Benthic Community Structure Response to Flow Dynamics In

BENTHIC COMMUNITY STRUCTURE RESPONSE TO FLOW DYNAMICS IN

TROPICAL ISLAND AND TEMPERATE CONTINENTAL STREAMS

Dissertation

Submitted to

The College of Arts and Sciences of the

UNIVERSITY OF DAYTON

In Partial Fulfillment of the Requirements for

The Degree

Doctor of Philosophy in Biology

Kathleen R. Gorbach, M.S.

UNIVERSITY OF DAYTON

Dayton, Ohio

December, 2012

BENTHIC COMMUNITY STRUCTURE RESPONSE TO FLOW DYNAMICS IN

TROPICAL ISLAND AND TEMPERATE CONTINENTAL STREAMS

Name: Gorbach, Kathleen R.

APPROVED BY:

______Albert J. Burky, Ph.D. M. Eric Benbow, Ph.D. Faculty Advisor Faculty Advisor

______Karolyn Hansen, Ph.D. Mollie D. McIntosh, Ph.D. Graduate Committee Member Graduate Committee Member

______Mark Nielsen, Ph.D. P. Kelly Williams, Ph.D. Graduate Committee Member Graduate Committee Member

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Kathleen R. Gorbach

2012

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ABSTRACT

BENTHIC COMMUNITY STRUCTURE RESPONSE TO FLOW DYNAMICS IN

TROPICAL ISLAND AND TEMPERATE CONTINENTAL STREAMS

Name: Gorbach, Kathleen R. University of Dayton

Advisors: Drs. Albert J. Burky & M. Eric Benbow

Hydraulic characteristics in lotic ecosystems are influential in the structure and function of aquatic benthic communities. Human activities and the increased demand for freshwater have caused the modification of natural flow regimes worldwide.

Hydrological alterations, such as dams, diversions, and channelizations, are associated with ecological change and known to have detrimental effects on benthic communities.

As a whole, this dissertation investigated the effects of hydraulic variables on the spatial distribution of macroinvertebrates and habitat template characteristics in tropical and temperate freshwater streams of the West Maui Mountains, Maui, Hawaii, and in Dayton,

Ohio.

The first two studies took place in Hawaiian mountain streams that have been diverted, often removing >95% of base flow, for development, agriculture and tourism, thus modifying the natural flow and altering habitat and species composition. A transplant study investigated the effects of water removal and increased density on

! iv! dispersal and upstream migration of N. granosa. Initial mean upstream migration rate was 0.25, 0.66 and 1.16 m/d under reduced flow, natural flow and natural flow with increased snail density, respectively. Through calculations using rates from published studies of neritids migrating en masse or in long lines, we generated realistic time frames for N. granosa to migrate above diversions, ranging from 72 days to 2.5 years (aggregate) and 29 days to 1.1 years (long narrow line). By understanding upstream migration, recommendations for migratory pathway and population restoration can be applied globally for tropical amphidromous species.

Secondly, habitat template, discharge, habitat flow, and macroinvertebrate insect indices were evaluated within riffle and cascade microhabitats upstream and downstream of the highest elevation diversion in four streams of the West Maui Mountains. A significant 44% reduction in macroinvertebrate density downstream of diversions was found when streams and sites were pooled (p = 0.0009, df = 1, F = 11.49). Microhabitat had a significant effect on the ratio of native to introduced taxa densities, with the amphibious splash zone home to significantly more endemic taxa compared to riffles.

Non-native taxa were dominant (> 95% by density) and ubiquitous in riffle habitats. Our findings contribute to ongoing water management and restoration efforts focused on the conservation of native species and habitat integrity in tropical streams worldwide.

Finally, in the Little Miami River, Ohio, the physical template and macroinvertebrate community were compared between riffle and run habitats. Mean flow velocity and macroinvertebrate densities were significantly greater in riffle (Flow: mean ± SE = 0.74 ± 0.04 m/s; Density: 1892 ± 200.2) than run (Flow: 0.32 ± 0.01 m/s;

Density: 540.3 ± 76.8) habitats. Linear regression found a positive and significant

! v! relationship (y = 4097x – 115.1, p < 0.0001) where 49% of variation in macroinvertebrate density was explained by mid-column velocity. Our results call for the need of future analyses using simple and complex hydraulic variables to accurately predict the distribution of invertebrate communities.

In conclusion, comprehensive understanding of how flow variation affects stream ecosystems is necessary for the development of future management practices that promote balance between economic and environmental benefits.

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Dedicated to my parents, Fred and Anne Jennings, who have encouraged me to seize every opportunity, supporting me with excitement and

enthusiasm, and showering me with unconditional love.

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ACKNOWLEDGMENTS

I have had the pleasure of interacting with wonderful people throughout this doctoral graduate program, each impacting my life in a profound way. It’s amazing where life takes you, and I’m fortunate to have had this wonderful experience. Family, friends and colleagues have made it all possible, contributing immensely to this final product. Whether it was support at home, another warm body in the field, hours spent picking bugs at the microscope, or good laughs to ease the frustration; my relationships with others have made this achievable.

I am deeply grateful to my two advisors, Drs. Albert Burky and Eric Benbow.

They have shown me the world, taught me invaluable life lessons, dedicated hours and hours to me – talking, teaching and helping me better my critical thinking and writing skills, all the while developing a friendship that will always be true. You are two of the most outstanding and important men in my life. Thank you to my Graduate Advisory

Committee – Drs. Carl Friese, Karolyn Hansen, Mark Nielson, Mollie McIntosh, and

Kelly Williams. I appreciate the continued guidance and time they gave to developing me as a contributing scientist. I would also like to express thanks to the University of

Dayton Graduate School and Biology Department for their financial assistance, providing teaching and travel opportunities, and maintaining a successful graduate program.

Going through it together, I’d like to thank my fellow graduate students who provided much camaraderie – especially Casey Hanley, Andy Lewis, Rachel Barker, Jen

! viii! Lang, Tracy Collins, and Elizabeth Rhodes. A special and sincere thanks to my sidekick, partner in crime and friend through it all, Megan Shoda. Her positivity, intellectual insights, brute strength, and smiles made all the difference. The endless hours together – in the stream, at home, driving the stream mobile, and of course in the lab sitting around our microscopes, were more than enjoyable. Fortunate for me, she saved me from wild dogs and bioluminescent algae scam artists, and the GIS expertise she provided has always been top notch.

My fieldwork and the laboratory aftermath would not have been humanly possible without the help of numerous undergraduates. Thank you to Doug Vonderhaar, Tiffany

Blair, Jon White, Carolyn Teter, Maggie Ernst, Ryan Lemier, Ryan Andrews, Allison

Gansel, Ian Barron, Gustavo Diaz, Elise Grotehouse, Jack Farrely, Jessica Teater,

Charlotte Perko, John Kurzawa, Liz Grazdick, and Melanie Aldaharian. I am grateful for the time and effort they put into my work. Their extra hands and eyes, and more importantly their companionship, made the cold mornings in the stream, long sampling days, and crowded lab benches in 226, quite memorable. It was a pleasure working with each one of them and I wish them all a wonderful future.

Finally, my family has given me the strength to always keep going. Thank you to my husband, John, because while he doesn’t always ‘get it’, he loves and supports me everyday, gently pushing me to get it done! My mom, sister, two brothers and their families have been by my side every step of the way – showing interest in my passion, trusting in my travels, helping with Charlie and around the house so I could get work done, and loving me no matter what. I am truly grateful.

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TABLE OF CONTENTS

ABSTRACT………………………………………………………………………....iv

DEDICATION……………………………………………………………………... vii

ACKNOWLEDGMENTS…………………………………………………………. viii

LIST OF FIGURES………………………………………………………………. xiii

LIST OF TABLES………………………………………………………………... xv

CHAPTERS

I. INTRODUCTION & LITERATURE REVIEW……………………….1

Introduction………………………………………………………… 1

Literature Review………………………………………………….. 3

References………………………………………………………….. 15

II. DISPERSAL AND UPSTREAM MIGRATION OF AN AMPHIDROMOUS NERITID SNAIL: IMPLICATIONS FOR RESTORING MIGRATORY PATHWAYS IN TROPICAL STREAMS……………………………..20

Summary…………………………………………………………… 20

Introduction…………………………………………………………22

Methods……………………………………………………………. 25

Results……………………………………………………………… 31

Discussion ………………………………………………………….33

Acknowledgments ………………………………………………….39

Tables………………………………………………………………. 41

! "! Figures………………………………………………………………50

Appendices………………………………………………………….54

References………………………………………………………….. 63

III. VARIABILITY IN HABITAT TEMPLATE AND BENTHIC COMMUNITY RESPONSE TO ANTHROPOGENIC WATER REMOVAL IN TROPICAL MOUNTAIN STREAMS.……………………………………………...69

Abstract……………………………………………………………..69

Introduction ………………………………………………………...70

Methods …………………………………………………………….74

Results ……………………………………………………………...79

Discussion ………………………………………………………….85

Acknowledgments ………………………………………………….91

Tables………………………………………………………………. 92

Figures……………………………………………………………... 97

References………………………………………………………….. 106

IV. BENTHIC COMMUNITY STRUCTURE UNDER DIFFERENT FLOW AND SUBSTRATE CONDITIONS IN THE LITTLE MIAMI RIVER, OHIO…………………………………………………………… ……...112

Abstract……………………………………………………………..112

Introduction…………………………………………………………113

Methods……………………………………………………………..115

Results………………………………………………………………118

Discussion…………………………………………………………..121

Figures………………………………………………………………124

References………………………………………………………….. 131

! "#! FUTURE DIRECTIONS………………………………………..…………………. 135

! "##! !

LIST OF FIGURES

Note: Titles have been abbreviated.

CHAPTER II Figure 1: Map of the Hawaiian Islands with Maui and study streams identified ...... 50 Figure 2: Individual snail captures made during each treatment ! 16 days after release plotted on an x, y coordinate system (release point = 0,0)...... 51 Figure 3: Mean (SE) Euclidean migration rates (EMR; black bar) and Upstream migration rates (UMR; white bar) between treatment conditions...... 52 Figure 4: a) Mean upstream migration rate (UMR) over all initial (! 6 days) search days for each treatment. b) Mean (SE) upstream migration rates for RF, NF and NF+D treatments with captured snails that traveled ! 8m in ! 6 days, pooled longer-term captures made 16, 33 and 63 post-release, and ‘rapid’ snail captures that traveled " 8m in ! 6 days...... 53

CHAPTER III Figure 1: Map of West Maui, Hawaii, with study watersheds highlighted and corresponding study locations as black dots...... 97 Figure 2: Mean (SE) measured discharge in all streams, upstream and downstream of the highest elevation diversion...... 98 Figure 3: Mean relative percent of available habitat within the upstream and downstream 100 m study reaches...... 99 Figure 4: a) Mean (SE) riffle macroinvertebrate non-corrected density and b) mean (SE) habitat-corrected density for upstream and downstream reaches within each stream...... 100 Figure 5: Mean (SE) macroinvertebrate density between riffle habitat and amphibious and torrenticolous microhabitats of cascades upstream and downstream of diversions, with all streams pooled...... 101 Figure 6: Community composition of riffle, torrenticolous and amphibious habitats for all streams and sites pooled...... 102

! xiii! Figure 7: Mean (SE) macroinvertebrate density for habitat and location upstream and downstream of the highest elevation diversions of all streams (data pooled)...... 103 Figure 8: Index of Nativity (ratio of native taxa density to introduced taxa density) for each habitat and location upstream and downstream of the highest elevation diversion in each stream (data pooled)...... 104 Figure 9: NMDS ordination with habitat overlay...... 105

CHAPTER IV Figure 1: Hydrograph depicting mean daily discharge (m3/s) for the Little Miami River (USGS gage 03240000), near Oldtown, Ohio from May through September 2008...... 124 Figure 2: Riffle and run habitat substrate particle size frequency distribution...... 125 Figure 3: Mean (SE) flow velocity measured at each benthic sample in the riffle and run habitats...... 126 Figure 4: Mean (SE) macroinvertebrate density for habitat and sampling time...... 127 Figure 5: Linear regression relationship between mid-column velocity (m/s) and macroinvertebrate density...... 128 Figure 6: Community composition of riffle and run habitats for all sites pooled...... 129 Figure 7: NMDS ordination with habitat overlay...... 130

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LIST OF TABLES

Note: Titles have been abbreviated.

CHAPTER II

Table 1: Shell length, width and height of Neritina granosa collected from Honomanu Stream, Maui, Hawaii...... 41

Table 2: Habitat and hydraulic variables for each treatment during Neritina granosa migration in Iao Stream, Maui, Hawaii...... 42

Table 3: Three types of snail capture data...... 44

Table 4: Snail Upstream Migration Rate (UMR as m d-1), under historic discharge quantiles, Q50, Q70, Q90, and minimum discharge needed to reach the ocean ...... 46

Table 5: Migration as described in previous published studies compared to results of the current study...... 47

CHAPTER III

Table I: Two-way ANOVA statistics for physical habitat template characteristics and macroinvertebrate density and diversity for the riffle habitat...... 92

Table II: Mean (± SD) density of represented macroinvertebrate taxa in the riffle habitats among the study streams and between upstream and downstream of diversion...... 94

Table III: Two-way ANOVA statistics for the Index of Nativity for each microhabitat and site within each stream...... 96

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CHAPTER I

INTRODUCTION & LITERATURE REVIEW

INTRODUCTION

The importance of flow to the ecology of aquatic benthos is unquestionable in lotic ecosystems (Nowell & Jumars 1984). The integrity of these freshwater habitats depends on how various benthic species assemble into structural communities, provide ecosystem functions and contribute to complex food webs (Covich et al. 1999). Benthic organisms, also known as macroinvertebrates, include crustaceans, gastropods and aquatic larval and pupal forms of terrestrial insects that reside on and within the stream substrate. Through the use of macroinvertebrates, physical-biological coupling can aid our understanding of the dynamic organization of the ecological structure of streams and rivers.

The literature review and experimental work presented throughout the following chapters indicate the importance of riverine hydraulic variables in governing habitat template characteristics and spatial distribution of benthic organisms. Developing a more complete understanding of how spatial and temporal flow variation affects the structure and function of stream ecosystems will provide deeper insight into ecological organization, improve our ability to predict how flow alterations caused by human activities affect these vital ecosystems, and guide water management practices that would

! achieve a better balance between economic and environmental benefits (Hart & Finelli,

1999).

Although the dissertation work was completed in two different locations – tropical island and temperate continental freshwater streams – unifying ecological themes permeate throughout such as Community Ecology, Disturbance Ecology, Hydraulic

Stream Ecology, and the potential application of Conservation and Restoration Ecology.

The tropical island stream studies, presented in Chapters II and III, were carried out in four streams of the West Maui Mountains on Maui, Hawaii. They investigated the effects of anthropogenic stream flow removal on macroinvertebrate community composition and upstream migration rate of a native neritid snail. Our findings are of timely importance due to a recent legal battle regarding the return of water to the streams under study.

Experimental efforts in the temperate continental stream, the Little Miami River, in

Dayton, Ohio, presented in Chapter IV, examined the spatial distribution of macroinvertebrate communities between habitats of differing flow conditions and habitat template characteristics. The current conclusions and those gathered through the completion of future analyses can contribute substantially to fundamental knowledge of

Stream Ecology.

! LITERATURE REVIEW

Part 1: Review of flow effects on benthic organisms

The role of macroinvertebrates in stream ecosystems

Benthic invertebrates (i.e. macroinvertebrates) are abundant and typically diverse in freshwater sediments. While macroinvertebrates are ideal biological indicators of overall stream health and integrity (Rosenberg 1993), their functional importance generally goes unnoticed until unexpected changes occur in ecosystems (Covich et al.

1999). They play a fundamental role in the biological community and food web of aquatic ecosystems, serving as food for fish and facilitating ecosystem functions such as sediment mixing, nutrient cycling and energy flow. Each species is uniquely important, adapted to functioning under variable conditions and performing distinct ecosystem services; the addition or loss of a single species can dramatically alter food web dynamics

(Covich et al. 1999).

Effects of flow

Flow is recognized in running water ecology as the fundamental abiotic factor controlling ecological processes and patterns in stream ecosystems (Hart & Finelli 1999).

The natural flow regime is the range and variation of flows over recent historical time, and sets the template for contemporary ecological processes (Resh et al. 1988, Doyle et al. 2005), evolutionary adaptations (Lytle & Poff 2004), and native biodiversity maintenance (Bunn & Arthington 2002, Poff & Zimmerman 2010). Streams are known to provide patchy landscape where hydraulic and structural heterogeneity have been proposed to be major determinants of macroinvertebrate community organization.

! Small-scale differences in hydraulic conditions created by combinations of velocity, depth and substrate roughness play an important role in the spatial distribution of macroinvertebrates in riffle habitats (Brooks et al. 2005). As summarized by Hart &

Finelli (1999), flow affects the ecological processes of benthic organisms directly or indirectly through multiple causal pathways. Processes affected include dispersal, habitat use, resource acquisition, competition, and predator-prey interactions (Hart & Finelli

1999). Not only are organisms influenced individually, but also hydraulic characteristics affect entire assemblages (Statzner et al. 1988).

Many of the flow forces and processes affecting benthic organisms (i.e. drag, lift, diffusivity, and mass transfer) vary as a function of velocity (Denny 1993, Vogel 1994,

Hart & Finelli 1999) and thus, flow characteristics related to velocity are often considered as having the greatest relevance. The problem with “mean velocity” is that it is only one characteristic of moving water and can hardly be used to comprehensively explain the physical environment experienced by the organism (Statzner et al. 1988). Flow in a natural channel is three-dimensional which means that each fluid particle may travel in the upstream—downstream direction, from bank to bank and from bottom to surface

(Statzner et al. 1988).

