Cannatelli and Curran 2012).

Ecological Consequences of Dam Removal

Ecological consequences of dam removal still remain largely unresolved. Studies investigating these ecological effects have mostly focused on freshwater fish and mussels

(Gangloff 2013). Most studies indicate that dam removals have a variety of effects on mussel beds, while other studies indicate ambiguity in the effectiveness of dam removals on restoring habitat connectivity for native fish populations (Gangloff 2013). Studies that have focused on macroinvertebrates have been equally equivocal, with a large variety of results indicating that ecological effects may be determined by multiple factors, including local geomorphology and the resilience of the native macroinvertebrate species to sedimentation (Renöfält et al 2013). Some hypotheses also suggest that the rapid disturbance of accumulated sediments might also mobilize previously buried toxins.

However, a study on the Pawtuxet River during the removal of a lowhead dam measured no significant fluctuations in the levels of polycyclic aromatic hydrocarbons or polychlorinated biphenyls (Cantwell et al 2014). Additionally, a study within the

Olentangy River has indicated that dam removal can actually leads to a reduction in fish- community contamination, due to alterations in food-web patterns (Davis et al 2017).

Dam removal also triggers immediate downstream decreases in the density of benthic organisms, due to scouring and burial (Chiu et al 2013).

Some studies have suggested that small dams are actually beneficial to mussels, extending the growing season and providing an increase in suspended food particles, leading some researchers to advocate maintaining lowhead dams to assist imperiled mussel populations (Singer and Gangloff 2011). Dam removal, on the other hand, tends to greatly imperil mussel beds (Doyle et al 2005). A study in Koshkonong Creek,

Wisconsin, on unionid mussels demonstrated extremely high mortality rates, due to exposure and desiccation of populations in the former reservoir, and heavy burial in sediment downstream, leading to the complete extirpation of one rare species (Quadrula pustulosa) in the area (Sethi et al 2004). However, other studies have indicated that impounded mussel beds suffer from decreased population densities and extant species richness, along with limiting the upstream distribution of certain species (Tiemann et al

2007, Shea et al 2013).

The impact of dam removal on fish species is also variable. Dams and other forms of flow regulation tend to cause habitat fragmentation, leading to populations becoming divided, and frequently leading to declines and occasional extirpation of some species

(Perkin et al 2015). An intact dam generates selection pressures that favor tolerant, warmwater and lentic species, while cold and cool-water lotic species experience steep decline

(Cooper et al 2016). For highly mobile species, especially anadromous fishes, removing dams helps to restore historical levels of habitat connectivity (DeHaan et al 2011, Hogg et al 2013). A restoration of anadromous movement also restores transport of nutrients from the ocean into riparian ecosystems, with marine nutrients being found in piscivorous birds within a year of dam removal (Tonra et al 2015). Even less mobile species of fish

6 have shown increased habitat utilization following dam removals (Burroughs et al 2010).

Increased species homogenization in populations formerly restrained by artificial barriers is frequent, increasing total biomass available to the ecosystem (Kornis et al 2014).

Evidence has also been found indicating that the formation of riffles following dam removal influences fish assemblages (Cook and Sullivan 2018). However, other studies have indicated fish assemblages only making slight shifts towards those in free-flowing rivers, and occasionally showing a decrease in assemblage richness and density in the first three years (Maloney et al 2008). In some cases, dam removal may aid in the dispersal of exotic fish species and diseases, causing a negative impact on upstream fish communities (McLaughin et al 2013). Additionally, the removal of a dam can eliminate microhabitats generated by changes in water quality and geomorphic features, causing a reduction in species diversity (Kornis et al 2014). Full restoration of fish assemblages may take over three years, especially in areas where hydrological conditions discourage the formation of habitats for lotic species, or those species which preferentially inhabit cold pools (Poulos et al 2014).

Aquatic larval invertebrates might be expected to recover quickly after dam removal, due to their rapid life cycles and tendency to respond quickly to ecological change (Barbour et al 1999). However, there are also indications that population densities may remain depressed (Hansen and Hayes 2012). In some cases, taxonomic richness did not recover, and in fact continued to decline after the dam was removed (Renöfält et al

2013). In spite of these findings, a large-scale study on two dams in the Pacific Northwest determined that the presence of a lowhead dam generated a greater ecological disturbance

7 than the pulse disturbance generated by unregulated flow and sediment release associated with a dam removal. Once again, ecological recovery from the press disturbance occurred within the year, despite signs of continuing geomorphic instability (Tullos et al 2014).

Additionally, evidence has been found to suggest that seasonality plays a strong role in the response and recovery of aquatic macroinvertebrate communities following dam removal (Sullivan and Manning 2017, Cook and Sullivan 2018).

Emergent Aquatic Insects – Connecting Aquatic and Terrestrial Ecosystems

Emergent aquatic insects, that is, insects that mature in an aquatic environment, and then emerge from the water as adults, are important for riparian ecosystem processes

(e.g., Kautza and Sullivan 2016). For example, emergent insects capable of flight functionally integrate aquatic and terrestrial environments (Baxter et al 2005, Sullivan and Rodewald 2012). This transport of biomass, nutrients, and potential contaminants is also a major feature of the constant flux between the aquatic environment and the surrounding riparian zone. For example, the bulk of the diet of several bat species consists of emergent aquatic insects such as mosquitoes, and this aquatic subsidy can amplify terrestrial primary production as it is deposited in the form of guano (Fukui et al

2006). Riparian swallows also rely heavily on subsidies from emergent aquatic insects, especially in urban environments (Alberts et al 2013). Emergent aquatic insects are also major components in the diets of numerous riparian spiders, including those in family

Tetragnathidae. Riparian tetragnathids can feed almost exclusively on emergent aquatic insects (Kato et al 2004). For instance, Kato et al (2003) found that in riparian

8 tetragnathids, 50-90% of the dry mass of their diet was aquatic in origin. The reliance of tetragnathids on emergent aquatic insects varies with the environment and rates of disturbance (Krell et al 2014, Tagwireyi and Sullivan 2016). These and other riparian arthropods also serve as contributors to the input of organic matter from the terrestrial system to the river, leading to a continuous state of flux (Baxter et al 2005). Nutrient flux is heavily affected by disturbance and land use. Agricultural land use has a tendency to increase the raw biomass of emergent aquatic insects, while also modifying the community to favor small bodied-insects with consequences for terrestrial consumers.

For example, agricultural streams may favor consumers that rely on high volumes of small insects, while undisturbed forests favor greater numbers of large-bodied emergent aquatic insects and their predators, such as birds (Stenroth et al 2015).