Not only is flow multi-dimensional, but flow conditions experienced by benthic organisms differ from those experienced farther above the stream bed due to the presence of a velocity gradient, or the “boundary layer,” which is created by friction between the moving water and the stationary bed (Nowell & Jumars 1984, Statzner et al. 1988, Hart

& Finelli 1999). Many invertebrate taxa are constrained by these near-bed flow conditions, living within this layer where turbulent flows are highly irregular,

! unpredictable and uniquely dependent on the shape, size and arrangement of roughness elements (Dolédec et al. 2007). Unfortunately, such complex topography makes it extremely difficult and currently impossible to predict near-bed velocities using simple formulae such as the log-linear relationship between velocity and height above the bed

(Nowell & Jumars 1984, Hart & Finelli 1999).

Habitat preferences of benthic organisms

Hydraulic preferences of lotic invertebrates can be explained by tradeoffs between energy costs and oxygen consumption, biotic interactions and the ability to colonize habitats (Phillipson 1956, Edington 1968, Peckarsky et al. 1990, Doledec et al. 2007,

Fonseca & Hart 2001). While previous studies have shown that habitat preferences are influenced by flow conditions (Merigoux & Dolédec 2004), specific relationships have not been addressed that could be used to confidently predict habitat preferences (Benbow et al. 1997). Benbow et al. (1997) found a significant relationship between water micro- flow parameters in torrential benthic habitats where chironomid larval density was inversely correlated with depth and positively correlated with bottom velocity. In this study, the heterogeneity of the substrate and the near bottom water flow were found to interact to produce areas of flow refugia affecting species abundance, distribution, and habitat availability. A similar study of blackfly larvae determined that the current speeds measured 10 mm above the bed were poor predictors of speeds measured at 2 mm, limiting our understanding of the quantitative and qualitative importance of flow to benthic stream organisms (Hart et al. 1996). Further, blackfly larvae have been observed positioning their body below the boundary layer, while extending their filter feeding structures in the high velocity flow (Hart et al. 1996).

! Hydraulic preferences may also be identified and explained through morphological and behavioral observations of these organisms. Many benthic organisms lower the risk of dislodgement by reducing the drag force experienced with a streamlined body shape or a smaller size, which allows them to inhabit zones of greatly reduced velocities. They may also have specialized structures or modify their position with the flow to avoid dislodgement. In contrast, some benthic organisms actively expose themselves to flow for dispersal. This has been documented where the rate at which black fly larvae enter the water column decreases with increasing water velocity (Hart,

1999). It has also been shown that larval simuliids (Chance & Craig 1986), various pulmonate snails (Dussart 1987), dorsoventrally flattened insects and streamlined limpets

(Smith & Darnall 1980, Statzner & Holm 1982, McShaffrey & McCafferty 1987) negotiate rather complicated flow and consequently endure the forces of flows (Statzner et al. 1988).

Exploring these linkages between the organism and the abiotic environment, specifically hydraulic preferences, is required prior to developing predictive models regarding the structure and function of ecosystems. Such information will allow quantification of flow regimes necessary for maintaining biotic and abiotic processes within stream ecosystems (Dolédec 2007, Hart & Finelli 1999).

Part 2: Anthropogenic disturbances in stream ecosystems

Modification of the flow regime

Disturbance of river communities has been a central focus in the growth of lotic science, with much thought given to defining and quantifying disturbance in systems that are inherently dynamic (Resh et al. 1988, Poff 1992, Kinzie et al. 2006). Disturbance can 6!

! arise from many sources, but most attention has been devoted to disturbances resulting in changing the flow conditions of a river (Kinzie et al. 2006). These hydrological modifications include any anthropogenic disruption of flow in terms of magnitude, frequency, duration, timing or rate of change of timing; thus changing the natural flow regime (Standford et al. 1996, Hart & Finelli 1999, Rosenberg 2000, Poff & Zimmerman

2010) and impacting the physical, chemical and biological structure and function of rivers and streams (Ward & Standford 1983, Allan 1995).

Ecosystems worldwide are threatened by demands on freshwater resources due to increased population growth and consumption (Postel 1997, Benstead et al. 1999, March et al. 2003). To account for these demands, many streams and rivers have undergone hydrological alterations such as dams and surface-water diversions, stream channelization, and intercatchment water transfer (Rosenberg et al. 2000, Dudgeon 2000,

Baker et al. 2011). Alterations can occur simultaneously at different scales, such as landscape (watershed), stream reach and microhabitat. For example, urbanization (a landscape scale process) is typically accompanied by channelization and the removal of riparian canopy cover (a reach scale process), resulting in higher water temperatures, increased daily temperature fluctuations, increased siltation, and decreased substrate size

(microhabitat scale processes) (Brasher 2003).

In a review by Poff & Zimmerman (2010) of 165 studies that reported either aquatic or riparian responses to flow regime alteration, 92% of the studies found negative ecological changes in response to a variety of types of flow alteration. These modifications can include decreased flow and changes in the riffle – run sequence of streams, thus altering available habitats and species composition. Such consequences

! include changes in species distribution, abundance, composition and diversity of aquatic communities, recruitment failure and loss of native species, isolation of populations and local extinction, and the invasion and success of exotic and introduced species (Bunn &

Arthington 2002).

Knowledge of how organisms colonize and persist in a habitat after such disturbance can provide insight into the impact of human processes on natural systems

(Dolédec et al. 2007). Thus, a better understanding of physical-biological coupling in streams will enhance our ability to predict how flow alterations caused by various human activities affect these vital system, advance efforts to restore structure and function, and enable solutions to some of our most pressing environmental problems (Hart & Finelli

1999).

Stream diversions in the Hawaiian Islands

While freshwater ecosystems worldwide are threatened by increased demands on freshwater resources, tropical stream habitats are undergoing substantial alteration as human population increases and watersheds become far different from those that once sustained native stream communities (Brasher 2003). Streams throughout the tropics have been altered by water diversion, channel modification, introduced species and water quality degradation (Brasher 2003). The impact of water removal and subsequent changes within the aquatic ecosystem have been extensively studied in temperate regions; whereas, relatively little research has been conducted in tropical regions leaving these ecosystems and human impacts poorly understood (Benke et al. 1988, Flowers 1991,

Jackson & Sweeney 1995).

! This is especially true of tropical islands, such as Hawaii, where freshwater resources are naturally limited to surface streams and groundwater supplied by heavy tropical rains (Fitzsimons et al. 1997, Brasher 2003, March et al. 2003). To original

Hawaiian civilizations, water resources were culturally important and used in native practices of taro cultivation and habitat to native stream macrofauna, such as fish, shrimp, snails and prawns that were harvested as a food source. However, through Western colonization and associated development of large-scale commercial sugarcane plantations in the mid-1800s, stream diversions and extensive tunnel transport systems were built to translocate water from the wet, windward watersheds to the dry leeward areas of growth and development. At least 58% of the estimated 366 perennial streams had experienced some type of streamflow alteration by 1978, with water being exploited for anthropogenic uses, such as agriculture, development and tourism (Parrish et al. 1978,

HCPSU 1990).

Many of these diversions remove 90 – 100% of base flow volume, altering the natural flow regime which is important for sustaining native biodiversity and ecosystem integrity (Wilcox 1996, Poff et al. 1997, Benbow 1999, McIntosh et al. 2002, Brasher

2003, McIntosh et al. 2008). During major precipitation events, water may breech some diversions; however, the quantity and magnitude of the flood events are reduced compared to natural conditions. Diversions can have serious impact on native, amphidromous organisms by disrupting larval drift to the ocean and obstructing postlarval recruitment (return migration) by reducing or eliminating flow downstream

(Timbol & Maciolek 1978, Kinzie 1990). Disruption of this lifecycle could significantly lower the number of breeding populations of these native Hawaiian species. Removal of

! stream flow can also result in alterations in downstream habitat subsequently altering the physical and chemical conditions that regulate biological communities; thus, limitations in habitat availability may shift species interactions, eliminate native and introduced aquatic species, potentially increase the number of invasive species, and alter the entire food web (Cowx et al. 1984, Poff et al. 1997, Brasher 2003). This is of particular concern in Hawaii because of several native and endemic species that are sensitive to changing environmental conditions and introduced species (Shoda et al. 2010).

Na Wai Eha Watershed

The ecosystems under study are of great interest and have been the topic of a heated debate. In 2004, on the island of Maui, an initiative by local interest groups, in cooperation with government agencies began efforts to determine interim stream flow levels for the N! Wai ‘Eh!. The N! Wai ‘Eh! is considered the four major watersheds of the West Maui Mountains (Waikapu Stream, Iao Stream, Waiehu Stream and Waihe’e

River). One of the major objectives of this effort was to determine adequate flow volumes necessary to support healthy stream biological communities. It is understood that water must be returned to these streams to again provide a continuum of stream flow and habitat for aquatic organisms; however, it is the minimum amount of stream flow needed to maintain reproductive populations that remains unknown. Tentative restoration plans include controlled releases for a final restoration level equal to the estimated long-term minimum daily mean flow. However, in order to understand biological responses of these aquatic communities, baseline data on these amphidromous and non-amphidromous invertebrate communities under historic diverted conditions is needed.

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! Water Management Initiatives

Increasing demands for water in the state of Hawaii have put tremendous pressure on water managers who require a sound basis for making water allocation decisions

(DAR 1996, Kinzie et al. 2006). Of particular interest is the importance of naturally occurring flow variability which influences channel maintenance, clearing of debris dams, promotion of migration and movement of amphidromous species, cues for reproductive cycles of stream organisms, enhancement of primary or secondary production, control of alien species, provision of habitat heterogeneity, and maintenance of benthic communities (March et al. 2003, Brasher 2003, Kinzie et al. 2006). Because flow in many Hawaiian streams is presently diverted (Timbol & Maciolek 1978, Wilcox

1996), understanding human impacts on flow in stream systems is critical for management and mitigation (HCPSU 1990, March et al. 2003, Kinzie et al. 2006).

The task of determining minimum flows necessary to prevent detrimental effects, while fulfilling freshwater needs, has become a popular, yet difficult task for water resource managers (Bunn & Arthington 2002, Dewson et al. 2007). Total elimination of dams and water diversions in tropical streams is not an appropriate solution, instead, impacts can potentially be reduced through various structural and operational changes such as increased minimum flow, installation or improvement of fish and shrimp ladders and periodic releases of flushing flows (March et al. 2003, see Bednarek & Hart 2005). If ecological structure and function is to be restored and maintained to stream networks, water-conservation strategies and a strong commitment to sustainable use of water resources must be implemented, especially on tropical islands where freshwater resources are already limited and endemism is high (Smith et al. 2003). In the last decade, it has

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! become increasingly evident that conserving aquatic ecosystems while balancing the anthropogenic need for freshwater is a pressing issue around the world – urging a deeper understanding of the direct and indirect impacts of flow modifications on aquatic communities and overall ecosystem services, for the development of effective management strategies (Power et al. 1996, Jackson et al. 2001, Baron et al. 2002, Brasher

2003).

Part 3: Tropical vs. temperate streams

The ecology of insular tropical stream and river drainages are not well understood. Most tropical streams differ significantly from those found on continents by having relatively short, straight and steep channels in comparatively small, narrow watersheds (Smith et al. 2003). Temperate streams are more seasonally driven whereas tropical stream networks tend to be more event-driven, which may be relatively unpredictable (Smith et al. 2003). Although extensive research has documented the impacts of hydrologic alterations on biotic communities in temperate systems, far less is known about biotic responses to hydrologic modifications in tropical systems (Pringle et al. 2000, Smith et al. 2003).

It is not yet clear how the concepts derived from studies of continental stream ecosystems in the temperate zone can be effectively applied to understand and manage streams and rivers on islands in tropical areas (Smith et al. 2003). There is a growing recognition, from studies of both temperate continental and tropical insular ecosystems, of the importance of conserving the biological resources of watersheds as well as the other “ecosystem services” they provide such as a clean water supply, fishery resources, and recreation (Smith et al. 2003). The demand on the freshwater resources provided by 12!

! these island watersheds is especially critical as this demand will continue to increase due to fast growing populations, which may ultimately transform and irreversibly alter these ecosystems (Brasher 2003, March et al. 2003, Smith et al. 2003).

Comparison among stream systems is necessary because the vast amount of research undertaken on north temperate streams has given rise to ecological models (i.e. the River Continuum Concept: Vannote et al. 1980; the Riverine Productivity Model:

Thorp & Delong, 1994) that are now being used to guide research questions and management approaches in rivers worldwide (Boulton et al. 2008). Evaluations between temperate and tropical streams are confounded by immense variability inherent in these systems, and the wide range of climatic, geomorphology and hydrological conditions that may generate a habitat template for ecological differences (Boulton et al. 2008). Further, understanding diversity differences have been hampered by the imbalance in research between these climatic zones. For example, lists of species are scarce for the tropics – especially for macroinvertebrates – and the identification of tropical species has been difficult for non-specialists (Boulton et al. 2008).

While these systems may not be routinely compared in singular studies, Boulton et al. (2008) summarized in a review of the literature, that there were no consistent differences in food web structure, productivity, organic-matter processing or nutrient dynamics, and response to disturbance between tropical and temperate systems and that the adjective ‘tropical’ has no particular significance when applied to stream ecology.

Instead, ecological processes in tropical streams tend to be driven by the same variables that are important in temperate ones. Consequently, valid extrapolation of models and management strategies may not be so much an issue of latitude (tropical vs. temperate)

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! but of ensuring suitable comparability at an appropriate scale. However, this review clearly demonstrated that whereas ecological mechanisms may be similar, the organisms involved can and do differ (Boulton et al. 2008).

Similarly, Greathouse & Pringle (2006) investigated the application of the river continuum concept on a tropical island stream and found that while collector-filterers showed a trend opposite to that predicted by the model, patterns in basal resources suggest that this was consistent with the central theme: longitudinal distributions of FFGs follow longitudinal patterns in basal resources. Their results indicated that the river continuum concept generally applies to tropical streams, however they concluded that additional theoretical and field studies across a broad array of stream types was necessary to examine whether the river continuum concept needs to be refined to reflect the potential influence of top down trophic controls on FFG distributions (Greathouse &

Pringle 2006).

Although these studies have begun to investigate the relationship between tropical and temperate stream systems, further comparative studies are necessary to understand the application of ecological models in tropical systems. Further, there is a growing need to establish more unified themes of research, management and conservation in tropical stream ecosystems (Smith et al. 2003).

14!

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19!

CHAPTER II

DISPERSAL AND UPSTREAM MIGRATION OF AN AMPHIDROMOUS NERITID

SNAIL: IMPLICATIONS FOR RESTORING MIGRATORY PATHWAYS IN

TROPICAL STREAMS

With kind permission from Wiley-Blackwell publishing K.R. Gorbach, M.E. Benbow, M.D. McIntosh, A.J. Burky (2012). Dispersal and upstream migration of an amphidromous neritid snail: implications for restoring migratory pathways in tropical streams. Freshwater Biology, 57, 1643-1657.

SUMMARY

1. The amphidromous lifecycle of several species of neritid snails, shrimp and gobies throughout the tropics includes juveniles that migrate from the ocean to breed in fresh water. In many Hawaiian streams, the decline of Neritina granosa, an endemic gastropod, has been associated with habitat degradation and water withdrawal, which are common factors affecting tropical rivers around the world.

2. We investigated the effects of water withdrawal and density on dispersal and upstream migration of N. granosa using three experimental treatments: 1) reduced flow due to a stream diversion, 2) natural flow, and 3) natural flow with artificially increased snail density. For each treatment snails were differentially tagged and released in a stream without a natural, extant population of N. granosa.

! 20 3. Capture rates ranged from 17 – 65% over a 63-day period following release. Captures on 2 – 6 days after release measured initial dispersal and migration, whereas longer-term migration rates were calculated from snails captured 16 – 63 days after release. Snails under natural flow displayed positive rheotactic behaviour, with only 3 – 12 % demonstrating initial downstream movement. Under reduced flow 22 – 77 % of snails moved downstream or showed no bias either way.

4. Initial mean upstream migration rate was 0.25, 0.66 and 1.16 m d-1 under reduced flow, natural flow and natural flow with increased snail density, respectively. Longer- term migration rates did not differ significantly among treatments and the overall mean was 0.62 m d-1.

5. Principal component analysis and generalized linear models were used to identify habitat characteristics important to upstream migration rate, with habitat and reach-scale hydraulics as the most important factors. The relationship between discharge and upstream migration rate suggested it would take 11 – 35 years for snails to migrate past the most upstream water diversion. However, rates from published studies of neritid snail species migrating en masse or in long lines under natural situations, suggested that N. granosa could migrate above stream diversions within 72 days to 2.5 years (when in an aggregation) and 29 days to 1.1 years (when following in long lines).

6. An understanding of upstream neritid snail migration can be used for management and conservation of this and other migratory species in tropical streams.

! 21 INTRODUCTION

Dispersal, and often migration, is an essential component of the life history and ecological niche of many organisms (Dingle & Drake, 2007). Migration has been documented for gastropod snails in tropical (Paulini, 1963; Ford & Kinzie, 1982;

Schneider & Frost, 1986; Wallace, 1992; Blanco & Scatena, 2005, 2006, 2007; see Pyron

& Covich, 2003), temperate (Bovjberg, 1952; Ball, Wojtalik & Hooper, 1963; Houp,

1970; Mancini, 1978; Burris, Bamford & Stewart, 1990) and intertidal habitats (Garrity

& Levings, 1981; Crowe, 1996). Migration from the ocean into fresh water occurs in two subclasses, three orders and 10 gastropod families (Hutchinson, 1967; Graham, 1985;

Huryn & Denny, 1997).