The Olentangy River

The Olentangy River is a 156-kilometer tributary of the Scioto River flowing through Richland, Crawford, Marion, Morrow, Delaware, and Franklin counties in the state of Ohio, with a drainage area of 1406 square kilometers (Ohio EPA 2010, US

Geological Survey 2011). The river’s headwaters are located east of the city of Galion, and joins with the Scioto River within the city limits of Columbus. The section of the river featured in the study extends from the Franklin county line to the Olentangy’s confluence with the Scioto River, and is marked by heavy urbanization. The riverbed is composed heavily of alluvial silt and glacial sand and gravel. The course of the river is fairly unmodified over much of its length, developing with normal sinuosity and in-

9 stream complexity, with a variety of riffles, runs, and pools. In two key areas, however, the course of the river has been highly modified. In the 1960s in the vicinity of

Worthington, a substantial portion of the river was diverted 304 meters east, straightened, and lined with rip-rap, with five lowhead dams installed for aesthetic purposes.

Additionally, several lowhead dams were erected in the vicinity of the Ohio State

University for the purpose of providing a power plant with water for steam (Friends of the Lower Olentangy River Watershed 2003, Army Corp of Engineers 2004). In recent years, however, there has been a large-scale effort to remove unnecessary dams from the

Olentangy River. Between 2004 and 2010, six lowhead dams in or around the city of

Delaware, OH were successfully removed, with EPA testing indicating improvements with biotic indicators such as the Index of Biotic Integrity (IBI) and invertebrate community index (ICI) scores in four monitoring sites (EPA 2010, EPA 2010, EPA

2013)

The 5th Avenue Dam was constructed in 1935 to provide cooling water for a power plant on the Ohio State University Campus. After the plant became more efficient, and new environmental laws were passed, the dam fell out of use, with some modifications being made in the 1960s. An inspection of the dam in 1999 found some areas of deteriorated concrete, but no physical hazards or indications of impending dam failure (Army Corp of Engineers 2004). During an Ohio EPA inspection of the

Olentangy River, the region upstream of the 5th Avenue Dam was found not to meet ecoregional biocriteria for modified warmwater habitats, with highly tolerant macroinvertebrate communities, contaminated sediments, and pollution due to combined

10 sewer outflows (Ohio EPA 1999). Following a mussel survey in 2010, which determined that no threatened or endangered species would be imperiled by the project, the dam was breeched and destroyed in late summer of 2012, while the exposed bed of the dam pool was restored and revegetated with riparian vegetation in 2013 (Ohio EPA 2011).

Objectives

Although numerous studies have been conducted on how dam removals impact in-stream aquatic communities, few have examined their impact on emergent aquatic insects. Within this context, the primary goal of my project was to evaluate the short-term effects of lowhead dam removal on the flux of emergent aquatic insects to the riparian zone. This work builds on ongoing food-web research in the study system. An additional lowhead dam upstream from the removal site acted as control, allowing the study to follow a modified BACI (Before-After Control-Impact) design. My research objectives were to:

1. Quantify adult aquatic (i.e. emergent) macroinvertebrate communities at multiple

time steps following the dam removal.

2. Relate aquatic insect community characteristics (e.g. diversity, abundance) to

hydrogeomorphic and habitat characteristics associated with dam removal.

Although the study was conducted in the Olentangy River, I anticipate that the results will be broadly applicable, and could help inform management and restoration efforts related to lowhead dam removal.

References Cited

Ahearn, D.S. & Dahlgren, R.A. (2005) Sediment and nutrient dynamics following a low-

head dam removal at Murphy Creek, California. Limnology and Oceanography,

50, 1752-1762

Alberts, J.M., Sullivan, S.M.P., & Kautza, A. (2013) Riparian swallows as integrators of

landscape change in a multiuse river system: Implications for aquatic-to-terrestrial

transfers of contaminants. Science of the Total Environment, 463, 42-50

Baldigo, B.P. & Smith, A.J. (2012) Effects of flow releases on macroinvertebrate

assemblages in the Indian and Hudson Rivers in the Adirondack Mountains of

Northern New York. River Research and Applications, 28, 858-871

Barbour, M.T., Gerritsen, J., Snyder, B.D., & Stribling, J.B. (1999) Rapid bioassessment

protocols for use in streams and wadeable rivers: periphyton, benthic

microinvertebrates and fish. 2nd ed. United States Environmental Protection

Agency, Washington, DC

Bascietto, J., Hinckley, D, Plafkin, J. & Slimak, M. (1990) Ecotoxicity and ecological

risk assessment – regulatory applications at EPA 1. Environmental Science and

Technology, 24, 10-15

Baxter, C.V., Fausch, K.D. & Saunders, W.C. (2005) Tangled Webs: reciprocal flows of

invertebrate prey link streams and riparian zones. Freshwater Biology, 50, 201-

220

Burroughs, B.A., Hayes, D.B., Klomp, K.D., Hansen, J.F., & Mistak, J. (2010) The

effects of the Stronach Dam removal on fish in the Pine River, Manistee County,

Michigan. Transactions of the American Fisheries Society, 139, 1595-1613

Bushnaw-Newton, K.L, Hart, D.D., Pizzuto, J.E., Thomson, J.R., Egan, J., Ashley, J.T.,

Johnson, T.E., Horwitz, R.J., Keeley, M., Lawrence, J., Charles, D., Gatenby, C.,

Kreeger, D.A., Nightengale, T., Thomas, R.L., & Velinsky, D.J. (2002) An

integrative approach towards understanding ecological responses to dam removal:

The Manatawny Creek study. Journal of the American Water Resources

Association, 38, 1581-1599

Cannatelli, K.M., & Curran, J.C. (2012) Importance of hydrology on channel evolution

following dam removal: Case study and conceptual model. Journal of Hydraulic

Engineering-ASCE, 138, 377-390

Cantelli, A., Paola, C., & Parker, G. (2004) Experiments on upstream-migrating erosional

narrowing and widening of an incisional channel caused by dam removal. Water

Resources Research, 40, doi:10.1029/2003WR002940

Cantwell, M.G., Perron, M.M., Sullivan, J.C., Katz, D.R., Burgess, R.M., & King, J.

(2014) Assessing organic contaminant fluxes from contaminated sediments

following dam removal in an urbanized river. Environmental Monitoring and

Assessment, 186, 4841-4855

Cheng, F, Granata, T & Nechvatal, M. (2008) Discharge and suspended sediment

transport during deconstruction of a low-head dam. Journal of Hydrological

Engineering, 134, 652-657

Chiu, M., Yeh, C., Sun, Y., & Kuo, M. (2013) Short-term effects of dam removal on

macroinvertebrates in a Taiwan stream. Aquatic Ecology, 47, 245-252

Cook, D.R., & Sullvan, S.M.P. (2018) Associations between riffle development and

aquatic biota following lowhead dam removal. Environmental Monitoring and

Assessment, 190, 339

Cooper, A.R., Infante, D.M., Wehrly, K.E., Wang, L., & Brendan, T.O. (2016)

Identifying indicators and quantifying large-scale effects of dams on fishes.