Many migratory taxa move between marine and fresh water to complete their lifecycle, dominating tropical island systems in numbers and biomass (Gross, Coleman &

McDowall, 1988; Blanco & Scatena, 2007). More unusual is the amphidromous lifecycle of several decapod crustaceans, neritid snails and gobiid fishes of Oceania, the

Neotropics, Indomalaya, Japan and New Zealand (Benbow & McIntosh, 2009;

McDowall, 2004 & 2007). This lifecycle is a type of diadromy where adults live, breed and deposit eggs in streams; upon hatching, planktotrophic larvae drift to the ocean where they grow and develop, after which postlarvae move back into stream mouths and begin an upstream migration to the adult habitat where they mature and breed (McDowall,

1998, 2004, 2007; Hodges & Allendorf, 1998; Crandall et al., 2010). Migration has been described as movement that is regular in terms of season, direction and life stage, and where the organisms occupy two distinct and well-separated habitats (see McDowall,

2007).

! 22 Neritina granosa (Sowerby) is an amphidromous endemic Hawaiian gastropod found on bedrock, boulders and large gravel in clear, cool, well oxygenated, and fast flowing perennial streams (Ford, 1979). After the pelagic larval stage, spat (snail postlarvae ! 5 mm in shell length) settle at a stream mouth and begin an upstream migration until finding adult habitat (Ford, 1979; Way et al., 1993; Hau, 2007). During the upstream migration, spat (! 5 mm) and juveniles (" 6 mm) grow to about 9 mm over several months; growth then continues at a rate of 1 – 3 mm y-1, slowing until they reach a mean maximum adult length of 29 mm (Brasher, 1997). Ford (1979) suggested a 10- year life span, while studies undertaken by Brasher (1997) implied a 6 – 10 year life span.

However, we recovered tagged snails in Kinihapai Stream, Maui in 2009 that were released in 1994, confirming a potential life span > 10 years (Benbow, unpublished).

Migrating spat, juveniles and other pre-adults of N. granosa remain on the underside of rocks by day, feed nocturnally, and usually inhabit areas of strong flow. However, adults can be found in pools mating and grazing during the day (Maciolek, 1978). Brasher

(1997) proposed that smaller snails (! 9 mm) are the migrating stage, while adults lose the rheotactic response.

Upstream migration of neritid snails has attracted much research attention

(Schneider & Frost, 1986; Schneider & Lyons, 1993; Blanco & Scatena, 2005 & 2007). It has been suggested that migration relates to the search for food and space (Paulini, 1963), predator avoidance (Gross et al., 1988; Schneider & Lyons, 1993; Blanco-Libreros &

Arroyave-Rincón, 2009), a response to accidental downstream drift (Carpenter, 1928;

Schneider & Frost, 1986), constraints imposed by body architecture and hydrodynamics

(Haynes et al., 1985; Way et al., 1993; Huryn & Denny, 1997), the availability of

! 23 breeding sites (Söderström, 1987) and variation among adults and juveniles in tolerance of physical factors such as salinity and temperature (Pyron & Covich, 2003). Further, seasonal change, channel substratum, distance from the ocean, flow hydraulics and water depth may affect migration at different spatial scales (Way et al., 1993; Pyron & Covich,

2003; Blanco & Scatena, 2005). Massive upstream snail migration has been hypothesized to be density and body size dependent; spat and juvenile forms (generally ! 9 mm in length) have been observed to move together in long lines and/or dense aggregations

(Schneider & Frost, 1986; Schneider & Lyons, 1993; Brasher, 1997; Pyron & Covich,

2003; Blanco & Scatena, 2005; Hau, 2007). In Puerto Rico, Blanco & Scatena (2006 &

2007) concluded that younger snails prefer fast, turbulent and erosive habitats, and that densities were greater in deep habitats with heterogeneous substrata thus indicating that migration might ultimately be influenced by stream discharge and channel hydraulics.

Diversions and dams are responsible for modified flow regimes, fragmented populations, obstructed breeding migrations and the loss of navigational cues and endemic species in tropical streams (Drinkwater & Frank, 1994; Benstead et al., 1999;

Pringle et al., 2000; Dudgeon, 2003). Further, natural amphidromous populations are important components of tropical stream communities but are profoundly affected by altered flow regimes (McIntosh et al., 2002; Brasher, 2003). To our knowledge, this is the first experimental field study using a transplant and monitoring approach to evaluate the effects of stream diversions on the upstream migration of an amphidromous, neritid snail, with application for stream management and conservation in other regions of the world. Our objectives were to determine patterns of N. granosa spat initial dispersal and upstream migration rates under reduced flow conditions and increased spat density. We

! 24 hypothesized that reduced stream flow would negatively affect rheotaxis and initial dispersal with ultimate consequences for migration rate and distance, while increased snail density would have opposite effects. We also sought to understand the relationship between amphidromous snail upstream migration rates and hydraulic habitat characteristics and, further, to use published stream discharge data and neritid migration rates to predict the time taken to migrate by tropical amphidromous neritid snails, for application in catchment management and conservation.

METHODS

Snail collection & tagging

Neritina granosa were collected from Honomanu Stream, East Maui (Fig. 1).

While a diversion on this stream ~2.5 km upstream of the ocean usually results in a dry streambed, a natural spring ~250 m upstream of the tidal influence maintains a wetted channel at the stream mouth. A large number of N. granosa accumulate below this spring, where we collected them for the transplant study.

On 11 June 2001, rocks were carefully lifted and N. granosa ranging from 2.6 –

7.0 mm in shell length, including spat (! 5mm) and young juveniles (6 – 7 mm), were collected and placed in a cooler of stream water. Immediately following length and width

(±0.1 mm) measurements using an electronic caliper (Mitutoyo 700-103), 3.0 mm coloured and numbered bee tags (The Bee Works, US; Canada) were applied with Super

Glue ® to 396 snails and allowed to air dry. Snails were returned to aerated buckets of stream water at low densities resulting in little mortality (< 0.5%) and resumed a nocturnal movement pattern (crawling in circles) for at least six hours before being transplanted to the study stream. On 5 July 2001, an additional 2,600 snails were

! 25 measured of which 198 were tagged using different colours from the previous release, while 2,157 were left untagged (Table 1). In previous research, snails with these tags have been recovered after >10 years; the tag numbers are often not legible, but are in place and sometimes covered by nacreal shell formation that, when removed, reveals number and colour (Benbow, unpublished). We monitored initial snail dispersal and migration rates under three experimental treatments: 1) reduced flow conditions (RF) due to a stream diversion, 2) natural flow conditions (NF) upstream of the diversion, and 3) natural flow conditions with increased spat density (NF+D).

Release sites and treatments

The release sites were in Iao Stream, a 13.4 km second order stream, located in the Wailuku catchment on the northeastern side of the West Maui Mountains, Hawaii

(Fig. 1). Iao Stream is diverted at three locations, withdrawing ~0.8 m s-1 for agricultural and developmental purposes (Shade, 1997; Benbow, 1999). Stream flow breaches all three diversions and flows to the ocean only after high rainfall (Oki, 2007; Oki et al.,

2010). Because of these diversions, constructed in the mid-1800’s, natural populations of

N. granosa were not found in the study reaches; this was tested in hundreds of person- hours doing snorkel searches during the weeks before snail releases and for nearly six years of prior routine surveys (Benbow, unpublished). The absence of natural populations facilitated snail release and capture and also ensured no pre-existing slime trails could affect movement and migration (Pyron & Covich, 2003; Hau, 2007). The NF and NF+D release site was ~1.0 km above the highest altitude diversion while the RF site was located ~100 m downstream of this diversion, where flow was reduced by 92 – 98% of daily base flow (Fig. 1; McIntosh et al., 2002).

! 26 On 12 June 2001, 198 tagged snails were released at both the NF and RF sites.

The NF study reach was on average 8 m wide and characterized by deep, fast and turbulent flow, numerous riffles and cascade habitats, while the RF study reach was on average 4 m wide and characterized by shallow and slow flow, infrequent riffles and small shallow pools. On 13 July 2001 an additional group of 198 tagged plus 2,157 untagged snails was released under natural flow (NF+D) at the NF site. The release sites in each study reach appeared similar – small pools along the bank in a 0.06 m2 area of no flow, ~0.25 m in depth and protected from the thalweg. When standardized for benthic area, the density of released snails was 3,168 m-2 for RF and NF treatments, and 37,680 m-2 for NF+D. Immediately after release, snails were observed for 15 minutes and then the small pool was covered with a large, flat rock suspended above the water to provide cover against birds.

Snail capture

We searched for snails on days 2, 3, 5, 6, 16, 33 and 63 after they were released.

Two to three trained researchers, adjacent to each other across the stream, worked upstream searching under and around all substrata using facemasks and snorkels. When large, immovable boulders were encountered, effort was made to feel in and around all accessible surfaces and interstitial spaces to locate spat. Each observed snail was considered a capture with its location marked by an x-y coordinate system, where the origin (0,0) was the point of release, x the perpendicular distance to the bank of origin, and y the parallel distance up- or downstream along the bank of origin; measurements were made to the nearest 0.1 m using a tape measure stretched between a depth rod at the capture point and a person on the adjacent bank. At each point, snail tag number and

! 27 colour, x-y coordinate, depth and water velocity using an Ohio Professional electronic impeller flow meter (The Great Atlantic Trading Co. Ltd.) were recorded; measurements were subsequently used to calculate Froude and Reynolds numbers (Statzner, Gore &

Resh, 1988). Unsafe stream conditions prevented search efforts on day 2 of the NF and day 6 of the NF+D study periods. Captures made on days 2 – 6 after release were used to determine initial dispersal and migration whereas longer-term migration rates were calculated from captures made on days 16, 33 and 63 after release. Capture rates were calculated as the percentage of the total number of snails released on the original date that were found on a subsequent search day.

Discharge and water temperature were measured on each search day and the volume of diversion flow removal was estimated by measuring discharge above and below it using the velocity-area method (Gordon, McMahnon & Finlayson, 1992). In addition, stream discharge data were retrieved from a U.S. Geological Survey gauge

(16604500) immediately upstream of the diversion. Because attempts to normalise data were unsuccessful, Kruskal-Wallis non-parametric one-way analysis of variance

(GraphPad Prism 5.0 Software) was used to test habitat differences among treatments, while Mann-Whitney tests were used when only two treatments could be compared (i.e. unsafe stream conditions).

Initial snail dispersal and upstream migration

Snails could move in any direction from the release location, so for the purposes of this study any upstream movement was considered migration. Initial snail dispersal during days 2 – 6 after release was described as downstream, upstream or neutral

(remaining at release site) from the release location. Snails that moved up- or downstream

! 28 not only moved parallel (y-coordinate) to the flow but also perpendicular to it (x- coordinate) as they dispersed. To represent the degree of snail aggregation, the coefficient of variation (CV) of the total captured snail population movement along the x-axis was compared among days and treatments: a low CV represented aggregation, while a larger

CV represented more independent movements. Further, over the six days after release, the Euclidean distance between the re-captures of any particular snail was used to calculate a Euclidean Migration Rate (EMR as m d-1). Movement parallel to stream flow between re-captures defined an Upstream Migration Rate (UMR as m d-1). Kruskal-

Wallis non-parametric one-way analysis of variance with Dunn’s post-hoc tests and nonparametric two-way ANOVA with Bonferroni post-hoc tests for pairwise comparisons were used (GraphPad Prism 5.0 Software) to test for differences in EMR and UMRs among treatments and days. Mann-Whitney tests were used when data for all three treatments were unavailable.

Snail movement that demonstrated rapid migration (" 8 m in ! 6 d, ‘rapid’ snails) was estimated to provide a maximum upstream migratory potential not weighted by a few individuals that affected initial movement and migration dynamics. To assess migration beyond our initial 6 day period, mean longer-term UMRs were calculated from fewer

(N=31) captures made 16, 33 and 63 d after release. Differences in UMR among snails that migrated ! 8 m in ! 6 days, the ‘rapid’ snails and the longer-term estimates were tested using Kruskal-Wallis non-parametric one-way analysis of variance.

Migration – habitat relationship

The same habitat variables measured or calculated on each search date were also used to evaluate the possible effect of previous habitat conditions encountered the day

! 29 before a re-capture. Several analytical steps were employed to test for relationships between habitat variables and upstream migration rates to represent a gradient of migratory conditions (Supporting Information, Appendix S1).

Application to restoration ecology of amphidromous migration

Although microhabitat characteristics may be more specific, overall stream discharge is often readily available from gauging stations and may be a practical tool for predicting upstream migration rates of amphidromous species. Measured mean daily discharge was used to predict mean daily UMRs using simple linear regression. These regression models were then used to predict migration rates under different quantiles of published stream discharge conditions from 1985-2005 (Oki, 2007). The Q50, Q70, and

Q90 were determined using duration curves for three diverted West Maui Streams (Oki,

2007): Iao Stream, Waiehu Stream north and south branches, and Waihee River. For these streams, Oki (2007) also provided the minimum discharge necessary to maintain stream flow to the ocean (Qmin). We used these quantiles and regression models for each stream to estimate the time (in days and years) necessary for N. granosa to migrate from the ocean to the reach upstream of the highest altitude diversion. Further, in order to provide estimates that would include the effect of aggregate migration, we compared our upstream migration rates to published studies of other neritid snails under natural conditions.

! 30 RESULTS

Habitat characteristics & snail capture

There was a 98% reduction in stream discharge below the diversion that corresponded to a 49% reduction in available habitat and warmer water temperatures

(Mann-Whitney U = 1316, P < 0.0001) (Table 2). Snails in the NF+D treatment population used habitats with greater hydraulic intensity; mean velocity, Froude and

Reynolds numbers followed the trend of RF < NF < NF+D (Table 2). Interestingly, the

CV for the hydraulic variables was greater in the NF population compared to both RF and

NF+D (Table 2), suggesting that microhabitat conditions of NF snails were more variable than the other populations. Capture rates over the first 6 d ranged from 17 – 65%, with the highest from the NF+D population (Table 3). The mean capture rate among all snails was 39%.

Initial snail dispersal and upstream migration

Over the first 6 d, mean initial dispersal patterns revealed substantially more downstream/neutral snail movement in the RF (43.3%) compared to NF (5.7%) or NF+D

(5.3%) populations (Fig. 2; Table 3). Snails moved upstream, downstream and also perpendicular to the flow. As a measure of aggregation, the CV for lateral movement across the channel decreased from 1.01 on 2 d to 0.53 on 6 d, from 0.62 on 3 d to 0.27 on

6 d, and from 0.65 on 2 d to 0.29 on 5 d, for the RF, NF and NF+D populations, respectively, inferring increased movement in aggregation with higher discharge and snail density. The mean 6 d EMR for NF+D snails was significantly higher than the NF snails, while the latter were almost 5x faster than RF snails (Fig. 3). Similarly, mean

UMR more than doubled from RF to NF and from NF to NF+D (Fig. 3). There were

! 31 significant treatment (F = 17.93, P < 0.0001, df = 2), day (F = 12.67, P = 0.0004, df = 1) and interaction (F = 5.94, P = 0.0028, df = 2) effects on UMR between days 3 and 5 (Fig.

4a). The mean UMR of the RF population differed over all days, driven by an increase in upstream migration on day five, whereas it was consistent across days for the other populations (Fig. 4a).

Four, three and 31 snails exhibited ‘rapid’ migration (" 8 m in ! 6 days) under

RF, NF and NF+D conditions (Fig. 2); however, upstream migration rate was not different among the populations (Kruskal-Wallis, H = 1.19, P = 0.55). Thus, these data were pooled for a mean ± SE maximum potential UMR of 2.18 ± 0.14 m d-1 (N = 38)

(Fig. 4b). Similarly, upstream migration rates from 16, 33 and 63 days after release did not differ among populations (Kruskal-Wallis, H = 3.99, P = 0.14) and the mean longer- term UMR from pooled data was 0.62 ± 0.06 m d-1 (N = 31) (Fig. 4b).

Migration – habitat relationship

The cumulative variation in upstream migration rate explained by retained habitat variables in PCA ordinations (Appendix S2) was 77%, 83% and 80% for RF, NF, and

NF+D populations, respectively. Using generalized linear models, habitat-scale hydraulics, including velocity, Froude and Reynolds number, were positively related with upstream migration rate across all populations (Appendices S3 & S4). Snail spatial configuration (aggregation) explained 3 – 31% of the residual variation, and position in the thalweg influenced migration rates.

! 32 Application to restoration ecology of amphidromous migration

We found a positive linear relationship between mean daily discharge and UMR

2 (y = 0.43x + 0.45, R = 0.23, P = 0.16). Using this model to predict UMRs under Qmin,

Q50, Q70, Q90 discharge quantiles it would take N. granosa spat from 11 years at Q50 to 35 years under Qmin conditions to migrate beyond the highest altitude abstraction point

(Table 4). Because our migration rates are from a transplant experiment in higher altitude reaches, and do not represent natural aggregate conditions reported elsewhere to be much faster, we used published studies of N. granosa and related Neritina spp. to provide a range of time estimates needed to restore naturally migrating neritid populations in West

Maui streams (Tables 4 & 5). In general, non-aggregate migration < aggregate migration

< migration in long lines.

DISCUSSION

Water flow alterations by dams and diversions change downstream habitat for aquatic organisms (Dewson, James & Death, 2007). In Iao Stream, discharge was significantly lower downstream of the diversion, negatively affecting water depth and the wetted substratum available for benthic movement, and water temperature was higher.

Depth did not differ between NF and NF+D, because these treatment conditions were in the same reach during different study periods. However, habitat hydraulic variables associated with snail capture locations were significantly higher in the NF+D population, which may indicate snail preference for faster flowing microhabitats, or that the overall channel reach had been modified. Further, the coefficient of variation for these variables was less than those of the NF population, illustrating the use of more similar hydraulic habitats for snails under the increased density treatment.

! 33 Mean capture rate over all treatments was 39% and greatest at 65% for the NF+D population. Other studies have documented higher capture rates, such as 95% by Pyron

& Covich (2003); however, 40-60% of these snails remained at the release site. Such variation may be due to stream habitat differences, organism size, tagging and search techniques or season (Covich, Crowl & Scatena, 2003; Pyron & Covich, 2003). In our reduced flow treatment there was a greater new capture rate, which may have been due to restricted depth and a narrow channel that reduced the overall search area.