Ecological Indicators, 61, 646-657

Costigan, K.H., Ruffing, C.M., Perkins, J.S., & Daniels, M.D. (2016) Rapid response of a

sand-dominated river to installation and removal of a temporary run-of-the-river

dam. River Research and Applications, 32, 110-124

Davis, R.P., Sullivan, S.M.P., & Stefanik, K.C. (2017) Reductions in fish-community

contamination following lowhead dam removal linked more to shifts in food-web

structure than sediment pollution. Environmental Pollution, 231, 671-680

DeHaan, P.W., Brenkman, S.J., Adams, B., & Crain, P. (2011) Genetic population

structure of Olympic Peninsula bull trout populations and implications for Elwha

Dam removal. Northwest Science, 85, 463-475

Doborek, A., Sullivan, S.M.P., & Kautza, A. (2015) Short-term consequences of lowhead

dam removal for fish assemblages in an urban river system. River Systems, 21,

125-139

Doyle, M.W., Stanley, E.H., Orr, C.H., Selle, A.R., Sethi, S.A., & Harbor, J.A. (2005)

Stream ecosystem response to small dam removal: Lessons from the Heartland.

Geomorphology, 71, 227-244

Fukui, D., Murakami, M., Nakano, S., & Aoi, T. (2006) Effect of emergent aquatic

insects on bat foraging in a riparian forest. Journal of Animal Ecology, 75, 1252-

1258

Gangloff, M.M. (2013) Taxonomic and ecological tradeoffs associated with small dam

removals. Aquatic Conservation, 23, 475-480

Gartner, J.D., Magilligan, F.J., & Renshaw, C.E. (2015) Predicting the type, location and

magnitude of geomorphic responses to dam removal: Role of hydrologic and

geomorphic constraints. Geomorphology, 251, 20-30

Gonzales, J.M. & Graca, M.A.S. (2003) Conversion of leaf litter to secondary production

by a shredding caddis-fly. Freshwater Biology, 48, 1578-1592

Hansen, J.F. & Hayes, D.B. (2012) Long-term implications of dam removal for

macroinvertebrate communities in Michigan and Wisconsin Rivers, United States.

River Research and Applications, 28, 1540-1550

Hart, D.D., Johnson, T.E., Bushaw-Newton, K.L., Horwitz, R.J., Bednarek, A.T.,

Charles, D.F., Kreeger, D.A., & Velinsky, D.J. (2002) Dam removal: Challenges

and opportunities for ecological research and river restoration. Bioscience, 52,

669-681

Hogg, R., Coghlan, S.M. Jr., & Zydlewski, J. (2013) Anadromous sea lampreys

recolonize a Maine coastal river tributary after dam removal. Transactions of the

American Fisheries Society, 142, 1381-1394

Kato, C. Iwata, T., Nakano, S., & Kishi, D. (2003) Dynamics of aquatic insect flux

affects distribution of riparian web-building spiders. Oikos, 103, 113-120

Kato, C., Iwata, T., & Wada, E. (2004) Prey use by web-building spiders: stable isotope

analysis of trophic flow at a forest-stream ecotone. Ecological Research, 19, 633-

643

Kautza, A., & Sullivan, S.M.P. (2016) The energetic contributions of aquatic primary

producer to terrestrial food webs in a mid-size river system. Ecology, 97, 694-705

Klecka, J. & Boukal, D.S. (2012) Who eats whom in a pool? A comparative study of

prey selectivity by predatory aquatic insects. PLoS ONE, 7, e37741

Kornis, M.S., Weidel, B.C., Powers, S.M., Diebel, M.W., Cline, T.J., Fox, J.M., &

Kitchell, J.F. (2015) Fish community dynamics following dam removal in a

fragmented agricultural stream. Aquatic Sciences, 77, 465-480

Kraus, J.M. & Vonesh J.R. (2012) Fluxes of terrestrial and aquatic carbon by emergent

mosquitoes: a test of controls and implications for cross-ecosystem linkages.

Oecologia, 170, 1111-1122

Krell, B., Röder, N., Link, M., Gergs, R., Entling, M.H., & Schäfer, R.B. (2015) Aquatic

prey subsidies to riparian spiders in a stream with different land types.

Limnologica, 51, 1-7

Lake, P.S. (2000) Disturbance, patchiness, and diversity in streams. Journal of the North

American Benthological Society, 19, 573-592

Maloney, K.O., Dodd, H.R., Butler, S.E., & Wahl, D.H. (2008) Changes in

macroinvertebrate and fish assemblages in a medium-sized river following a

breach of a low-head dam. Freshwater Biology, 53, 1055-1068

McLaughin, R.L., Smyth, E.R.B., Castro-Santos, T., Jones, M.L., Koops, M.A., Pratt,

T.C., & Veles-Espino, L.A. (2013) Unintended consequences and trade-offs of

fish passage. Fish and Fisheries, 14, 580-604

Ohio EPA (1999) Biological and water quality study of the Olentangy River and selected

tributaries 1999 – Delaware and Franklin Counties. OEPA Technical Report

MAS/2000-12-6, State of Ohio Environmental Protection Agency, Division of

Surface Water, Columbus, Ohio

Ohio EPA (2011) Finding of no significant impact to all interested citizens,

organizations, and government agencies: Columbus 5th Avenue dam removal and

river restoration. Columbus, Ohio Environmental Protection Agency

Orr, C.H., Kroiss, S.J., Rogers, K.L., & Stanley, E.H. (2008) Downstream benthic

responses to small dam removal in a coldwater stream. River Research and

Applications, 24, 804-822

Perkin, J.S., Gido, K.B., Costigan, K.H., Daniels, M.D., & Johnson, E.R. (2015)

Fragmentation and drying ratchet down Great Plains stream fish diversity.

Aquatic Conervation: Marine and Freshwater Ecosystems, 25, 639-655

Poulos, H.M., Miller, K.E., Kraczkowski, M.L., Welchel, A.W., Heineman, R., &

Chernoff, B. (2014) Fish Assemblage Response to a Small Dam Removal in the

Eightmile River System, Conneticul, USA. Environmental Management, 54,

1090-1101

Renöfält, B.M., Legon, A.G.C., Jonsonn, M., & Nilsonn, C. (2013) Long-term taxon

specific responses of macroinvertebrates to dam removal in a mid-sized Swedish

stream. River Research and Applications, 29, 1082-1089

Sethi, S.A., Selle, A.R., Doyle, M.W., Stanley, E.H., & Kitchel, H.E. (2004) Response of

unionid mussels to dam removal in Koshkonong Creek, Wisconsin (USA).