Under reduced flows, initial snail movement appeared disoriented, displaying non-rheotaxis and nearly random movement from the release location. On day 2, 77% of these snails did not move, or if they did, moved downstream. This may be related to the pool-like habitat ~4 m up- and downstream of the release location, characterized by shallow water, very slow flow and back eddy currents; a current which could result in downstream dispersal, as the snails would have moved against the flow, displaying rheotaxic behaviour (Vermeij, 1969; Ford, 1979; Schneider & Frost, 1986). Thus, large- scale discharge removal changes smaller-scale hydraulic conditions that could be disorienting to migrating spat, mediating dispersal patterns and upstream migration. A small cascade (0.25 m vertical drop), about 4 – 5 m upstream of the release point did generate increased flow, possibly providing a migratory cue not initially detected by the released snails, but later in migration. Although cascades may act as a migration barrier,

N. granosa can overcome cascades as long as there is some flow and wetted substratum

(Ford, 1979). The few RF snails that reached the cascade and beyond moved 14 m in 6 d, similar to migration rates and distances of both natural flow snail populations. This may indicate that, once snails detect flow, they display a more natural upstream migration.

! 34 The two populations under natural flow exhibited very little (<10%) downstream/neutral movement, with a pronounced pattern moving upstream and toward the thalweg. These pattern differences may be due to increased density of migrating snails, the use of more similar and hydraulically intense microhabitats, or both. Ford (1979) suggested that N. granosa spat migration may not be density dependent; however, our results indicate that increased numbers enhance upstream migration.

According to Michel, McIntyre & Chan (2007), snail movement rates have been calculated for surprisingly few species. To contribute to this knowledge base, we have described neritid migration under altered flow and density conditions. Under reduced flow, upstream migration rate was half that under natural flow, and this rate increased by two-fold with increased spat density. The initial release density was high; however, visual observations suggested that such conditions did not persist as the snails moved within the available habitat. Even though our daily densities did not remain as high as some recorded neritid mass migrations, where >5,000 m-2 have been documented (Blanco &

Scatena, 2007), our experimentally increased density did have a significant positive effect, indicating that naturally occurring en masse aggregative migration is probably much faster than in our transplanted snails. Further, upstream migration rate was consistent over all recovery days under natural flow conditions, whereas the rates of reduced flow snails significantly increased 5 d after release. This could have been due to the small cascade upstream of the release location. Our results demonstrate that N. granosa spat potentially have accelerated migration rates under natural flow and aggregative conditions.

! 35 The migratory behaviour of natural populations of Neritidae has been documented in several studies. Long lines of snails moving over a few centimetres have been measured and, when extrapolated, range impressively from 15 – 250 m d-1 (Ford, 1979;

Schneider & Frost, 1986; Schneider & Lyons, 1993; Pyron & Covich, 2003). Although it is unlikely this rate of travel would continue throughout the day and over seasons, snails are capable of large-scale coordinated upstream migration (Schneider & Frost, 1986).

Notably, upstream migration rate estimates calculated from Brasher (1997), where young snails 11 mm in shell length had a mean upstream movement of 21 m in one month, or about 0.7 m d-1, are very similar to our mean upstream migration rate of 0.66 m d-1 under natural flow conditions (Table 5).

Gastropoda are widely known to produce adhesive mucus trails as they move, generating a trail-following phenomenon (Bretz & Dimock, 1983; Denny, 1989; Smith &

Morin, 2002). A possible, yet untested, explanation for the higher migration rate and movement pattern in the increased density population could be from active mucus trail- following – directional information from chemical cues within the mucus or energy conservation by moving over previously laid trails (Shaheen et al., 2005; Stafford &

Davies, 2005; Alfaro, 2007; Davies & Blackwell, 2007). Although mucus decay is not fully understood in lotic systems, conclusions drawn from marine intertidal studies

(Connor, 1986; Calow, 1979; Herndl & Peduzzi, 1989; Davies & Beckwith, 1999; Davies

& Blackwell, 2007) indicate that trails laid by snails under natural flow conditions would not have affected snails under increased density because of degradation between study periods (four weeks). Further research is needed to understand more of the effects of mucus trails on aggregate migration of freshwater snails.

! 36 The importance of flow velocity in the ecology, behaviour and physiology of stream organisms is well established for many temperate stream organisms (Statzner et al., 1988). Our results identified habitat- and reach-scale hydraulic variables as significant predictors of upstream migration rate of an amphidromous tropical snail. The remaining variation may be attributed to variable periphyton productivity, microhabitat hydraulic conditions, substratum variability, and the use of more dense aggregate movement, slime trails or hitch hiking behaviours (Schneider & Lyons, 1993; Pyron &

Covich, 2003; Blanco & Scatena, 2005 and 2007; Hau, 2007; Kano, 2009). While other environmental characteristics may influence movement and upstream migration, our results concur with Way et al. (1993), who proposed that micro-scale habitat flows directly influence paths of grazing, refugia from predators, migratory pathways and spawning areas of N. granosa. Additional studies have noted that large numbers of smaller snails (<8 mm) tend to be found in, and move against, fast and turbulent flow

(Vermeij, 1969; Ford, 1979; Schneider & Frost, 1986; Way et al., 1993; Brasher, 1997;

Huryn & Denny, 1997; Blanco & Scatena, 2005, 2006 & 2007). Huryn & Denny (1997) proposed that hydrodynamic drag on the shell generates torque on the foot and causes the snail to rotate until its anterior end faces upstream, essentially steering snails upstream.

Without sufficient velocity to produce such torque, movement may not occur in an upstream direction. In high-flow events, common to Hawaiian streams (Moore, 1964), N. granosa are usually found on the underside of boulders or in crevices (Brasher, 1997), and are well adapted to maintain position on the substratum (Ford, 1979); again illustrating the role of flow in determining snail movement and placement.

! 37 Across Hawaii and other tropical islands, water withdrawal from streams continues to occur (Hawaii Cooperative Park Service Unit, 1990). Our results indicate that 98% of base flow was removed from Iao Stream, which can lead to the decline of native and endemic fauna (Oki et al., 2010; Shoda et al., 2010). While this threat is understood and acknowledged, migration time and habitat requirements have not been fully investigated. This study demonstrates that N. granosa can successfully be transplanted to dewatered streams; however, given the current diverted conditions, sustained recruitment for long-term restoration of resident populations is not possible.

Using our regression models, we estimated that it would take N. granosa from 11 – 35 years to migrate above the highest altitude abstraction points in the West Maui catchments. These are overly conservative and unrealistic estimates, considering a potential 6 – 15+ year life span (Ford, 1979; Brasher, 1997; Benbow, unpublished). Such unrealistic estimates could be because spat were transplanted from a lower to higher altitude and under relatively low densities compared to other studies of N. granosa aggregations (Ford, 1979; Hau, 2007), thus, creating experimental conditions that were not representative of natural high density spat aggregations at stream mouths. Because of such limitations, it was necessary to use other Neritidae migratory studies to estimate time frames that may be broadly applied to other regions. Snails reportedly moving en masse allow us to consider that N. granosa would have the ability to migrate above stream diversions in realistic time frames under natural stream flow conditions within 72 days to 2.5 years, when moving as an aggregate, and 29 days to 1.1 years, when traveling in long lines. Perennially restored flow could have effects on N. granosa migration, where large densities would probably entrain at stream mouths to form natural

! 38 aggregations, like those reported by Ford (1979), and be capable of migrating to altitudes of at least 400 m (Maciolek, 1978), ultimately enabling natural populations to re-establish at higher altitudes.

Dudgeon (2000) drew attention to the need for tropical research with international relevance. In this context, our study contributes to the growing literature on tropical stream organisms and, more specifically, on the factors affecting the upstream migration of ubiquitous but understudied, amphidromous neritid snails. As the threat of decline of native populations continues, discussions of restoration and the development of management plans are becoming increasingly important. If restoration efforts use this migratory information on the slowest amphidromous species, surely it is possible for faster species to be restored concurrently (Benbow et al., 2002). Therefore, understanding time frames necessary for these species to reach natural habitats and adult breeding grounds may facilitate mitigation practices that would restore pathways for all amphidromous organisms and could be applied in other tropical regions experiencing similar disturbances and population decline.

ACKNOWLEDGMENTS

Support for this study was provided by the University of Dayton Graduate School, Learn,

Lead and Serve Program and Department of Biology, the Earthwatch Institute and Center

For Field Studies, and the Office of Hawaiian Affairs. We would like to acknowledge T.

Fernandes, L. Orzetti, M. Shoda and Earthwatch Institute volunteers that assisted during the 2001 season. We would also like to thank the Hawaii Department of Natural

Resources, Division of Aquatic Resources for granting research permits, S. Hau for

! 39 continued support of our research, and two referees and Alan Hildrew for improving earlier drafts.

! 40 Table 1: Shell length, width and height of Neritina granosa collected from Honomanu Stream, Maui, Hawaii. Height was measured for only the tagged snails of the Natural flow + increased density (NF+D) treatment. Snail density was determined from snail counts in nine, 1m2 quadrats, where water depth was also measured.

Measurement N Range (mm) Mean (mm) ± SE C.V. (%)

Tagged snails

Length 594 2.6 - 7.0 5.2 ± 0.05 23.7

Width 594 2.1 - 6.9 3.9 ± 0.04 23.8

Height 198 1.3 - 2.7 1.9 ± 0.02 13.3

Untagged snails

Length 2402 2.2 - 22.3 6.8 ± 0.05 35.7

Width 2402 1.6 - 20.1 5.2 ± 0.04 38.9

Height 2402 0.4 - 8.0 2.6 ± 0.01 27.5

Quadrat measurements

Density (snails m-2) 9 17 - 1192 309 ± 130.10 126.3

Depth (m) 9 0.04 - 0.54 0.32 ± 0.06 52.9

41 Table 2: Habitat and hydraulic variables for each treatment during Neritina granosa migration in Iao Stream, Maui, Hawaii. The three treatments were Reduced Flow (RF), Natural Flow (NF) and Natural Flow + Increased Density (NF+D). The Range, Mean (SE), Coefficient of Variation (CV%), Kruskal-Wallis analysis of variance test statistics, and the results of Dunn’s post-hoc tests are provided.

Kruskal- Wallis test Environmental Variable Treatment Range Mean ± SE C.V. (%) N P Value statistic

Water Temperature (°C) RF 19.5 - 21.0 20.3 ± 0.44 3.8 3 NF 18.5 - 19.0 18.7 ± 0.17 1.6 3 0.0538 5.84 NF+D 20.5 - 21.5 20.8 ± 0.33 2.8 3

Stream Width (m) RF 1.0 - 7.4 4.3 ± 0.61 50.6 13 *t=3.84, NF 4.0 - 12.5 8.4 ± 0.90 36.7 12 0.0008 df=23 NF+D 4.0 - 12.5 8.4 ± 0.90 36.7 12

Depth (m) RF 0.12 - 0.67 0.39 ± 0.01 30.6 132 NF 0.08 - 0.81 0.48 ± 0.01 40.7 263 <0.0001 48.88 NF+D 0.03 - 0.88 0.51 ± 0.01 38.9 294

Surface Velocity (m s-1) RF 0.0 - 0.71 0.19 ± 0.02 78.7 40 NF 0.0 - 1.95 0.37 ± 0.02 100.1 252 <0.0001 27.63 NF+D 0.0 - 1.50 0.43 ± 0.02 74 294

Froude RF 0.0 - 0.34 0.10 ± 0.01 74 39 NF 0.0 - 0.77 0.11 ± 0.01 122.1 293 <0.0001 67.51 NF+D 0.0 - 0.61 0.19 ± 0.01 69.7 294

42 Table 2 continued.

Reynolds number RF 0.0 - 317 79 ± 10.59 84 40 NF 0.0 – 1380 174 ± 14.12 138.2 293 <0.0001 36.56 NF+D 0.0 – 952 294 ± 12.61 86.7 294

Reach Discharge (m3 s-1) RF 0.002 - 0.12 0.03 ± 0.02 188.3 5 NF 0.94 - 1.22 1.09 ± 0.06 10.9 4 0.016 8.27 NF+D 0.56 - 1.53 0.92 ± 0.31 58.9 3

USGS Discharge (m3 s-1) RF n/a n/a n/a n/a NF 0.91 - 2.52 1.64 ± 0.23 36.5 7 0.06 *U=9.5 NF+D 0.68 - 1.84 1.08 ± 0.15 37.5 7

Note: When only two groups were compared a Student’s t-test was used between upstream and downstream study sites, and Mann Whitney U-test between NF and NF+D treatments.

43 Table 3: Three types of snail capture data: a) Number of new captures, identified as snails that had not previously been captured, and the corresponding percentage of the total number of snails released on the original release date, b) Total number of captures for each treatment and capture rates calculated as the percentage of the total number of snails released on the original release date found on each search date and c) Total number of captured snails found downstream or at the release site (neutral) and its percentage of the overall total captured. n/a indicates no data. See Table 2 for explanation of treatments.

Days After Release Capture type Treatment 2 3 5 6 16 a.) New captures RF 75 46 50 33 n/a Percentage 100 57 56 30

NF n/a 115 81 65 28 100 32 20 21

NF+D 117 84 67 n/a n/a 100 39 28

b.) Total captures RF 78 46 54 33 n/a Percentage 39 23 27 17

NF n/a 116 81 66 28 59 41 33 14

NF+D 128 89 77 n/a n/a 65 45 39

44 Table 3: continued.

c.) Downstream/Neutral total captures RF 60 23 12 8 n/a Percentage 77 50 22 24

NF n/a 12 3 2 n/a 10 4 3

NF+D 15 0 3 n/a n/a 12 0 4

45 -1 Table 4: Snail Upstream Migration Rate (UMR as m d ), under historic discharge quantiles, Q50, Q70, Q90, and minimum discharge needed to reach the ocean (Qmin) (based on Oki [2007]) in three West Maui streams, were calculated using the discharge – migration rate regression equation (y = 0.43x + 0.45, R2 = 0.23, P = 0.16). The “Individual Snail UMR Maximum” and “Mean UMR Maximum Potential” calculated from ‘rapid’ snails are provided. The approximate distance from the ocean to the highest altitude diversion in each stream are as follows: Iao = 6900 m, Waihee = 4500 m, N. Waiehu = 3600 m, S. Waiehu = 4000 m. Based on these distances, the time necessary to migrate beyond these diversions was estimated, in days and years, for all upstream migration rates.

Mean UMR Discharge - Migration Regression Equation Individual Snail UMR Maximum Maximum Potential RF NF NF+D Rapid

Stream Measure Q50 Q70 Q90 Qmin Treatment Treatment Treatment Snails

Iao UMR 0.92 0.79 0.69 0.55 5.6 4.9 3.77 2.18 Days 7492 8744 9929 12592 1232 1408 1830 3165 Years 20.5 23.9 27.2 34.5 3.4 3.9 5 8.7

Waihee UMR 1.09 1 0.9 0.47 5.6 4.9 3.77 2.18 Days 4126 4517 4988 9598 804 918 1194 2064 Years 11.3 12.4 13.7 26.3 2.2 2.5 3.3 5.7

Waiehu N. UMR 0.51 0.5 0.49 0.47 5.6 4.9 3.77 2.18 Days 7016 7242 7368 7587 643 735 955 1651 Years 19.2 19.8 20.2 20.8 1.8 2 2.6 4.5

Waiehu S. UMR 0.51 0.49 0.48 0.47 5.6 4.9 3.77 2.18 Days 7810 8093 8315 8498 714 816 1061 1835 Years 21.4 22.2 22.8 23.3 1.9 2.2 2.9 5

46 Table 5: Migration as described in previous published studies compared to results of the current study. The organism of interest, study location and a brief explanation of study observations are provided. All non-aggregate and aggregate mean rates were calculated using details within authors’ observations. Results in cm min-1 were extrapolated to m d-1 and assuming migration takes place 8-12 hours day-1 (Benbow et al., 2002). Migration time frames, in days, were estimated using stream distance to highest diversion (in parentheses) and calculated migration rates.

Calculated Migration Days to migrate beyond highest (m d-1) diversion Migration Author Organism, Observation Mean N. S. Type location Migration 8 hrs of 12 hrs of Iao Waihee Waiehu Waiehu Rate migration migration 6900m 4500m (3600m) (4000m) (m d-1)

Non- Gorbach N. three 0.25 under aggregate et al. granosa, treatments, reduced -- -- 27600 18000 14400 16000 (2012) Maui, 800 flow Current Hawaii observations 0.66 under Study of 590 tagged natural -- -- 10455 6818 5455 6061 snails over 6 flow days 1.16 under following natural release; 3 – 7 -- -- 5948 3879 3103 3448 flow + mm in shell density length

47 Table 5: continued.

Non- Pyron & Neritina 1140 0.153 mean aggregate Covich, punctulata, observations of upstream -- -- 45098 29412 23529 26144 (2003) Puerto 274 tagged rate Rico snails over 15 0.81 weeks; 24 mm greatest -- -- 8519 5556 4444 4938 in shell length mean rate 2.38 maximum -- -- 2899 1891 1513 1681 upstream rate

Non- Brasher, N. 769 tagged aggregate (1997) granosa, snails Hawaii recaptured after 0.7 -- -- 20909 13636 10909 12121 1 month; 11-13 mm in shell length 762 tagged snails recaptured after 0.17 -- -- 40588 26471 21176 23529 2 months, 16-33 mm in shell length

48 Table 5: continued.

Aggregate Pyron & N. ~307 snails, 7.4 -- -- 932 608 486 541 Covich, punctulata, ~12 mm in (2003) Puerto Rico shell length

Aggregate Schneider Neritina >500,000 50 (ranged & Lyons, latissima, snails, <7 mm -- -- 138 90 72 80 30 -110) (1994) Costa Rica in shell length

Following Ford, N. granosa, 80 snails in a in narrow (1979) Hawaii chain, 274 - 179 - 143 - 159 - 50.4 16.8 25.2 line !5 mm in shell 411 268 214 238 length

Following Schneider N. latissima, 140 snails/m, in narrow & Frost, Costa Rica trail 32m long, 76 - 50 - 44 - 181.44 60.48 90.72 40 - 60 line (1986) 2-6 mm in shell 114 74 66 length

Following Schneider N. latissima, 20 snails in a in narrow & Lyons, Costa Rica chain moving 2 55 - 36 - 32 - 250.56 83.52 125.28 29 - 43 line (1993) cm, <7 mm in 83 54 48 shell length

Following Pyron & N.punctulata, 20 snails in a 114 - 74 - 66 - in narrow Covich, Puerto chain, ~12 mm 120.96 40.32 60.48 60 - 89 line (2003) Rico in shell length 171 112 99

49 !