Hydrobiologia, 525, 157-165

Shea, C.P., Peterson, J.T., Conroy, M.J., & Wisniewski, J.M. (2013) Evaluating the

influence of land use, drought, and reach isolation on the occurrence of freshwater

mussel species in the lower Flint River Basin, Georgia (USA). Freshwater

Biology, 58, 382-395

Singer, E.E., & Gangloff, M.M. (2011) Effects of a small dam on freshwater mussel

growth in an Alabama (U.S.A.) stream. Freshwater Biology, 56, 1904-1915

Stenroth, K., Polvi, L.E., Fältström, & Jonsson, M. (2015) Land-use effects on terrestrial

consumers through changed size structure of aquatic insects. Freshwater Biology,

60, 136-149

Sullivan, S.M.P., & Manning, D.W.P. (2017) Dam removal disturbance leads to

seasonally distinct taxonomic and functional shifts in macroinvertebrate

communities. PeerJ. 5:e3189; DOI 10.7717/peerj.3189.

Sullivan, S.M.P., & Rodewald, A.D. (2012) In a state of flux: The energetic pathways

that move contaminants from aquatic to terrestrial environments. Environmental

Toxicology and Chemistry, 31, 1175-1183

Tiemann, J.S., Dodd, H.R., Owens, N., & Walh, D.H. (2007) Effects of lowhead dams on

unionids in the Fox River, Illinois. Northeastern Naturalist, 14, 125-138

Tonra, C.M., Sager-Fradkin, K., Morley, S.A., Duda, J.J., & Marra, P.P. (2015) The rapid

return of marine-derived nutrients to a freshwater food web following dam

removal. Biological Conservation, 192, 130-134

Tullos, D.D., Finn, D.S., & Walter, C. (2014) Geomorphic and ecological disturbance

and recovery from two small dams and their removal. PLoS ONE 9(9): e108091.

doi:10.1371/journal.pone.0108091

Wesner, J.S. (2012) Predator diversity effects cascade across an ecosystem boundary.

Oikos, 121, 53-60

Chapter 2: Emergent Aquatic Insects Show Muted Responses to Lowhead

Dam Removal

Introduction

The installation of a lowhead dam can generate a press disturbance, which is an ongoing alteration to a region’s ecological processes with no clearly defined end point

(Lake 2000). Accumulation of sediment and alteration of streamflow patterns generate selective pressures favoring lentic species above the dam, while also selecting against lotic species and flooding the banks (Smith et al 2017). Below the dam, streamflow is drastically reduced, leading to drying events towards the banks and a narrower channel

(Tullos et al 2014). The removal of a dam, by contrast, has been considered a pulse disturbance (Tullos et al 2014, Doborek et al 2015), whereby the event altering the ecosystem has a definitive beginning and end (Lake 2000). When a lowhead dam is breached and removed, there is a sudden release of previously accumulated sediments and a localized flooding event as the dam’s reservoir empties, leading to mechanical scouring and sediment deposition downstream of the dam (Ahearn and Dahlgren 2005,

Cheng et al 2008). Upstream of the dam, the water level of the former reservoir rapidly drops, along with channel degradation due to the release of sediments, followed by further degradation and channel widening (Doyle et al 2001).

The removal of lowhead dams exerts significant impacts on the ecological state of rivers (Doyle et al 2005, Hansen and Hayes 2012, Baldigo and Smith 2012). Aside from the disruption caused by the heavy machinery used to breach and remove dams in a

22 controlled manner, removing a dam can result in releases of sediments downstream, an extreme pulse in productivity downstream, and changes in erosion patterns both upstream and downstream (Hart et al 2002, Ahearn and Dahlgreen 2005, Doyle et al 2005). This resulting habitat modification can lead to extensive disruptions in the population dynamics of fish, mussels, and macroinvertebrates, with mussels being strongly affected due to their nature as sessile filter-feeders (Sethi et al 2004, Gangloff 2013, Renöfält et al

2013). Studies on the Olentangy River have also indicated that benthic macroinvertebrates showed signs of recovery linked to seasonal changes after 15 months

(Sullivan and Manning 2017). An analysis of fish assemblages in the same study area revealed a sharp decrease in species diversity below the dam directly after the dam removal, followed by partial improvement, as well as colonization by ecologically sensitive species (Doborek et al 2015).

Insects that mature in aquatic environments and then emerge as adults (hereafter,

“emergent aquatic insects”) are important in both aquatic and riparian environments. In their larval form, emergent aquatic insects are sensitive to water pollution, making them valuable bioindicators (Barbour et al 1999). As larvae, the insects are also important prey for a variety of fish (Wesner 2010, Wesner 2013). Adult aquatic insects become a major driver of aquatic-to-terrestrial energy nutrient flux, and are integral to the diet of a large variety of terrestrial species, including spiders (Kato 2003, Kato 2004), aerial insectivorous bats and swallows (Murakami and Nakano 2002, Fukui et al 2006), and other terrestrial consumers (Kautza and Sullivan 2016). When these animals defecate, or the emerging aquatic insects die and decompose, they contribute their carbon and

23 nitrogen to terrestrial production, allowing for an increase in riparian vegetation (Fukui et al 2006, Bultman et al 2014).

2013). In spite of these findings, a large-scale study on two dams in the Pacific Northwest determined that the presence of a lowhead dam generates a greater ecological disturbance than the pulse disturbance generated by unregulated flow and sediment release associated with a dam removal. Once again, ecological recovery from the press disturbance occurred within the year, despite signs of continuing geomorphic instability (Tullos et al 2014).

Within the Olentangy River system, analysis of benthic macroinvertebrates indicated recovery in both density and diversity in three years following the removal of the dam, with seasonal changes in macroinvertebrate communities becoming less pronounced

(Sullivan and Manning 2017).

In this study, I compared emerging aquatic insect communities before and after lowhead dam removal at upstream/downstream and restored (i.e., active channel engineering vs. unrestored reaches in the Olentangy River (Columbus, Ohio, USA). I hypothesized that dam removal would cause a rapid increase in emerging aquatic insect community diversity in the reach directly upstream of the removed dam, as measured by family richness, Shannon’s Index, and Simpson’s Index, due to dewatering of the

24 reservoir and return of a consistent streamflow velocity. Second, fine sediment deposition downstream of the previous dam would cause a decrease in the density and diversity of emergent aquatic insects in the reach downstream of the lowhead dam, due to smothering and burial of complex habitat structure. Finally, the disturbance generated by active channel engineering would cause a short-term decrease in density and biodiversity at an active restoration site. I explored the influences of water chemistry and hydrogeomorphic

(sediment grainsize distribution and streamflow velocity) characteristics as potential mechanisms linked to changes in emergent insect communities.