Figure 1: Map of the Hawaiian Islands with Maui and study streams identified. a) Iao Stream in the West Maui Mountains and b) Honomanu Stream of East Maui. Insets highlight study reaches (black dots) and stream diversions (black rectangle).

! 50 ! Figure 2: Individual snail captures made during each treatment ! 16 days after release plotted on an x, y coordinate system (release point = 0,0): a) Reduced flow (RF), b) Natural flow (NF), and c) Natural flow + increased density (NF+D). The number (N) of snails captured on each day is provided in Table 3. Similar symbols represent captures across treatments: 2d = ܆, 3d = !, 5d = !, 6d = ", and 16d = # and the $ in NF+D treatment represent captured untagged snails of the density effect. !

! 51 !

! Figure 3: Mean (SE) Euclidean migration rates (EMR; black bar) and Upstream migration rates (UMR; white bar) between treatment conditions. UMR and EMR in the RF treatment were not significantly different, ns (P>0.05). EMR and UMR were significantly different within the NF and NF+D treatments, and when each migration rate was compared among treatments (Kruskal-Wallis H =203.0, P < 0.0001; H = 122.0, P < 0.0001; respectively). Different letters indicate significant differences between columns (P < 0.05). The number of observations (N) for the RF, NF and NF+D treatments were 204, 260 and 268, respectively.

! 52

Figure 4: a) Mean upstream migration rate (UMR) over all initial (! 6 days) search days for each treatment – RF (black), NF (grey) and NF+D (white). Negative UMR represents downstream movement. RF and NF+D UMR were significantly different on 2 d (Mann Whitney U = 642.5, P < 0.0001), and likewise, RF and NF were significantly different on 6 d (U = 776.0, P = 0.02). Mean UMR of the RF treatment differed over all days (H = 50.73, P < 0.0001), whereas in the other treatments, UMR was consistent (NF: H = 0.02, P = 0.99; NF+D: H = 3.87, P = 0.14). b) Mean (SE) upstream migration rates for RF, NF and NF+D treatments with captured snails that traveled ! 8m in ! 6 days (H = 131.1, P<0.0001), pooled longer-term captures made 16, 33 and 63 post-release, and ‘rapid’ snail captures that traveled " 8m in ! 6 days. Rapid mean UMR was significantly greater than snails that traveled ! 8m in ! 6 days (H = 165.2, P <0.0001) and the pooled longer- term mean UMR (U = 33.0, P <0.0001). Different letters indicate significant differences between columns (P <0.05). !

! 53 APPENDICES

METHODS

Habitat variables measured (or calculated) on each search day for each snail capture and included the following: discharge (Q), water column depth (D), surface flow velocity (SV), mid-column velocity (MCV), Froude (F), Reynolds number (R), and distance to nearest neighbour (NN). Additional covariates included each of these measured variables on the previous recapture day of a particular snail (denoted with p).

For example, pDischarge (pD) is the discharge measured on the previous recapture day.

These variable codes were used in subsequent multivariate analyses and corresponding tables.

A series of analytical steps, using JMP Version 4.0 (SAS Institute Inc., Cary, NC,

U.S.A.), was employed to test for significant relationships between habitat variables and upstream migration rates for each treatment and when treatments were pooled, representing a gradient of stream and migratory conditions. First, correlation matrices tested for multicollinearity and correlation coefficients summarized the strength of the linear relationship between each pair of response variables for each treatment, over all recovery days. Mahalanobis and Jackknife distances identified snail capture data that violated the correlation structure. These data were considered outliers and removed from further analyses (Jolliffe, 2002; Bayley & Prather, 2003; Venturelli & Tonn, 2005). For each treatment and when pooling all treatments, Principal Component Analysis (PCA) was performed on the correlated habitat variables to summarize the structure of intercorrelation and reduce the number of environmental variables (Graham, 2003; Roy et al., 2003; Lehman et al., 2005). The number of components retained was decided using

! 54 the strategy of Lehman et al. (2005). Varimax factor rotation was performed on the retained principal components and the new rotated factor loadings identified variables that loaded heavily on each component, which were then used to determine the construct measured by each component (Lehman et al., 2005). Standard least square linear regression models tested the relationship between the factor-rotated components and the four possible calculated response variables – EMR from the release point, EMR from the point of previous capture, UMR from the release point, and UMR from the point of the previous capture. The response variable best explained by the factor-rotated components was determined by comparing the adjusted R2 among the four models. Using the PCA factor-rotated components as explanatory covariates, and interactions among them, we developed a set of candidate models to relate upstream migration rate to habitat template covariates. We used generalized linear modeling (GLM) with an identity link function and maximum likelihood techniques to determine the best-fit model, explaining a significant portion of deviance between the response and explanatory covariates (Guisan,

Edwards & Hastie, 2002). Candidate models, differing in combinations of explanatory covariates, were compared using the Akaike Information Criteria adjusted for small sample size (AICc). The model with the lowest AICc value was accepted as the best-fit model of the candidate models considered (Akaike, 1973; Burnham & Anderson, 2002).

To account for any remaining variation not explained by the factor-rotated components, a standard least squares regression was used to examine the relationship between the studentized deviance residuals and the distance recaptured snails moved perpendicular to flow (x-coordinate), as a measure indicating usage of the thalweg and spatial aggregation

! 55 relative to the stream bank (i.e., how spread across the channel snails were as they migrated upstream).

RESULTS

For each treatment and when treatments were pooled, relationships between in- stream habitat template covariates and upstream migration rates were tested using a series of analytical steps. Individual captures that were identified as outliers, using Jackknife and Mahalanobis distances, were removed from each treatment (14 from RF, 10 from NF,

7 from NF+D treatment, and 23 from the pooled treatments). Correlation matrices determined strong multicollinearity among covariates and therefore Principal Component

Analysis (PCA) was used to summarize the structure of intercorrelation and reduce the number of environmental variables (Graham, 2003; Roy et al., 2003; Lehman et al.,

2005). The number of components retained in each PCA was determined and cumulatively the amount of variation in the response variable explained by the retained components in the PCA ordination was 77% in the RF treatment, 86% in NF, 83% in

NF+D and 80% when treatments were pooled (Appendix S2). Factor rotation was performed on the retained principal components and the new rotated factor loadings identified covariates that loaded heavily on each component. Based on the nature of these covariates, a construct was named for each component, depicting the biological measure described by the covariates (Lehman et al., 2005). The construct, “habitat-scale hydraulics (HH),” was used to define the biological nature of covariates including SV,

MCV, F, and R; the construct, “reach-scale hydraulics (RH),” described the discharge covariate; and the construct, “aggregation (A),” was used to define the distance to the nearest neighbor calculated from the spatial distribution of captures. Water column depth

! 56 did not load heavily on any retained component and was not described by a construct.

Not only were these constructs used to describe the covariates measured on each day of capture, but they also indicated the covariates measured on the previous day of capture.

Accordingly, we also evaluated habitat-scale hydraulics and reach-scale hydraulics measured on the previous day of capture (pHH and pRH) as constructs (Appendix S2).

The rotated factor component scores, identified by their construct names, were used as covariates. The adjusted coefficients of determination (R2) of linear regressions between the rotated factor component scores and the four possible response variables identified

EMR measured from the origin as the response variable that was used in subsequent analyses in all treatments except in the RF, where UMR from the origin was used.

Generalized linear modeling (GLM) tested candidate models by comparing AICc values (Appendix S3). Of the candidate models, AICc ranged from 313.6 to 304.3 in the

RF treatment, 377.5 to 366.3 in the NF treatment, 581.2 to 577.2 in the NF+D treatment, and 1797.2 to 1795.2 when treatments were pooled (Appendix S3). Although a general trend was not seen across all treatments, the models with the lowest AICc contained the constructs HH, pHH, RH and pRH and their interactions for NF and NF+D, while those for RF only included HH and RH (Appendix S4). As predicted, habitat-scale hydraulics was positively correlated with migration rate across all treatments. Further, HH and RH had the highest weight (parameter estimates = 0.596 and 0.510, respectively) for predicting migration rate when the treatments were pooled. From 3 – 31% of the residual variation in upstream migration was further explained by snail spatial configuration relative to the stream bank (x-coordinates). Even though these values are relatively small, usage of the thalweg appeared to influence upstream migration rate.

! 57 REFERENCES

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! !

! 58 Appendix S2: Principal component analysis results for each treatment and when pooled. Eigenvalues, the percent variance explained by each component, and the cumulative percent variance captured by the retained components are provided. Each component was identified by a construct that captured the biological significance of the covariates that loaded heavily on the component. These are habitat-scale hydraulics (HH), reach-scale hydraulics (RH) and aggregation (A). The letter “p” is added for that construct measured on the day previous to capture. See Table 2 for explanation of Treatments.

% of Cumulative Component Treatment Component Eigenvalue variance % construct RF 1 4.35 33.5 33.5 HH 2 4.05 31.1 64.6 pHH 3 1.61 12.4 76.9 RH

NF 1 5.67 43.7 43.7 pHH 2 3.24 24.9 68.6 HH 3 1.30 9.9 78.5 pRH 4 1.08 8.3 86.9 A

NF+D 1 5.67 43.6 43.6 pHH 2 3.77 29.0 72.6 HH 3 1.30 10.0 82.7 pRH

Pooled 1 5.88 45.2 45.2 pHH 2 3.31 25.5 70.7 HH 3 1.16 8.9 79.6 RH

59 Appendix S3: Generalized linear model results examining the effect of habitat template covariates on upstream migration rate. Chi- square test statistic (!2), K as the number of covariates in each model, the probability of significance, and overdispersion for each candidate model are provided. Corrected AIC values were used to compare candidate models; model with the lowest AICc was considered the best-fit model for each treatment. See Table 2 for explanation of Treatments. Candidate models were arbitrarily given letter identification a – h.

Treatment Candidate Model Model !2 K Prob>!2 overdispersion AICc RF a 68.03 2 <0.0001 0.29 304.31 b 68.21 3 <0.0001 0.29 306.14 c 68.49 4 <0.0001 0.29 308.10 d 68.53 5 <0.0001 0.29 310.22 e 69.94 6 <0.0001 0.29 311.99 f 69.55 7 <0.0001 0.29 313.59

NF a 120.06 8 <0.0001 0.29 366.30 b 121.75 9 <0.0001 0.29 366.83 c 122.70 10 <0.0001 0.29 368.13 d 123.31 11 <0.0001 0.29 369.78 e 124.16 12 <0.0001 0.28 371.21 f 124.64 13 <0.0001 0.28 373.05 g 124.75 14 <0.0001 0.28 375.27 h 124.90 15 <0.0001 0.28 377.48

NF+D a 178.30 5 <0.0001 0.56 577.19 b 178.40 6 <0.0001 0.56 579.22 c 178.55 7 <0.0001 0.56 581.23

Pooled a 385.39 6 <0.0001 0.87 1795.17 b 385.46 7 <0.0001 0.87 1797.16

60 Appendix S4: Best-fit generalized linear model results for each treatment and when treatments were pooled – !2test statistic (!2), K as the number of covariates in each model, the probability of significance, overdispersion, and AICc for each best-fit model. Each covariate included in the best-fit model along with their corresponding parameter estimates, standard error (SE) and probability of significance are provided. See Table 2 for explanation of Treatments. Note: Habitat-scale hydraulic covariate (HH), reach-scale hydraulic covariate (RH), aggregation covariate (A) and the letter “p” is added when referring to that covariate measured on the previous day of capture.

Over- Parameter Treatment Model !2 K Prob>!2 dispersion AICc Model Covariates Estimate SE Prob>!2 RF 68.03 2 <0.0001 0.29 304 Intercept 0.18 0.04 <.0001 HH 0.31 0.04 <.0001 RH -0.19 0.04 <.0001

NF 120.06 8 <0.0001 0.29 366 Intercept 1.05 0.04 <.0001 pHH -0.08 0.04 0.067 HH 0.08 0.04 0.027 pRH 0.14 0.04 0 A -0.38 0.04 <.0001 (HH-0.009)*(A-0.024) -0.19 0.05 0

(pHH+0.050)*(HH- 0.13 0.08 0.096 0.009)*(A-0.024)

(HH- 0.23 0.07 0.001 0.009)*(pRH+0.030)*(A- 0.024) (pHH+0.050)*(HH- 0.27 0.12 0.024 0.009)* (pRH+0.030)*(A-0.024)

61 Appendix S4: continued.

NF+D 178.3 5 <0.0001 0.56 577 Intercept 2 0.05 <.0001 HH 0.67 0.05 <.0001 pHH*HH -0.29 0.06 <.0001 pRH -0.09 0.06 0.1 pHH*pRH 0.17 0.07 0.013 pHH*HH*pRH 0.29 0.06 <.0001

Pooled 385.39 6 <0.0001 0.87 1795 Intercept 1.15 0.04 <.0001 pHH 0.22 0.04 <.0001 HH 0.6 0.04 <.0001 (pHH+0.021)*(HH- -0.13 0.04 0 0.006) RH 0.51 0.05 <.0001 (pHH+0.021)*(RH- -0.24 0.07 0 0.001) (HH-0.006)*(RH- 0.25 0.07 0 0.001)

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! 68! ! !

CHAPTER III

VARIABILITY IN HABITAT TEMPLATE AND BENTHIC COMMUNITY

RESPONSE TO ANTHROPOGENIC WATER REMOVAL IN TROPICAL

MOUNTAIN STREAMS

ABSTRACT

Mountain streams that originally supported Hawaiian cultural practices have been diverted for development, agriculture and tourism for over 150 years. Habitat characteristics and benthic macroinvertebrate community responses to water withdrawal were studied in four West Maui Mountain watersheds. We compared riffle and cascade habitats upstream and downstream of the highest elevation diversion in each watershed.

Riffles were shallow areas with moderate flow; while cascades were identified as high velocity water flowing over boulders, and separated into torrenticolous and amphibious microhabitats. Among streams, downstream discharge was reduced by 84 – 99%. Flow velocity was 4x greater upstream and depth was 50% lower downstream, corresponding to reductions in both downstream habitats. There was a significant 44% reduction in macroinvertebrate density downstream of diversions (t = 3.261, df = 136, p = 0.0014); however, density was not significant different among streams (F = 1.95, df = 3, p =

0.125). When density was corrected for habitat availability we found significantly greater proportions of native taxa in amphibious microhabitats compared to riffle and

!69! ! ! ! torrenticolous habitats. Non-native Chironomidae and Trichoptera (Cheumatopsyche sp. and Hydroptila sp.) were dominant (> 95%) and ubiquitous in riffle habitats, whereas native Limonia sp. dominated (30%) amphibious microhabitats. Macroinvertebrate community structure varied among streams, sites and microhabitats indicating inconsistency in response to water withdrawal, dependent upon watershed size and microhabitat conditions. Our findings contribute to ongoing water management and restoration efforts focused on conservation of native species and habitat integrity in tropical streams worldwide.

INTRODUCTION

Ecosystems worldwide are threatened by increased demands on freshwater resources due to growing populations and associated consumption (Postel, 1997;

Benstead et al., 1999; March et al., 2003). To account for these demands, many streams and rivers have undergone hydrologic alterations such as dams, surface water diversions, stream channelization, and inter-catchment water transfer (Rosenberg et al., 2000;

Dudgeon, 2000; Baker et al., 2011). These hydrologic modifications include any anthropogenic disruption in the magnitude or timing of river discharge; thus changing the natural flow regime (Standford et al., 1996; Hart & Finelli, 1999; Rosenberg et al., 2000) and impacting the physical, chemical and biological structure and function of rivers and streams (Ward & Standford, 1983; Allan, 1995).

The reduction or elimination of flow downstream of diversions has been shown to have significant impacts on native and amphidromous (a form of diadromous reproduction) stream organisms of many tropical regions by disrupting the connection between upstream habitats and the ocean, therefore preventing larval drift to the ocean

!70! ! ! ! and obstructing postlarval return migration (Timbol & Maciolek, 1978; Ford & Kinzie,

1982; Kinzie & Ford, 1982; Maciolek & Ford, 1987; Kinzie, 1990; Resh et al., 1992;

Benstead et al., 1999). Removal of streamflow can also alter downstream habitat which affects physical and chemical conditions that regulate macroinvertebrate insect communities; thus, limitations in habitat availability and/or quality may shift species interactions, threaten native and endemic aquatic populations and potentially increase the number of invasive species (Cowx et al., 1984; Poff et al., 1997; Brasher, 2003). This is of particular concern in Hawaii because several native and endemic species are sensitive to changing environmental conditions and community impacts of introduced species

(Shoda et al., 2010).