Methods

Study Site and Experimental Approach

The Olentangy River is a 156-km-long tributary of the Scioto River. Its watershed covers 1406 km2. The lower Olentangy flows through the city of Columbus, where it enters the Scioto River. Our 300-m study reaches were distributed both upstream and downstream of the 5th Avenue Dam, which was breached in August of 2012. The 5th

Avenue Dam was built in 1935 to provide cooling water for a now inactive power plant of The Ohio State University. Removal of the 5th Avenue Dam and channel restoration efforts were aimed at improving water quality and aquatic habitat (US Army Corps of

Engineers, 2004). In addition to the dam removal itself, restoration activities included channel engineering at sections of a 2.6-km river segment upstream of the previous dam, including reshaping the river channel, redeveloping and reconnecting floodplain

25 wetlands, and planting riparian vegetation (see Ohio EPA, 2011 for additional details).

Restoration activities were completed in September of 2014. Restoration efforts were focused on one segment of the previous impoundment, leaving an unrestored upstream reach (OR2; Figure 2.1). The upstream control reach, OR1 was located above an intact lowhead dam of comparable age and height to the 5th Avenue Dam. I assigned several treatments to our study reaches (e.g., upstream/downstream of dams, restored/unrestored sections, before/after) to assess the effects of dam removal and subsequent restoration activities on emergent insect responses. For clarity, all non-control reaches are designated as “experimental reaches”, following Dorobek et al (2015).

Figure 2.3 Map of the control reach (OR1), experimental reaches (OR2 – upstream unrestored, OR3 – upstream restored, OR4 - downstream), and previous 5th Avenue dam locations on the Olentangy River.

The study followed a modified BACI (before-after, control-impact) design

(Stewart-Oaten et al 1986, Downes et al 2002). Before data relative to emergent aquatic insects and water chemistry were available at a subset of study reaches and were collected as part of ongoing food-web research in the same river system (see Kautza and

Sullivan 2012, Alberts et al 2013, Tagwireyi and Sullivan 2015). The control sites represent an upstream (impounded) reach of an intact lowhead dam of comparable size

and age in the same river system. At each reach, three equidistant lateral transects

running across the river to the end of the immediate riparian zone were established).

Coordinated collections of after data (emergent aquatic insects, water chemistry, and

hydrogeomorphic parameters) were conducted along each transect at multiple intervals

from 9 to 21 months following dam removal: June 2013, August/September 2013,

November/December 2013, April 2014, and June 2014.

Emergent Aquatic Insects RO3RO2 Following Alberts et al (2013), six 1-m2 Mundie-style emerging aquatic insect

traps (Mundie 1964) were deployed along the transects (two per transect, one located

towards the left bank and one towards the right bank) for ten days at each of the four

reaches. At five and ten days, all emergent aquatic insects were collected, and

subsequently counted and sorted to family using Merritt et al (2008) as a guide.

Chemical Water Quality and Hydrogeomorphology

Water quality was measured at each transect, including temperature, conductivity,

dissolved oxygen, pH, ORP, flow, and turbidity. Water-quality measurements were made

once per sampling period. Temperature, conductivity, dissolved oxygen, pH, and ORP

were measured using a multiple parameter sonde (YSI 600 QS, YSI incorporated, Yellow

Springs, Ohio); streamflow velocity (m s-1) was determined using a portable flowmeter

(Flowmate 2000, Hach Flow, Loveland, Colorado), and turbidity (TPU) was measured

using a portable turbidity meter (Global Water Turbidometer Model WQ770-B, Xylem

Inc., College Station, Texas). Sediment size was determined using Wolman’s pebble count method (1954). Briefly, 200 haphazardly selected clasts at each reach were measured using a gravelometer and recorded by percentile (e.g., D50 = particle sizes for which 50% of the particles are finer).

Numerical and Statistical Analysis

Metrics utilized included family richness, Shannon-Weiner diversity index (H’), evenness (E), Simpson’s Index (D), and emergent insect density. The Shannon-Weiner index is an informational index in which both a greater number of species and a more even distribution contribute to greater H’

푠 ′ 퐻 = ∑ 푝푖 ln 푝푖 푖=1

Where pi is the proportion of the total represented by family i.

Family evenness (E) quantifies the relative abundance of families within the assemblage and ranges from 0 to 1 where communities with an equitability number closer to one represent greater evenness.

퐻′ 퐸 = 퐻푚푎푥

Where Hmax is the natural log of family richness.

Simpson’s Index (D) measures the probability that two individuals randomly selected from a sample will belong to the same family.

푛 퐷 = ∑( )2 푁

Where n is the total number of organisms of a particular family and N is the total number of organisms in the family. Due to the counterintuitive nature of Simpson’s Index,

Simpson’s Diversity Index (1-D) was used instead.

All data were transformed where necessary to meet assumptions of normality and homogeneity of variance. Non-metric multidimensional scaling (NMDS) was used to quantify changes in the aquatic emergent insect community composition among treatments. NMDS is an ordination technique that reduces multiple variables to a few graphically displayed interpretable dimensions (James and McCulloch 1990). Analysis of similarity (ANOSIM) was used to compare before-after aquatic insect community compositions (Somerfield et al 2002).

Linear mixed-effects models (LMMs) were used to assess potential differences in emergent insect response variables among reaches (representing control, upstream/downstream, restored/unrestored treatments) and time for “after” dam removal data for all Olentangy River reaches. Time, reach, and a reach × time interaction were included as fixed effects with transects nested within study reaches included as a random effect. Because “before” data was only available for the upstream control (OR1) and upstream actively restored (OR3) reaches, all Olentangy River time steps and reaches were not possible in one model. Therefore, paired t-tests were used to evaluate potential differences in response variables between before and after (year 1-3) dam removal at

OR1 and OR3. Post-hoc simple linear regressions were used to test for relationships between emergent-insect metrics (i.e., density, diversity measures) and water-quality and hydrogeomorphic data to explore potential environmental drivers. Due to the absence of

30 emergent insect activity from late fall to early spring, only data from the leaf-out period was utilized in the statistical analysis (June, August, and September). JMP 11.0 (SAS

Institute, Cary, North Carolina) was used for LMMs, simple linear regressions, and paired t-tests. R and the ‘vegan’ package were utilized for NMS and ANOSIM tests. In all statistical tests, p < 0.05 was used to indicate statistical significance; p < 0.10 was used as an indication of a trend.