Much of previous flow alteration research has focused on the direct downstream effects of dam construction and flow diversion on habitat conditions, invertebrate community structure, and biotic interactions (Rosenberg et al., 2000; Brasher, 2003;

Dewson et al., 2007; Miller et al., 2007); however, there are also upstream effects on migratory fauna and related basal food resources and invertebrate assemblages (for full review see Benstead et al., 1999 and Greathouse et al., 2006). While the impacts of dams have received some attention, few studies have assessed the direct physical and hydraulic effects of flow diversion on stream communities (Baker et al., 2011) even though the construction of small low head dams have undergone substantial proliferation in tropical watersheds (Benstead et al., 1999). Further, hydrologic alterations and subsequent ecosystem changes have been extensively studied in temperate regions; whereas, far less is known about these impacts in tropical regions, especially islands, leaving these

!71! ! ! ! ecosystems poorly understood (Benke et al., 1988; Benstead et al., 1999; Pringle et al.,

2000; March et al., 2003; McIntosh et al., 2008; Shoda et al., 2010).

The increasing development and urbanization of tropical island nations will continue to strain freshwater resources (Jackson & Sweeney, 1995; Benstead et al., 1999;

Smith et al., 2003). Hydrologic changes do not occur in isolation but interact with other threats to biodiversity, such as introduced species and water quality degradation

(Dudgeon, 2000; Brasher, 2003). Although the study of tropical streams has grown, further studies of biodiversity are necessary for continued and improved conservation efforts (Jackson & Sweeney, 1995; Smith et al., 2003; Kinzie et al., 2006). Further, it is not yet clear how ecological theory derived from studies in temperate continental stream ecosystems can be effectively applied to understand and manage tropical stream systems

(Smith et al., 2003).

As the most isolated island archipelago in the world, freshwater in the Hawaiian

Islands is naturally limited to surface streams and groundwater supplied by heavy tropical rains (Fitzsimons et al., 1997; Brasher, 2003; March et al., 2003). To original Hawaiian civilizations, water resources were culturally important and used in the native practice of taro cultivation. Flowing water was also important to the habitat of native stream macrofauna, such as fish, shrimp, snails and prawns that were harvested as a food source.

However, through Western colonization and associated development of large-scale commercial sugarcane plantations in the mid-1800s, stream diversions and extensive tunnel systems were built, transporting freshwater from the wet, windward watersheds to the dry leeward areas for agricultural production and population growth and development. At least 58% of the estimated 366 Hawaiian perennial streams had

!72! ! ! ! experienced some type of streamflow alteration by 1978, with water being exploited for anthropogenic uses, such as agriculture, development and tourism (Parrish et al., 1978;

Hawaii Cooperative Park Service, 1990). Depending on diversion structure and capacity, water removal varies substantially across watersheds. Many of these diversions remove from 90 – 100% of base flow volume, altering the natural flow regime which is important for sustaining native biodiversity and ecosystem integrity (Wilcox, 1996; Poff et al.,

1997; Benbow, 1999; McIntosh et al., 2002; Brasher, 2003; McIntosh et al., 2008).

The native fauna of Hawaiian freshwater streams is represented by relatively few species (Brasher, 2003; Font, 2003). Only 100 – 150 of the 795 species of Hawaiian aquatic insects and mites inhabit freshwater streams, of which more than 85% are endemic (Resh & Szalay, 1995). The native macroinvertebrate species richness that represents island streams is lower than continental streams. The native Hawaiian stream fauna composition is remarkably different than typical continental streams, with no native species of Ephemeroptera, Megaloptera, Plecoptera, or Trichoptera (Resh et al., 1990;

Howarth & Polhemus, 1991; Larned et al., 2003). Further, these streams have been invaded by introduced species, which now far outnumber populations of native fauna

(Devick, 1991; Cowie, 1997; Brasher, 2003; Shoda et al., 2010). Unfortunately, extensive stream alterations throughout the Hawaiian Islands have created degraded conditions that are more suitable for introduced species than the original native species (Maciolek, 1978;

Brown et al., 1999; Brasher, 2003). These introduced species have broad environmental tolerances, whereas, native species evolved in watershed conditions with natural flow regimes that are remarkably stochastic (Norton et al., 1978; Timbol & Maciolek, 1978).

Not only can introduced species prey on, and often outcompete native species (Brasher,

!73! ! ! !

2003), but these effects are exacerbated by disrupted flow regimes that have differentially more negative effects on at least some native macroinvertebrates (Brown et al., 1999;

McIntosh et al., 2003; Shoda et al., 2010; Gorbach et al., 2012).

This study identified the effects of water removal on habitat template characteristics and biological communities in four streams of the West Maui Mountains,

Hawaii, USA. We hypothesized that streamflow removal would negatively affect riffle habitat template characteristics with negative effects on macroinvertebrate community structure. Further, we compared our results from riffle habitats to a companion study of cascade habitats and amphibious and torrenticolous microhabitats (Shoda et al., 2010) to identify native and introduced species response differences among habitats. We predicted that stream diversion effects on the habitat template would be associated with shifts in macroinvertebrate density and community composition and that these effects would vary by watershed size and extent of water withdrawal.

METHODS

Study sites

This investigation took place in the four watersheds of the West Maui Mountains, collectively known in Hawaiian as the N! Wai ‘Eh!: Waikapu Stream, Iao Stream,

Waiehu Stream (North Branch) and Waihe’e River (Fig. 1; Oki et al., 2010). These naturally perennial, 3rd order, windward streams have a large proportion of high velocity turbulent flow habitats under natural conditions, but have been extensively diverted to transport water to the dry, more arid leeward side of Maui where agriculture, development and tourism dominate land use. The existing diversions on these streams have the combined capacity to remove 75 million gallons of water per day (3.29 m3/s);

!74! ! ! ! often removing all, or nearly all, of base flow (Oki et al., 2010). The summit of the West

Maui Mountains receives > 8.89 m of annual rainfall, and supplies the headwaters of Iao

Stream and Waihe’e River, two of the largest streams on Maui based on flow volume

(Oki et al., 2010). Oki (2007) described mean base flow discharge (Q70) for each of these streams during climate years 1984 – 2005. Q70 discharge represents the daily streamflow volume that is equaled to or exceeded 70% of the time in historical records, and was defined as base flow conditions for Hawaiian streams (Oki et al., 2010). For the study

3 watersheds, the Q70 discharge was as follows: 1.27 m /s (Waihe’e River), 0.101 – 0.118 m3/s (N. Waiehu Stream), 0.788 m3/s (Iao Stream), 0.171 – 0.227 m3/s (Waikapu

Stream). Thus, to evaluate watershed size differences, reflected by base flow, the four streams were divided into ‘small’ – Waiehu Stream and Waikapu Stream – and ‘large’ –

Iao Stream and Waihe’e River.

To study the effects of water removal in each watershed, study sites were established as stream reaches 500 m upstream (US) and downstream (DS) of the highest elevation diversion in each stream, and identified through collaboration with the U.S.

Geological Survey (USGS) who conducted hydrologic surveys of these watersheds during the same study period. Habitat template characteristics (e.g. flow velocity, water depth, channel width, and substrate size) associated with randomly sampled biological samples (e.g., macroinvertebrates, chlorophyll a, etc. see below) were taken from a defined 100 m study reach at each site, which was further divided into 10 transects perpendicular to flow and representative of all habitat types (i.e. riffle, pool, run, cascade). For a further detailed description of these watersheds and hydrological conditions refer to Oki et al. (2010) and Shoda et al. (2010). Based on funding, logistical

!75! ! ! ! and accessibility constraints, all sites were sampled twice in August 2007 and once in

May 2008, which was similar to sampling frequency for other published bioassessment studies of Hawaiian streams (McIntosh et al., 2002, 2003, 2008; Brasher et al., 2004;

Wolff, 2005; Shoda et al., 2010). !

Physical habitat template

The physical habitat of each reach was characterized and described in detail by

Shoda et al. (2010) and included reach discharge, wetted channel width, percent canopy cover, substrate size, reach gradient, habitat extent, water temperature and dissolved oxygen. In addition to discharge measurements made in the field, undiverted discharge data were obtained from USGS streamgages upstream of the highest diversions in Iao

Stream (station #16604500) and Waihe’e River (station #16614000); there were no functional gages on the other streams. Water column depth and flow velocity (at 0.6 depth) were measured every 0.5 m along each transect. Discharge and flow velocity were measured using a SonTek Doppler Flowtracker (SonTek/YSI, 2005). We measured

Seston fine particulate organic matter (FPOM), chlorophyll a and the ratio of epilithic biomass to chlorophyll a, or the Autotrophic Index (Kinzie et al., 2006; Hauer &

Lamberti, 2006). We quantified habitat availability by measuring the area of all geomorphic channel units within the 100 m study reach using a range finder and measuring tape.

Macroinvertebrate sampling

Six benthic samples were randomly collected from riffles at each site using a modified Surber sampler (0.0625 m2) and methods described by McIntosh et al. (2002,

!76! ! ! !

2003, 2008). Prior to each benthic sample, we measured water column depth and flow velocity profiles (at 0.2, 0.6 and 0.8 depth) at four equidistant points within the sampled area. In a companion study (Shoda et al., 2010), we identified six cascade habitats, in which two microhabitats were randomly sampled from each – torrenticolous (submerged habitat) and amphibious (wetted splash zones on adjacent exposed rock) – using methods of Benbow et al. (1997, 2003, 2005) and Shoda et al. (2010). We preserved all macroinvertebrate samples in 70-90% ethanol for laboratory sorting and identification.

Macroinvertebrates were counted and identified to genus level; however, order and family resolution was used when genus was not practical because of damaged specimens or if keys were unavailable (Terry, 1913; Denning & Beardsley, 1967;

Denning & Blickle, 1971; Beardsley et al., 1998; Zimmerman, 2001; Merritt et al., 2008;

Shoda et al., 2010). We pooled members of Chironomidae, excluding Telmatogeton sp., which was easily identifiable as an endemic midge (Benbow, 2008; Shoda et al., 2010) and verified identifications using voucher specimens identified by EcoAnyalyst Inc.

Based on the findings of previous studies (Williams, 1983; Polhemus, 1995; Englund et al., 2007), and similar to Shoda et al. (2010) we classified all members of Procanace,

Limonia, Ephydridae, Telmatogeton sp., Hyposmocoma sp., Megalagrion sp. and Atyoida bisulcata as endemic to the Hawaiian Islands.

Statistical analysis

Total macroinvertebrate density and habitat-corrected reach densities, determined by multiplying the macroinvertebrate density by the total area of available riffle habitat within each reach, were used to describe macroinvertebrate abundance, whereas

Simpson’s Diversity Index expressed the extent of biodiversity at each site. Cascade

!77! ! ! ! habitats and their corresponding microhabitats were analyzed in full detail in Shoda et al.

(2010). For purposes of this paper, cascade data from Shoda et al. (2010) were included to investigate density and taxa response differences between cascade and riffle habitats, which was not previously done. We calculated an Index of Nativity (I.N.), described as the ratio of native to introduced taxa densities, to describe differential responses of native and introduced taxa to reduced discharge. All data were appropriately transformed to meet assumptions of normality and homogeneity of variances when possible; however, when transformations were not effective, we used nonparametric analyses. A student’s t- test evaluated the effect of diversion between upstream and downstream sites when total densities and habitat-corrected reach densities were pooled among streams. We performed two-way ANOVA with Bonferroni post-tests using GraphPad Prism 5.0

(GraphPad Software) to test differences among streams, between upstream and downstream sites and the interaction for mean macroinvertebrate densities, habitat- corrected reach densities, Simpson’s Diversity Index, I.N., microhabitat-specific densities and habitat template variables, including FPOM, chlorophyll a, reach discharge, and transect flow velocities and depths. Because similar hypotheses were tested for each of these response variables, a Bonferroni adjusted p-value = 0.005 was used to interpret statistically significant main effects. Similar analyses of additional habitat template variables, including wetted channel width, percent canopy cover, substrate size, and reach gradient were presented in Shoda et al. (2010, Table 1).

Macroinvertebrate community composition differences among streams, between sites and between riffle and cascade microhabitats were evaluated using Non-metric multidimensional scaling (NMDS) followed by multi-response permutation procedures

!78! ! ! !

(MRPP) using PC-Ord (McCune & Mefford, 1999). The Sørensen distance measure was used with a random starting configuration and 250 runs with real data (McCune & Grace,

2002). We identified indicator taxa for each habitat or microhabitat and between sites using Indicator Species Analysis as described by in McCune & Grace (2002), and results from Monte Carlo tests with 4999 randomizations were used to determine significant differences in community structure similar to the multivariate approach by Shoda et al.

(2010).

RESULTS

Physical habitat template

Water removal significantly altered the physical habitat template of the downstream sites. Mean daily discharge recorded during the study by USGS streamgages in Iao Stream and Waihe’e River was 1.23 m3/s and 0.97 m3/s, respectively. When we measured instantaneous discharge, there were significant differences among streams and sites (Table I). Discharge was reduced by 84 – 99% downstream of the diversions and did not statistically differ among downstream reaches of the streams. Upstream discharge significantly varied among streams: Iao Stream and Waihe’e River was significantly greater than both small streams, with Waihe’e River significantly greater than Iao Stream

(t = 5.087, p < 0.001), while the two small streams did not significantly differ (t = 0.088, p > 0.05). When the two large streams were pooled, lower discharge below the diversions was associated with significantly reduced riffle and cascade habitat, (US to DS, riffle:

29.0% to 6.0% of total reach area and cascade: 20.6% to 3.5% of total reach area). For the two small streams, cascade habitat decreased (17.0% to 14.6%) but riffle habitat increased (14.6% to 22.6%) downstream of the diversion (Fig. 3).

!79! ! ! !

Streamflow ranged from 0.004 m3/s downstream in Waikapu Stream and Waiehu

Stream to 1.33 m3/s upstream in Waihe’e River (Fig. 2). Mean flow velocity and depth measured along reach transects were also significantly different among streams and sites with velocity following trends of measured discharge and depth significantly reduced downstream in all streams (Table I). Nonparametric correlation analysis found a significant positive relationship between percent cascade habitat and depth (Spearman r =

0.833, p = 0.015). However, there were no other significant relationships between percent habitat (riffle and cascade) and depth and velocity.

There was a significant stream effect (F = 7.297, df = 3, p = 0.0007; Table I) on

FPOM (mg/L); however, the effect of diversion on FPOM varied among watersheds.

There were no significant differences among the upstream sites, while downstream

FPOM in Waihe’e River was significantly greater than all other downstream sites.

Further, upstream FPOM in Iao Stream was significantly greater than downstream (t =

2.97, p < 0.05). There were significant main effects of stream and site (Table I) on chlorophyll a (!g/L), with upstream significantly greater than downstream in Iao Stream

(t = 2.94, p < 0.05). The Autotrophic Index was greater downstream in all streams (Table

I); among all downstream sites, the Autotrophic Index in Waikapu Stream was significantly greater than Waihe’e River (t = 2.45, p < 0.05).

Macroinvertebrate density and diversity

There was a significant 44% reduction in macroinvertebrate density in riffle habitats downstream of diversions when streams and sampling dates were pooled (t =

3.261, df = 136, p = 0.0014). When the effects of water removal on macroinvertebrate density were analyzed for individual streams, there was a significant effect of site (F =

!80! ! ! !

11.49, df = 1, p = 0.0009; Table I, Figure 4a); however, this was not the case among streams (F = 1.95, df = 3, p = 0.125). Within Waihe’e River, there was a significant reduction in mean macroinvertebrate density downstream of the diversion (t = 4.92, p <

0.0001). Further, there were significantly greater densities in the upstream site in Waihe’e

River compared to the two smaller streams: Waikapu Stream (t = 2.95, p < 0.01) and

Waiehu Stream (t = 4.12, p < 0.001); but the upstream densities in Iao Stream were not significantly different from the other upstream sites. There was no significant difference in macroinvertebrate density among downstream sites.

When densities were corrected for available riffle habitat within each reach, there were significant stream (F = 24.43, df = 3, p < 0.0001; Table I, Figure 4b), site (F =

89.81, df = 1, p < 0.0001) and interaction effects (F = 33.08, df = 3, p < 0.0001).

Habitat-corrected macroinvertebrate density was significantly greater upstream than downstream in both Iao Stream and Waihe’e River (t = 10.55, p < 0.001; t = 9.462, p <

0.001; respectively). However, this was not the case for Waikapu Stream and Waiehu

Stream (t = 0.3134, p > 0.05; t = 0.7006, p > 0.05; respectively). When comparing habitat-corrected densities between streams, Waikapu Stream had significantly greater density than Waiehu Stream, both upstream and downstream. Conversely, Waiehu

Stream downstream density was similar to downstream sites of the larger streams.

Further, habitat-corrected density did not significantly differ between the downstream sites in Waikapu Stream and Waihe’e River (t = 1.580, p > 0.05), whereas upstream densities significantly differed among all streams following the trend of Waiehu Stream <

Waikapu Stream < Iao Stream < Waihe’e River (Fig. 4b). Finally, Simpson’s

Biodiversity Index was not different among streams or between sites (Tables I and II).

!81! ! ! !

To understand the differential effects of water removal on the macroinvertebrate communities occupying different habitats, we pooled data from all streams and compared macroinvertebrate density in riffles versus cascades (pooled torrenticolous and amphibious microhabitats, as given in Shoda et al., 2010). There was significantly greater macroinvertebrate density in upstream habitats (F = 25.77, df = 1, p < 0.0001), but this depended on habitat (F = 138.9, df = 1, p < 0.0001) (interaction: F = 17.99, df = 1, p <

0.0001). The cascade habitat was then separated into amphibious and torrenticolous microhabitats, and similarly, there were significant site (F = 15.02, df = 1, p = 0.0001), microhabitat (F = 76.69, df = 2, p < 0.0001) and interaction (F = 9.308, df = 2, p =

0.0001) effects on macroinvertebrate densities (Fig. 5). However, macroinvertebrate density was not significantly different between upstream and downstream sites within amphibious (t = 0.3112, p > 0.05) or torrenticolous (t = 0.6425, p > 0.05) microhabitats; further, habitat-corrected density calculations were not possible for these microhabitats.