Results

Summary data for all predictor and response variables are found in Table 2.1 and

Appendix A. A total of 13,155 emergent insects were collected from the four study reaches, representing 48 families. For sampling times during leaf-out (spring and summer), mean density ranged from 10.3 ind. m-2 (OR3 – upstream restored) to 263.3

(OR1 - control) before dam removal and from 6.83 ind. m-2 (OR3) to 402.3 (OR1) following dam removal. Density after dam removal was not significantly different than before dam removal in the upstream control reach (OR1 - control: paired t-test, t = 0.44, df = 3, p = 0.692, Table 2.2). Density was also not significantly different pre- and post- dam removal in the restored reach (OR3: paired t-test, t = 0.08, df = 3, p = 0.941, Table

2.2). Time, reach, and the interaction of date × reach did not emerge as significant influences on density following dam removal (LMM, p > 0.10 for all three, Table 2.3 and

Figure 2.2a). Mean density initially experienced a spike at OR1 in June 2013, but returned to levels comparable with other reaches in the following months. At OR2

(upstream unrestored), on the other hand, mean density gradually declined (Figure 2.2a).

Table 2.1 Summary statistics for emergent insect and water-quality variables by sampling period, including mimina (Min), medians, maxima (Max), mean, and standard deviation

(SD). Also see Appendix A for summary statistics by year and reach.

Min Median Max Mean SD Biotic Variable Date Emergent Insects Family Richness Jun 2013 1 4 5 3.5 1.3 Aug 0 4 6 3.6 2.0 2013 Apr 2014 1 2 3 1.8 0.7 Jun 2014 0 4 8 3.7 2.7

Simpson’s (1 – D) Jun 2013 0 0.032 0.495 0.091 0.140 Aug 0 0.155 0.438 0.161 0.129 2013 Apr 2014 0 0.169 0.500 0.102 0.146 Jun 2014 0.062 0.149 0.332 0.167 0.089

Evenness Jun 2013 0 0.062 0.614 0.134 0.177 Aug 0 0.246 0.977 0.276 0.256 2013 Apr 2014 0 0.143 1.000 0.236 0.294 Jun 2014 0.143 0.199 0.497 0.266 0.121

Density (no. m-2) Jun 2013 23.0 151.0 608.0 186.6 173.0 Aug 0 87.2 279.5 100.9 87.2 2013 Apr 2014 0.5 12.2 59.0 18.1 19.5 Jun 2014 0 49.7 215.5 56.0 60.5

% EPT Jun 2013 0 0.002 0.065 0.014 0.023 Aug 0 0 0.073 0.012 0.028 2013 Apr 2014 0 0.023 1 0.146 0.303 Jun 2014 0 0.048 0.055 0.025 0.023 Water Chemistry Temperature (°C) Jun 2013 20.37 23.95 24.86 23.27 1.48 Aug 19.49 22.26 23.25 21.76 1.27 2013 Apr 2014 10.52 10.68 10.85 10.67 0.11 32

Jun 2014 24.09 24.38 25.09 pH Jun 2013 6.27 6.73 6.77 6.69 0.13 Aug 6.44 6.56 7.70 6.77 0.45 2013 Apr 2014 8.87 9.01 9.06 8.98 0.07 Jun 2014 8.22 8.30 8.41 8.31 0.06

Conductivity ([mS cm2]-1) Jun 2013 2.433 2.543 2.612 2.547 0.052 Aug 2.051 2.113 2.368 2.168 0.129 2013 Apr 2014 0.732 0.814 0.844 0.799 0.039 Jun 2014 0.470 0.485 0.499 0.485 0.010

ORP (mV) Jun 2013 59.5 80.1 88.7 77.4 9.9 Aug 60.2 79.7 89.6 77.8 8.6 2013 Apr 2014 98.2 117.0 128.0 115.1 9.2 Jun 2014 48.7 98.7 143.0 100.0 25.8

Turbidity (NTU) Jun 2013 22.4 32.4 219.5 52.3 54.7 Aug 7.0 13.3 112.9 23.8 29.5 2013 Apr 2014 - - - - - Jun 2014 46.1 51.2 59.9 51.8 5.0 Hydrogeomorphology

Streamflow velocity (m s-1) Jun 2013 0.02 0.23 0.72 0.32 0.27 Aug 0.01 0.46 0.99 0.43 0.29 2013 Apr 2014 - - - - - Jun 2014 0.38 0.71 1.90 0.95 0.52

D50 (mm) Jun 2013 1.9 1.9 46 7.7 13.0 Aug 1.9 1.9 28 9.4 10.5 2013 Apr 2014 1.9 4.0 29 9.3 9.6 Jun 2014 1.9 5.6 14 6.5 4.8

Table 2.2 Results of paired t-tests comparing emergent aquatic insect metrics at OR1 and

OR3 before (June 2011-August 2012) and after (June 2013-June 2014) dam removal.

Statistical significance at p < 0.5 marked with *, indication of a trend at p < 0.10 marked with **.

OR1 t df p Emergent Insect 0.4359 3 0.692 Density Family Richness -0.5170 3 0.641 1-D -0.9297 3 0.421 E -0.5921 3 0.596 %EPT -0.4919 3 0.657 OR3 Emergent Insect 0.0802 3 0.941 Density Family Richness 0.2789 3 0.798 1-D -0.3430 3 0.754 E Time 0.4561 Reach 3 Time*Reach0.679 %EPT F df p2.9963 F df p3 F df 0.029* p Density 2.25 3 0.102 1.02 3 0.396 1.04 9 0.428 Family 3.22 3 0.036* 0.93 3 0.435 1.94 9 0.082** Richness D 2.96 3 0.047* 3.94 3 0.017* 1.24 9 0.307 E 0.82 3 0.490 2.25 3 0.101 2.08 9 0.062**

For sampling times during leaf-out (spring and summer), mean family richness ranged from 2 (OR3) to 13 (OR1) before dam removal and from 1 (OR3 and OR5) to 8 34

(OR1) following dam removal. Family richness after dam removal was not significantly different than before dam removal in the upstream control reach (OR1: paired t-test, t = -

0.52, df = 3, p = 0.641, Table 2.2). Likewise, family richness after dam removal showed no significant difference from family richness before dam removal in the restored reach

(OR3, paired t-test, t = 0.28, df = 3, p = 0.798, Table 2.1). Time (LMM: p = 0.036, Table

2.3 and Figure 2.2b), but not reach (LMM: p > 0.10, Table 2.3 and Figure 2.2b), had a significant influence on family richness. The interaction for time × reach was not statistically significant, but showed evidence of a trend (LMM: p = 0.082, Table 2.3 and

Figure 2.2b). OR2 showed a steady rate of decline in mean family richness over the sampling period, while trends in family richness for OR1 and OR3 were closely aligned, with overlaps in confidence intervals indicating possible convergence (Figure 2.2b).