Community composition

Riffle macroinvertebrate communities were primarily composed of three introduced taxa, including Chironomidae (excluding Telmatogeton sp.) and two

Trichoptera, Cheumatopsyche sp., and Hydroptila sp. (Fig. 6, Table II). When all streams and sites were pooled, introduced Chironomidae dominated the riffle communities at

58.6%, followed by Cheumatopsyche, 27.9%, and Hydroptila, 8.6%. Similarly, introduced Chironomidae dominated the torrenticolous (64.9%) and amphibious (46.5%) communities. Cheumatopsyche, 21.1%, and Hydroptila, 7.0%, also composed the torrenticolous community, however, these introduced Trichoptera contributed much less to the amphibious community, 2.2% and 2.3%, respectively. Instead, amphibious

!82! ! ! ! communities included native taxa, Limonia (30.4%) and Ephyridae (10.7%). Finally,

Telmatogeton sp. made up 3.0% and 1.8% of the torrenticolous and amphibious communities, respectively. In all habitats, other taxa, including native, endemic and introduced taxa, were present but no other taxa represented more than 3% of the total community. When streams and sites were considered independently, the only additional taxa that made up greater than 3% were introduced Physa sp. (Waikapu US: 4.7%) and

Oligochaeta (Waiehu DS: 3.5%; Waihe’e US: 4.0%).

To evaluate the general impact of water removal on non-native versus native taxa, streams were pooled to compare native and introduced taxa density differences related to stream diversions (Fig. 7). There was only a significant main effect of site on native taxa density (F = 18.40, df = 1, p < 0.0001), whereas there were significant effects of habitat (riffle vs. cascade; F = 39.42, df = 1, p < 0.0001), site (F = 13.21, df = 1, p =

0.0004), and their interaction (F = 9.911, df = 1, p = 0.0019) on introduced taxa density.

In the riffle habitat, the density of both native and introduced taxa was significantly greater upstream than downstream (t = 3.998, p < 0.001; t = 4.796, p < 0.001; respectively). Further, introduced taxa density was significantly greater upstream in the riffle than the cascade habitat (t = 6.665, p < 0.001; Fig. 7).

The cascade habitat was further separated into microhabitats where we found a significant main effect of microhabitat on the Index of Nativity in each stream; however, a significant main effect of site was only evident in the larger streams – Iao Stream (F =

7.385, df = 1, p = 0.0079) and Waihe’e River (F = 4.503, df = 1, p = 0.0365). Further, the

I.N. was significantly greater in the upstream compared to the downstream amphibious microhabitat in Iao Stream (t = 4.309, p < 0.001) and Waihe’e River (t = 3.309, p < 0.01).

!83! ! ! !

The index was not a different between sites in the two smaller streams – Waikapu and

Waiehu Streams (F = 1.785, df = 1, p = 0.1847; F = 0.3997, df = 1, p = 0.5293; respectively). Additionally, when streams were pooled, the amphibious I.N. was significantly greater than both riffle and torrenticolous habitats, between downstream sites (t = 4.417, p < 0.001; t = 4.205, p < 0.001; respectively) and between upstream sites

(t = 4.821, p < 0.001; t = 4.728, p < 0.001; respectively) (Fig. 8). Conversely, I.N. did not differ between riffle and torrenticolous habitats in any stream or between downstream (t

= 0.207, p > 0.05) or upstream sites (t = 0.092, p > 0.05).

Using NMDS with MRPP we found a significant difference in macroinvertebrate community structure among habitats (Euclidean distance measure, A = 0.185, p =

0.00000; Figure 9). There also was a significantly different community composition between upstream and downstream sites (A = 0.0086, p = 0.0029) and among streams (A

= 0.0092, p = 0.011). Specifically, this was evident between Waikapu and Iao Streams (p

= 0.01) and between Waikapu Stream and Waihe’e River (p = 0.003). We found that

Chironomidae, Cheumatopsyche and Hydroptila accounted for the highest percentage of perfect indication of riffle habitat with 75%, 68%, and 67%, respectively (p = 0.0002, for each). Endemic Telmatogeton sp. and Procanace were 27% and 10% indicators of the torrenticolous microhabitat, respectively (p = 0.0002 & 0.0086), whereas, native

Tipulidae (i.e. Limonia) and Ephydridae were significant indicators of the amphibious microhabitat (61% and 29%, respectively; p = 0.0002 for both).

!84! ! ! !

DISCUSSION

Physical habitat template

The consequences of reduced streamflow on physical habitat are consistent and well documented (see Fig. 1 in Dewson et al., 2007 for full review). Comparisons between upstream and downstream wetted width, percent canopy cover, substrate size and gradient for each of these West Maui watersheds were presented in Shoda et al.

(2010). Trends of lower discharge, FPOM and chlorophyll a due to stream diversions were evident across the four watersheds. The two larger watersheds, Iao Stream and

Waihe’e River, experienced significant water reduction downstream of the diversions; although discharge downstream of diversions in the smaller streams was substantially lower, it was not statistically different from upstream conditions. While the large and small streams differed in stream size and upstream discharge volumes, strikingly, downstream discharge and flow velocity were similar across the four streams, indicating the diversions varied in capacity and efficiency, yet created relatively similar downstream flow environments. Therefore, the extent to which the diversions affected physical habitat and biological communities was dependent upon stream size and diversion capacity.

Such conclusions could impact restoration and conservation efforts of the N! Wai ‘Eh!, especially during recent legal proceedings (Case no. CCH-MA06-01) concerning freshwater resources and the return of flow back into diverted streams (Eagar, 2008;

Hamilton, 2008).

Similarly, in a study comparing two streams on the island of Molokai, Brasher

(1997) found that as discharge decreased below a diversion, flow velocities, water depth and wetted channel width were reduced. We found similar effects of significantly

!85! ! ! ! reduced wetted width and depth downstream of the diversions in Iao Stream and Waihe’e

River; and depth was significantly lower downstream in the smaller streams. The similar levels of fine particulate organic matter among upstream sites of the four watersheds suggest comparable allochthonous input and breakdown; however, the transfer of FPOM to downstream sites varied considerably among streams. Further, chlorophyll a was significantly lower downstream of diversions, which was expected as primary production would decrease with increased canopy cover (Hauer & Lamberti, 2006) from encroaching riparian plants in the dry stream sides, which is what we found in this study. Kinzie et al.

(2006) found a similar trend of reduced benthic chlorophyll a at dewatered sites on the island of Kauai. Although our results indicate variability in chlorophyll a among streams, unfortunately our study only included one isolated sampling event for chlorophyll a and

FPOM from each study site, so these results should be interpreted with caution. A more thorough systematic sampling regime would be necessary for broader conclusions (Hauer

& Lamberti, 2006).

Reduced discharge affected habitat availability for macroinvertebrate colonization, where the relative percent of riffle, cascade, and pool habitats were lower downstream in the larger watersheds, while the relative percent of run habitat doubled. In the smaller streams, riffle and pool habitats actually increased downstream, whereas run and cascade habitats decreased. This is important when sampling macroinvertebrate communities, especially in cascade habitats, where it may be necessary to expand the size of a study reach during conditions of low flow (Benbow et al., 2003, 2005). In addition, overall habitat quality may be affected by reduced discharge (Miller et al., 2007) and impact the macroinvertebrate community composition and biomass as a result of altered

!86! ! ! ! habitat suitability for certain species (Gore et al., 2001; McIntosh et al., 2003, 2008;

Dewson et al., 2007; Benbow, 2008).

Macroinvertebrate density and diversity

Our results revealed that total riffle macroinvertebrate density in these West Maui watersheds responded negatively to reduced flows downstream of diversions, with a statistically significant difference found in Waihe’e River. Such a high density in the

Waihe’e River upstream site could be attributed to overall habitat quality or timing of sampling events. These results in Waihe’e River differ from a similar study by McIntosh et al., (2003) where macroinvertebrate densities were found to be similar between upstream and downstream reaches. Changes in hydrology and frequency of spate events may account for such variation in communities downstream of diversions that might mask the effect of diversions, perhaps because of low statistical power (McIntosh et al.,

2003). However, when a similar study was conducted in Iao Stream during the same summer (McIntosh et al., 2002), macroinvertebrate densities were significantly lower downstream of the diversion. It is important to note that habitat availability was not accounted for in either of these studies (McIntosh et al., 2002, 2003). Further, Kinzie et al. (2006) found that invertebrate abundance, diversity and biomass were reduced downstream of a Wainiha River diversion on Kaua’i, Hawaii.

When habitat availability was taken into account there were profound and significant reductions in macroinvertebrate density downstream of the diversions in the larger streams. Interestingly, total habitat-corrected macroinvertebrate density was greater downstream than upstream in the two smaller streams, attributable to greater riffle habitat availability (and thus associated reduced cascades). Further, downstream habitat-

!87! ! ! ! corrected macroinvertebrate densities were generally similar among streams, likely due to diversion differences in capacity and efficiency, creating comparable habitats. These data demonstrate that accounting for available habitat is important to evaluating ecological change as a response to reduced streamflow (Shoda et al., 2010).

Community composition

Similar to other Hawaiian habitats, non-native organisms have invaded the freshwater streams. Our findings suggest that more than 95% of riffle community composition in these West Maui watersheds was dominated by introduced taxa, whereas native taxa only comprised approximately 1% of the overall community structure, similar to the torrenticolous microhabitats reported by Shoda et al. (2010). Exotic Trichoptera and Chironomidae are common in continental streams and have successfully established themselves in the Hawaiian Islands (Howarth & Polhemus, 1991). Our findings are comparable to other recent Hawaiian studies, where alien species have been found to dominate riffle macroinvertebrate communities (McIntosh et al., 2002, 2003, 2008;

Brasher et al., 2004; Wolff, 2005; Kinzie et al., 2006).

When all data were pooled, the Index of Nativity identified the amphibious habitat as having the greatest proportion of native taxa (43.9%) compared to torrenticolous (4.7%) and riffle (1.0%) habitats. Interestingly, the Index of Nativity was not statistically different between the torrenticolous and riffle habitats; both compositions were dominated by introduced taxa. In all study sites upstream of diversions, the Index of

Nativity was greater in the amphibious habitat than in either of the other two habitats, suggesting this microhabitat under natural flow conditions is crucial for the persistence of native aquatic species in these tropical mountain streams. To our knowledge,

!88! ! ! ! macroinvertebrate community comparisons among Hawaiian stream microhabitats have not been addressed other than by Shoda et al. (2010). However, the value of native insects is widely acknowledged (Kido & Smith, 1997; Brasher et al., 2004; see Wolff,

2005) and they are considered potential indicators of Hawaiian stream habitat quality

(Kido & Smith, 1997; Shoda et al., 2010). Kido and Smith (1997) found that endemic genera such as Telmatogeton and Procanace were sensitive to reduced flow and could potentially act as indicators of moderate to excellent stream habitat quality. Our data from four Maui watersheds support this recommendation.

The native taxa Limonia, Ephydridae, Telmatogeton sp., and Procanace accounted for the highest percentage of perfect indication of cascade habitats. Non-native taxa dominated the riffle habitats and included the introduced Chironomidae,

Cheumatopsyche and Hydroptila as the indicator taxa. Similarly, Brasher et al. (2004) found that Cheumatopsyche pettiti dominated riffle communities in 9 streams on Oahu and Shoda et al. (2010) reported that native species represented amphibious communities.

Significant community composition separation was found between upstream and downstream sites suggesting that water withdrawal influenced overall macroinvertebrate community structure. Separation was further found between streams of different sizes, implying community structure similarity between streams of similar size and diversion efficiency.

Through years of continental watershed research, riffle habitats are widely accepted and recommended for biomonitoring (Hauer & Lamberti, 2006; Merritt et al.,

2008). However, our results here support a growing literature base of tropical island streams suggesting that riffles are dominated by introduced taxa (Kido & Smith, 1997;

!89! ! ! !

Kinzie et al., 1997; Brasher et al., 2004; Wolff, 2005); thus, biomonitoring programs in these tropical regions target non-native organisms. Our investigation of cascade habitats clearly indicates their importance in high gradient, flashy, tropical streams, for understanding macroinvertebrate community responses to anthropogenic disturbance.

Where water conservation and management are of concern in regions of high endemicity, we propose that biomonitoring programs include habitats where native and endemic organisms are most abundant and sensitive to appropriate anthropogenic stressors.

Identifying habitats, such as cascades, where native organisms can thrive, despite an overall community structure overrun by introduced taxa, may be necessary in establishing effective biomonitoring programs.

In the last decade, it has become increasingly evident that conserving aquatic ecosystems while balancing freshwater needs is a pressing global issue, with researchers in tropical regions urging a better understanding of the direct and indirect impacts of flow modifications on aquatic communities and overall ecosystem services for the development of effective management strategies (Power et al., 1996; Jackson et al., 2001;

Baron et al., 2002; Brasher, 2003). However, the task of determining minimum flows is a difficult and politically sensitive task for water resource managers (Bunn & Arthington,

2002; Dewson et al., 2007). If ecosystem structure and function is to be restored and maintained in stream networks, water-conservation strategies and a strong commitment to sustainable water is necessary, especially on tropical islands with high endemism and where freshwater resources are already limited (Smith et al., 2003).

!90! ! ! !

ACKNOWLEDGMENTS

The authors would like to acknowledge C. Hanley, A. Jennings and D. Vonderhaar for their invaluable assistance during field sampling. K. Sproat, I. Moriwake, D. Oki, J.

Duey, J. Verel, and the Pellegrino and Duberstein families provided logistical support, stream access and hospitality while conducting this research. This project was made possible through the financial support of the US Geological Survey, Earthjustice, the

Office of Hawaiian Affairs and the University of Dayton Graduate School and

Department of Biology.

!91! ! Table I: Two-way ANOVA statistics for physical habitat template characteristics and macroinvertebrate density and diversity for the riffle habitat. Site refers to upstream or downstream of water diversions among the study streams. Bonferroni corrected p-value = 0.005 was used to interpret statistical significance.

Parameter Source of Percent of F df P value variation variation value

Discharge (m3/s) Stream 24 49.98 3 <0.0001 Site 30.2 188.6 1 <0.0001 Stream x Site 22.2 46.16 3 <0.0001

Depth (m) Stream 7.6 63.56 3 <0.0001 Site 10.1 254.5 1 <0.0001 Stream x Site 1.3 10.96 3 <0.0001

Flow Velocity (m/s) Stream 0.7 5.28 3 0.0013 Site 9.9 211 1 <0.0001 Stream x Site 0.8 5.63 3 0.0008

FPOM (mg/L) Stream 30.3 7.3 3 0.0007 Site 3.1 2.21 1 0.1473 Stream x Site 22.5 5.43 3 0.0039

Chlorophyll a (µg/L) Stream 34.6 8.58 3 0.0003 Site 19.3 14.32 1 0.0006 Stream x Site 3.2 0.78 3 0.5124

Autotrophic Index Stream 16.9 2.8 3 0.0579 Site 19.6 9.7 1 0.0041 Stream x Site 5.1 0.85 3 0.4797 ! ! ! ! ! ! ! ! ! ! ! ! !

! 92! Table I: continued.

Total macroinvertebrate Stream 3.6 1.95 3 0.1256 -2 Density (# m ) Site 7.1 11.49 1 0.0009 Stream x Site 8.5 4.59 3 0.0043

Total Riffle-Corrected Density Stream 18 24.43 3 < 0.0001 (# m-2 per 100m) Site 22 89.81 1 < 0.0001 Stream x Site 24.3 33.08 3 < 0.0001

Diversity Stream 0.8 0.38 3 0.7704 Site 1.2 1.58 1 0.2104 Stream x Site 3.8 1.75 3 0.1601

! 93! Table II: Mean (± SD) density of represented macroinvertebrate taxa in the riffle habitats among the study streams and between upstream and downstream of diversion. !

Riffle taxa Waikapu Stream Waiehu Stream Iao Stream Waihe'e River US DS US DS US DS US DS

Megalagrion sp. 2 ± 2 1 ± 1 8 ± 10 4 ± 4 0 ± 0 0 ± 0 0 ± 0 0 ± 0

Dytiscidae 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0

Chironomidae sp. 185 ± 137 155 ± 134 272 ± 141 157 ± 147 215 ± 270 253 ± 316 430 ± 336 168 ± 214

Telmatogeton sp. 2 ± 1 1 ± 1 3 ± 1 1 ± 0 2 ± 1 1 ± 1 4 ± 1 2 ± 2

Procanace sp. 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 1 0 ± 0

Ephydridae 0 ± 0 0 ± 0 0 ± 1 0 ± 0 1 ± 1 0 ± 1 1 ± 2 0 ± 1

Limonia sp. 0 ± 0 0 ± 0 1 ± 3 0 ± 0 1 ± 2 0 ± 1 1 ± 1 0 ± 0

Cheumatopsyche analis 79 ± 78 47 ± 37 126 ± 129 48 ± 73 125 ± 183 53 ± 70 194 ± 152 53 ± 64

Hydroptila sp. 10 ± 15 1 ± 2 36 ± 50 18 ± 54 79 ± 101 44 ± 72 306 ± 471 28 ± 50

Hyposmocoma sp. 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0

Ferrissia sharpi 0 ± 0 0 ± 0 1 ± 1 2 ± 4 0 ± 0 0 ± 0 0 ± 0 0 ± 0

! 94! Table II: continued.