Simpson’s Index of diversity during leaf-out ranged from 0.01 (OR3) to 0.56

(OR1) before dam removal, and from 0 (indicating a monospecific sample, at OR1, OR3, and OR5) to 0.996 (OR5) after dam removal. Simpson’s Index measurements after dam removal were not significantly different from those before the dam removal for either the control reach (paired t-test, t = -0.93, df = 3, p = 0.421, Table 2.2) or the restored reach

(paired t-test, t = -0.34, df = 3, p = 0.754, Table 2.2). Both time (LMM: p = 0.05, Table

2.3 and Figure 2.2c) and reach (LMM: p = 0.020, Table 2.3 and Figure 2.2c), but not the interaction of time × reach (LMM: p > 0.10, Table 2.3 and Figure 2.2c) showed significant influences on Simpson’s Index. OR1 initially yielded low measurements in 1-

D, but diversity increased over the course of the year. OR2 displayed a spike in diversity

35 in August and April, while trends at OR3 and OR5 closely mirrored each other (Figure

2.2c).

Evenness during leaf-out ranged from 0.03 (OR3) to 0.67 (OR1) before dam removal, and from 0 (indicating a monospecific sample, at OR1, OR3, and OR5) to 1

(OR2) after dam removal. Evenness after the dam removal was not significantly different from evenness before dam removal for the control reach (OR1: paired t-test, t = -0.59, df

= 3, p = 0.596, Table 2.2). Similarly, evenness after dam removal was not significantly different from evenness before dam removal for the restored reach (OR3: paired t-test, t =

0.46, df = 3, p = 0.679, Table 2.2). There was no indication that time or reach had significant influence on evenness (LMM: p > 0.10, Table 2.3 and Figure 2.2d). However, the interaction of time × reach showed indications of a trend (LMM: p = 0.062, Table

2.3). OR1 experienced a steady increase in evenness over the study period, while OR2 experienced surges in evenness in August 2013 and April 2014 (Figure 2.2d).

a b

)

(ind. m

Emergent Emergent Density

c d

Figure 2.4 Emergent insect metrics for experimental and control reaches including (a)

mean emergent density, (b) mean family richness, (c) mean Simpson’s Index (1 – D), and

(d) mean evenness for Olentangy River reaches: Upstream control (OR1, black),

upstream unrestored (OR2, red), upstream restored (OR3, green), and downstream (OR5,

purple). Error bars are ± 1 SD. See Tables 2.2, 2.3 for statistical details.

Community Composition

The most common insect family collected both before and after dam removal was

Chironomidae, which comprised between 99.31 (OR1) and 52.38% (OR5) of insects collected at any given reach, respectively (see Appendix B). Before dam removal,

Ceratopogonidae and Cecidomyiidae were relatively common at OR3 (15.2 and 3.5%, respectively). Dolichopodidae were the second most common family found after dam removal, accounting for 1.3% of all the insects collected. Several families of insects were not found after the dam removal, including Lonchopteridae and Aleyrodidae.

The percentage of Ephemeroptera, Plecoptera, and Trichoptera (EPT) at OR3 increased significantly following dam removal (paired t-test, t= 2.99, df = 3, p = 0.029,

Table 2.2), whereas the percentage at OR1 did not show a significant increase (paired t- test, t = -0.49, df = 3, p = 0.657, Table 2.2). OR2 showed a spike in % EPT in April 2014, but the sampling periods before and after April found no EPT taxa at all (Figure 2.3).

Figure 2.5 Percentage of Ephemeroptera, Plecoptera, and Trichoptera present in samples for experimental and control reaches for Olentangy River reaches: Upstream control

(OR1, black), upstream unrestored (OR2, dark gray), upstream restored (OR3, gray), and downstream (OR4, light gray). Error bars are ± 1 SD. See Tables 2.2, 2.3 for statistical details.

Visual mapping of the NMS data illustrates these findings (Figure 2.5), showing that there were significant differences in the number of individuals in each family based on the year in which they were collected, but not based on the reach. Both linear and non- metric fit scores (R2) had values near 0.95, which indicates a high goodness of fit.

Analysis of similarity (ANOSIM) failed to reject the null hypothesis of equivalent or smaller differences between groups for reach (R = -0.078, p = 0.69), but detected a significant influence for year (R = 0.248, p = 0.02).

Figure 2.6 NMS plots for family metrics grouped by (a) reach and (b) date. Years are represented by open circles (2013) and open triangles (2014). Reaches are indicated by different colors (black, gray, blue, and orange for OR1, OR2, OR3, and OR4 respectively). The stress value was 0.167 for the NMS ordination.

Influences of Water Quality and Hydrogeomorphic Characteristics on Emergent Insects

Water temperature was positively related to emergent insect density (simple linear regression, p = 0.031, Figure 2.5a) and family richness (simple linear regression, p =

0.032, Figure 2.5b). However, there was no association between water temperature and evenness or Simpson’s Index (simple linear regression, p > 0.10, Table 2.3). Conductivity was significantly (and positively) related to increased emergent insect density (p = 0.001, 40

Figure 2.5c), but not with family richness, Simpson’s Index, or evenness (simple linear

regression, p > 0.10 for all, Table 2.3, data not shown). Although ORP levels did not

show any significant influence on the emergent insect metrics, there was evidence of a

trend correlating ORP with Simpson’s Index and evenness (simple linear regression, p =

0.097 for both, Table 2.3). There were no associations between any emergent-insect

metrics and pH or turbidity (simple linear regression, p > 0.10 for all, Table 2.3, data not

shown). Median sediment size (D50) showed a trend of correlation to evenness (simple

linear regression, p = 0.084, Figure 2.5d) but not to family richness, density, or

Simpson’s Index (simple linear regression, p > 0.10, Table 2.3, data not shown).

a b

)

Family Richness

Emergent Density (ind. m

Temperature (°C) Temperature (°C) Figure 2.7 Simple linear regressions for relationships between emergent insect (a) density

and temperature (R2 = 0.106, p = 0.031) (b) family richness and temperature (R2 = 0.103,

2 p = 0.032) (c) density and conductivity (R = 0.238, p = 0.001), and (d) evenness and D50

(R2 = 0.073, p = 0.084). Colored regions represent confidence curves at α = 0.05.

)

2 c

Evenness

Emergent Density (ind. m

Conductivity ([mS cm2]-1) D50 (mm)

Discussion

In the Olentangy River, we found limited evidence to support that dam removal

had a strong effect on emergent insect communities from 9 to 22 months following dam

removal. Seasonality appeared to dominate emergent aquatic insect community

responses, in spite of dam removal, a finding similar to larval invertebrate community

responses in the same study system (Sullivan and Manning 2017). Measures of diversity

(Simpson’s Index) and community composition (% EPT) appeared to be most strongly

affected by dam removal. The interaction of time × reach also had a noticeable effect on

some emergent insect responses, such as evenness and family richness. One explanation

would be that the pulse and press disturbances generated by the dam removal and

subsequent restructuring of the riverbed did not significantly impact adult aquatic insect

community structure. This interpretation is supported by previous studies, which indicate

that smaller lowhead dams, such as the one removed at 5th Avenue, do not completely

42 block sediment and organic material from flowing downstream (Tullos et al 2014). As a result, the communities were still partially interconnected, and the dam removal had a relatively small impact on the water chemistry and flow of nutrients in the affected reaches, as well as larval drift.

Emergent-insect density was not significantly different among reaches or before or after dam removal. This is likely due to ease of dispersal in emergent aquatic insects.

Many emergent insects are especially efficient at dispersing to new river segments within a single life cycle (Sondermann et al 2015). Given that some emergent insects have life cycles less than a month in length (Jackson and Füreder 2006), rapid recovery in population density can be expected after a pulse disturbance. The lack of difference before or after dam removal is additionally backed up by studies on macroinvertebrate communities, which found a return to pre-removal levels tended to occur after 15-20 months, dependent on factors including mean annual discharge, surrounding land use, and distance from the removed dam (Carlson et al 2017).

Overall, my hypothesis related to increased diversity in the upstream reaches was not supported. However, Simpson’s Index, a measure of community dominance, appeared to be the most strongly affected by dam removal, with some divergence in the response at the upstream unrestored region in comparison to the others. This could be explained by the intermediate disturbance hypothesis, where the periodic occasion of disturbances is responsible for maintaining diversity in an ecosystem (Wilkinson 1999).

However, previous studies have indicated that the effects of disturbance on evenness was also dependent on taxonomic richness and Simpson’s index at the site in question

(Tonkin et al 2012). This would explain the influence of the reach on evenness, while data gathered before the dam removal indicates there may have already been a seasonal factor in fluctuations in emergent insect metrics.

In regards to community composition, I did not find strong evidence for community shifts via NMS analysis. However, I did find that the percentage of

Ephemeroptera, Trichoptera, and Plecoptera at OR3 (the restored reach) increased significantly following dam removal, in contrast to the control reach of OR1, which did not change. These three orders are strong indicators of a lotic habitat, and are also useful bioindicators, as a majority of them are intolerant to pollution, sedimentation, or other forms of water contamination (Barbour et al 1999, Lenat 1988, Lenat and Penrose 1996,

Wallace et al 1996). The increase in EPT density is also supported by Sullivan and Cook

(2018), who found a benthic community dominated by Chironomidae, Hydropsychidae, and Baetidae. This finding implies a possible short-term improvement via channel engineering in the restored reach. While the numerical increase in EPT was relatively small, there is a solid potential for the increase to build upon itself in time (i.e., serve as initial colonizers) as new habitats are exploited.

I also hypothesized that disturbance generated by active channel engineering would cause a short-term decrease in density and biodiversity at the active restoration site. This hypothesis was not supported by any of the results. A possible explanation could be based on the timing of the active channel engineering efforts, which focused primarily during the spring, when streamflow velocity was significantly elevated to the point where it was deemed too high of a risk to take measurements in the thalweg. As a

44 result, fine sediments became less likely to be deposited, and thus had a reduced impact on emergent insect communities (Wood and Armitage 1999, Mathers et al. 2017).

However, it should be noted that both the family richness and emergent density were relatively low throughout the study, and as noted above, taxonomic richness is a driver in determining the impact of disturbance in riverine systems (Tonkin et al 2012).

Potential Drivers

One of the likely drivers for the increase in sensitive families found in the previously impounded area is the exposure of coarser sediment and the formation of riffles, as suggested by Egan (2001). It should also be noted that the increase in the percentage of Ephemeroptera, Plecoptera and Trichoptera at OR3 (the upstream restored reach) was not reflected at OR4 (the downstream reach), in spite of the continued disturbance generated by active channel reengineering. This could be due to an increase in fine sediments at OR4 (downstream reach), which was noted by Cook and Sullivan

(2018). The lack of increase in EPT percentage in the Olentangy contradicts previous studies, which indicate that sensitive taxa downstream of a removed dam should have recovered within the timeframe of the study (Carlson et al 2017). This finding would support my second hypothesis, except there was no significant difference found in any of the metrics before or after dam removal at OR4 (downstream) Factors other than dam removal could also alter geomorphic changes after dam removal, especially land use

(Tullos et al 2014, Ahearn and Dahlgren 2005). Findings by other studies examining the

Olentangy dam removal indicate that there may not have been a significant sediment

45 pulse downstream of the removed dam, which could be confirmed with an additional quantitative geomorphic research study (Doborek et al 2015).

I also observed that water temperature and conductivity were significantly and positively correlated with emergent insect density. This could be explained partially through the morphoedafic index (MEI), which is normally a tool used in fisheries management to indirectly measure food sources (Oglesby 1982), but could arguably be used in other aquatic communities. There is also a clear divide in conductivity between

2013 and 2014, which might be tied to elevated flow during the spring of 2014, which could be viewed as an additional pulse disturbance.

Conclusions

Dams are a major footprint of human impact and global environmental change, and thus a clear target for attempts to reverse ecological modifications. Over 1,000 dams have been removed in the U.S. alone, with hundreds more likely removed worldwide

(O’Connor et al 2015), and that number is likely to increase as more dams become structurally or functionally obsolete, marking a substantial ecological disturbance in riverine systems across the globe. Additionally, most of the studies conducted on dam removal focused on smaller streams surrounded by agricultural or forested land

(Burroughs et al 2010, Bushaw-Newton et al 2002, Maloney et al 2008). This study provides needed insight on the effects of dam removal on larger streams in an urban setting. I found only minor changes in the density or diversity of emergent aquatic insects, both compared to before the dam removal and to unimpounded portions of the

46 river, while the presence of sensitive taxa at the actively restored reach increased significantly. These results are important to consider, as emergent insects provide an important energetic linkage with terrestrial systems through riparian flux, driven by consumers such as swallows, bats, and tetragnathid spiders (Kato 2003, Fukui et al 2006,

Kautza and Sullivan 2016). However, these results will need to be considered in conjunction with other studies, which indicate a high probability of short-term disruption in other communities, such as fish (Doborek et al 2015) and mussels (Hornbach et al

2014, Sethi et al 2004). Also, since nine months had passed between the deconstruction of the dam and the beginning of the study, it is possible that additional ecological events could have affected the community in the first months after dam removal without being monitored. Where possible, future studies into dam removal and its impacts on stream ecology should be initiated more quickly to capture immediate changes in ecological communities. As my study only captured a short timeframe, I view these results as preliminary; ongoing research in the study system will further clarify the effects of dam removal on emergent aquatic insects.

References