Lymnaeidae 2 ± 6 0 ± 1 0 ± 0 0 ± 1 0 ± 0 0 ± 1 0 ± 0 0 ± 0

Physa sp. 11 ± 12 4 ± 7 1 ± 4 0 ± 0 1 ± 1 0 ± 0 2 ± 8 0 ± 0

Oligochaeta 2 ± 3 1 ± 2 9 ± 14 23 ± 75 6 ± 8 2 ± 3 56 ± 100 6 ± 14

Oribatei 0 ± 0 0 ± 0 2 ± 3 1 ± 3 0 ± 0 0 ± 0 1 ± 2 0 ± 1

Atyoida bisulcata 0 ± 0 0 ± 0 1 ± 3 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 1

Prostoma sp. 0 ± 1 0 ± 0 0 ± 1 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0

Turbellaria 0 ± 0 0 ± 0 0 ± 0 0 ± 0 1 ± 2 2 ± 3 0 ± 0 0 ± 0

Erpobdellidae 1 ± 3 0 ± 1 1 ± 2 0 ± 0 0 ± 0 0 ± 0 0 ± 0 0 ± 0

Glossiphoniidae 0 ± 0 0 ± 1 0 ± 0 0 ± 0 0 ± 1 0 ± 0 0 ± 0 0 ± 0

! 95! Table III: Two-way ANOVA statistics for the Index of Nativity for each microhabitat and site within each stream. Site refers to upstream (US) or downstream (DS) of water diversion among the study streams. Bolded p-values indicate statistical significance (! = 0.05) and are followed by Bonferroni pairwise comparisons between US and DS reaches for each microhabitat are presented (ns = not significantly different). !

Percent Bonferroni post-test Stream Source of variation of F-value df p-value comparisons between US & variation DS sites for each habitat

Waikapu Site 1.4 1.79 1 0.1847 Stream Microhabitat 19.8 12.58 2 < 0.0001 ns Site x Microhabitat 2.8 1.76 2 0.1774

Iao Stream Site 5.5 7.39 1 0.0079 Amphibious microhabitat: Microhabitat 12.5 8.39 2 0.0005 t = 4.309, p < 0.001 Site x Microhabitat 10.2 6.85 2 0.0017

Waiehu Stream Site 0.4 0.4 1 0.5293 Microhabitat 19.1 9.58 2 0.0002 ns Site x Microhabitat 0.6 0.28 2 0.7558

Waihe'e River Site 3.7 4.5 1 0.0365 Amphibious microhabitat: Microhabitat 17.2 10.39 2 < 0.0001 t = 3.309, p < 0.01 Site x Microhabitat 6.2 3.77 2 0.0266

! ! !

! 96!

Figure 1: Map of West Maui, Hawaii, with study watersheds highlighted and corresponding study locations as black dots – A) Waikapu Stream, B) Iao Stream, C) Waiehu Stream and D) Waihee River. At each location, an upstream (US) and downstream (DS) site was sampled above and below the highest elevation diversion in each stream.

! 97! !

1-00

2 1-,0 2

B:9 ,-00 A 6@ ,-,0 ,-,/ ,-,. ,-,, 3'"45678953'":&;'<5=#:>(";?'5 !"#$"%& !"#'(& )"* !"#('+'

! Figure 2: Mean (SE) measured discharge in all streams, upstream (black bar) and downstream (white bar) of the highest elevation diversion. The * represents a significant difference between upstream and downstream sites at p < 0.001 (Bonferroni post-tests). ! ! ! ! ! ! ! ! ! ! ! ! ! ! ! ! ! ! ! !

! 98! ! !

3440 &'()'*+, -.//0+ -12 %!!

=+'2,3+9)+2;,>'?.;'; "#

! 56 76 56 76 8'9:+,6;9+'<( 6<'00,6;9+'<( ! Figure 3: Mean relative percent of available habitat within the upstream and downstream 100 m study reaches, where the large (Iao Stream and Waihee River) and small streams (Waikapu and Waiehu Streams) were pooled. The habitats were Run (Black), Riffle (Diagonal Stripe), Cascade (Grey), and Pool (White).

! ! ! ! ! ! !

! 99! ";

4 3.,/0,,2 ; 3 0.-/0,,2

0.,/0,,2

8*=D"6#BEB7F7E -.,/0,

, !"#$"%& !"#'(& )"* !"#('+' 5'"6789:;75"<=*#6>'=?'@="?'7A'6B#?C

@; 5'"6789:;7K*=='

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0.,/0,,-

0.,/0,,- 0.,/0,,2 !"#$$%!&'(#" 87*=D"6#BEB7F7='"<(;

,1 5'"6789:;7K*=='

! Figure 4: a) Mean (SE) riffle macroinvertebrate non-corrected density and b) mean (SE) habitat-corrected density for upstream (Black) and downstream (White) reaches within each stream. The * represents a significant difference between upstream and downstream pairs at p < 0.001 (Bonferroni post-tests).

! 100! %&!!! 9 $!!! @ & #!!!

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:+;6<=>?@<:;851(6A+57+05;7+

Figure 5: Mean (SE) macroinvertebrate density between riffle habitat and amphibious and torrenticolous microhabitats of cascades upstream (Black) and downstream (White) of diversions, with all streams pooled. The * represents a significant difference between upstream and downstream pairs at p < 0.001 (Bonferroni post-tests).

! 101!

544

84 !<(%.B*;(";<%? 54 98'5%)%5".-1H? !"#$%&'("&')*+,-. 4 :2"(')'0";<% !"##$% &'((%)*"+'$',-. /012"3"',-

Figure 6: Community composition of riffle, torrenticolous and amphibious habitats for all streams and sites pooled. Taxa comprising less than 3% of the total community were pooled and defined as rare. Native taxa are denoted with an * and represented as solid bars whereas introduced species are represented by hashed lines.

! 102!

A & %&!!! %&! '()*+,-.(/(->:9D(8*4E4FE =,(8->?@A-B,84*)C-:1- < < $!!! $!

#!!! #!

"!!! "! =,(8->?@A-B,84*)C-:1- & ! ! A 78)9:6;5,6-.(/(->:9D(8*4E4FE 0*112, 3(45(6, 0*112, 3(45(6, 78)9:6;5,6-.(/( '()*+,-.(/( ! Figure 7: Mean (SE) macroinvertebrate density for habitat and location upstream (Black) and downstream (White) of the highest elevation diversions of all streams (data pooled). Introduced taxa density references the left y-axis and native density references the right y-axis. The * represents a significant difference at p < 0.001(Bonferroni post-tests). There was also a significant difference in introduced taxa density between riffle and cascade habitats, upstream of diversions (t = 6.665, df = 1, p < 0.001). ! ! ! ! ! ! ! ! ! ! ! ! ! ! ! ! ! !

! 103! 1./ 2 2

0./

-./ ) 3 -.0- ) 3 4$5(67"879)&:;:&< -.-/ ) ) -.-- !"#$%&'()* +,%&'()*

Figure 8: Index of Nativity (ratio of native taxa density to introduced taxa density) for each habitat (Riffle – Black, Torrenticolous – Grey, Amphibious – White) and location upstream and downstream of the highest elevation diversion in each stream (data pooled). Different letters indicate significant differences between columns (p < 0.001, Bonferroni post-tests).

! 104!

Figure 9: NMDS ordination with habitat overlay. Total stress = 15.71. Axis 1 explained 52.1% and axis 2 explained 18.7% of the variation in macroinvertebrate community structure. MRPP analysis demonstrated significant separation in community composition among habitats (A = 0.185, p < 0.0001).

! ! !

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! 111!

CHAPTER IV

BENTHIC COMMUNITY STRUCTURE UNDER DIFFERENT FLOW AND

SUBSTRATE CONDITIONS IN THE LITTLE MIAMI RIVER, OHIO

ABSTRACT

Streams and rivers are made up of a continuum of distinct habitats composed of macroinvertebrate and fish communities. Many studies have investigated the differences between these geomorphic units, but often focus on riffle and pool habitats. This study compares the physical template and the macroinvertebrate community, in June and

September 2008, between riffle and run habitats of a 1 km reach in the Little Miami

River, Greene County, Ohio. Mean flow velocity was significantly greater in riffle (mean

± SE = 0.74 ± 0.04 m/s) than run (0.32 ± 0.01 m/s) habitat units (p < 0.0001, Mann-

Whitney U = 28.0). Similarly, macroinvertebrate density was significantly greater in riffle habitats (mean ± SE; 1892 ± 200.2, 4018 ± 493.6) than in run habitats (540.3 ±

76.8, 930.1 ± 139.6), in both June and September, respectively. Further, linear regression found a positive and significant relationship (y = 4097x – 115.1, p < 0.0001) where 49% of variation in macroinvertebrate density was explained by mid-column velocity. Our results call for the need of future analyses and studies that focus on the relationship between flow and community dynamics, using simple and complex hydraulic variables in

! 112! an attempt to determine which more accurately predict the distribution of invertebrate communities.

INTRODUCTION

Streams are generally acknowledged as having a patchy landscape (Hildrew &

Giller 1994). Geomorphic units such as pools, riffles and runs are distinct habitats for specific assemblages of aquatic flora and fauna such as macroinvertebrates (e.g. Lium,

1974; Logan & Brooker, 1983; Pridmore & Roper, 1985; Brown & Brussock, 1991), macrophytes, and algae (e.g. Korte & Blinn, 1983), and represent distinct habitats for fish

(e.g. Saffel & Scarnecchia, 1995; Braaten & Berry, 1997). Ecologists use these various habitats as sampling units or for describing habitat use (Allen 1951, Pridmore & Roper

1985, Glova & Duncan 1985, Jowett 1993). However, these sampling units are often used to reduce or to demonstrate the variability caused by the physical differences in these habitat types, and thus further efforts are usually limited to one habitat type (i.e. riffle habitats) (Allen 1951, Lium 1974, Logan & Brooker 1983, Pridmore & Roper 1985,

Brown & Brussock 1991, Jowett 1993).

While many studies have compared the faunas of different stream habitats, most comparisons are concerned mainly with riffle and pool habitats (Dolling 1968, Lium

1974, Armitage 1976, Minshall & Minshall 1977, Scullion et al. 1982, Logan & Brooker

1983, O’Neill & Abrahams 1984, Brown & Brussock 1991, Jowett 1993). These habitats are known to differ significantly in flow velocity, depth and substratum (O’Neill &

Abrahams 1984) and therefore, differences in faunal abundance and diversity may be expected. However, variation among riffle and run habitats may not be as apparent, and thus, are not typically the focus of biological studies.

! 113! A riffle habitat is a geomorphic unit that can be defined as a “place where there is an obstruction in the stream, producing a ripple, or a stretch of shallow, rapid or choppy water, typically providing a well-oxygenated habitat” (Brown & Shoemake 1964). A riffle habitat is hydraulically rough and a complex patchwork of microhabitats (Pringle et al. 1988). Within a riffle habitat, sufficient microhabitat heterogeneity may exist within a quadrat to enable several species to coexist and minimize intra- and interspecific competition (Sites & Willig 1991). On the other hand, run habitats are hydraulically smooth, of increased depth and made up of fine sediments, homogeneous in nature. Run habitats often represent the “average” current velocity and depth in a lotic system

(Mosely 1982) and therefore are more similar in physical characteristics to riffles than pools (Pridmore & Roper 1985).

Accurate identification of distinct habitats in the aquatic environment has been encouraged as an important tool in monitoring programs (i.e National River Authority

1993 in Buffagni et al. 2000, Moore et al. 1997). However, pools, runs and riffles are not discrete units and instead form a continuum for which classification is arbitrary and consequently there will be some overlap in any subjective method of assessment (Jowett

1993). The functional habitat concept, proposed by Harper et al. 1992 and 1995, is a habitat-based approach that identifies distinct habitat units based on physical and biological attributes, using the distribution of macroinvertebrate assemblages (Buffagni et al. 2000). Effectiveness of the functional habitat approach strongly depends upon the existence of separate physical habitats that support different assemblages with distinct functional roles in the river environment (Buffagni et al. 2000).

! 114! Characterizing the hydrology, substrate and biological communities of river systems can be done at various scales that provide different ecological information regarding the structure and function of the stream ecosystem. In this study, our objective was to compare the physical habitat template and biological communities between riffle and run habitats in a 1 km reach of the Little Miami River. We hypothesized that macroinvertebrate density and diversity would be greater in riffle habitats with a positive relationship between flow velocity and density.

METHODS

Study site

The study took place in a 1 km reach of the Little Miami River, Ohio in Greene

County, directly upstream of the Jacoby Rd. canoe launch park, which was fully accessible throughout the study period. Here, the river is dominated by a natural sequence of riffle and run habitats, with very few pools. The Little Miami River flows

107 miles from its headwaters in southeastern Clark County to its confluence with the

Ohio River in Hamilton County east of Cincinnati, with a total drainage area of 1,757 square miles. The topography of the Little Miami River Watershed has been influenced by glacial activity, which left distinctive landforms and thick deposits of silt, sand, and gravel. Aquatic life use designations for the streams in the watershed reflect the generally good conditions in the watershed. The stream is designated as an Exceptional

Warmwater Habitat (EWH) and has a State and National Scenic River status. Land uses in the watershed are principally agricultural in the northern and eastern portions with relatively limited development near cities. Beginning with the Dayton-Xenia corridor there is an increasing impact from population and development. While most developing

! 115! areas in the Little Miami River watershed are not immediately adjacent to the river, the impacts of development are still a potential problem. Numerous residential, industrial, and commercial developments are recently completed, underway, or proposed within the watershed (OEPA 2000).

A USGS gauge station (03240000) on the Little Miami River near Oldtown, Ohio

(39°44’54” and 83°55’53”) was used to obtain discharge data throughout the 2008 study period and data was compared to field measurements made according to Gordon et al.

(1992).

Field collections

At the Jacoby Rd. canoe launch site within the Little Miami River, a 1 km reach was measured and five sequential riffle and five sequential run sites that ranged from

39°45.837, 83°54.137 to 39°45.998, 83°53.937 were identified based on substrate and flow characteristics. These were identified as Riffles 1 – 5 and Runs 1 – 5. A 100 square meter grid was placed within each riffle and run site and a random number generator provided six 1 x 1 m cell locations from which benthic samples were collected. Prior to each sampling, water depth and velocities were measured above the benthic habitat at four equidistant points within the modified Surber sampler area (0.0625 m2). Flow velocity measurements were taken at 0.2, 0.6 and 0.8x total depth to generate a flow profile within the water column. The benthic area was then scrubbed for 30s with a coarse brush and dislodged material was collected in a D-frame net placed directly downstream of the sample area. All samples are stored in 70% ethanol for laboratory sorting and identification. Sampling events took place once a month from June –

September in 2008, collecting six samples from each site each month. Samples were

! 116! sorted and organisms were identified to genus or lowest possible taxonomic resolution for the June and September sampling events.

Substrate size and canopy cover was described for each site once during the study period. The composition of substrate has been identified as being an important parameter affecting the spatial arrangement of benthic communities. Using a gravelometer and the

Wolman Pebble Count procedure, substrate size was measured by walking heel-toe across the wetted width of the channel, representing the most downstream transect of each habitat unit, and measuring the rock touching the toe after each step. If there are less than ten measurements along the transect, the recorder moved 1m upstream and continued the process. Therefore, at least 100 measurements were made to generate a substrate size frequency distribution for each habitat unit. These measurements were then plotted and a frequency distribution produced. Describing the amount of light infiltrating to the streambed is important for in-stream primary production, thus a potential food resource for the benthic organisms. This was accomplished by determining the percent canopy cover using a convex densiometer. Four estimates (one facing downstream, toward the left descending bank, upstream and toward the right descending bank) were taken in the middle of the wetted width at the most downstream transect of each habitat. From the four estimates, canopy cover was calculated for each site.

Statistical analyses

Total macroinvertebrate density described macroinvertebrate abundance, whereas

Simpson’s Diversity Index expressed extent of biodiversity at each site and was used to compare between riffle and run habitat units. To determine macroinvertebrate density

! 117! differences among riffle sites and among run sites, one-way ANOVA with Bonferroni post-tests were performed using GraphPad Prism 5.0 (GraphPad Software). After pooling all riffle sites and all run sites, two-way ANOVA with Bonferroni post-tests determined differences in these macroinvertebrate metrics between habitat units (riffle vs. run) and sampling events (June vs. September). All data were appropriately transformed to meet assumptions of normality and homogeneity of variances; however, when transformations did not produce normal distributions, nonparametric analyses were used.

A two-tailed Mann-Whitney U t-test was used to determine flow velocity differences between riffle and run habitats. Linear regression identified the relationship between macroinvertebrate density and flow velocity. Nonmetric multidimensional scaling

(NMDS) followed by multi-response permutation procedures (MRPP) using PC-Ord

(McCune & Mefford 1999), investigated overall macroinvertebrate community composition between habitats, and between sampling events. Prior to the NMDS ordinations, data were log (x + 1) transformed and Sorensen distance measure was used with a random starting configuration and 250 runs with real data (McCune & Grace

2002). As described by in McCune & Grace (2002), results from Monte Carlo tests with

4999 randomizations determined significant differences in community structure.

RESULTS

USGS mean daily discharge data was used to generate a hydrograph during the study period (Figure 1). On the two sampling dates, June 20 and September 13, 2008, the

Little Miami River average daily discharge was 3.34 m3/s and 0.71 m3/s, respectively.

When sampling dates and sites were pooled, average flow velocity was significantly

! 118! greater in riffle (mean ± SE = 0.74 ± 0.04 m/s) than run (0.32 ± 0.01 m/s) habitat units

(Figure 2; p < 0.0001, Mann-Whitney U = 28.0).

Substrate size was measured at each site and pooled for each habitat. Following the Wolman Pebble Count and distribution procedure, it was determined that 50% of the samples were equal to or smaller than 25 mm in the run habitat and 60 mm in the riffle habitat (Figure 2). Canopy cover was also measured at each site, facing four different directions and averaged. Riffle habitats had 77.9% canopy cover, whereas run habitats were slightly less at 72.5%.

Total macroinvertebrate density varied among riffle sites and among run sites.

One-way ANOVA results indicated that in June, macroinvertebrate density did not significantly differ among riffle sites (p = 0.0642, Kruskal-Wallis KW = 8.88), whereas density was significantly different among run sites (p = 0.004, KW = 15.53), with Dunn’s multiple comparison test identifying differences between Run 2 and 4, and between Run

3 and 4. In September, conversely, macroinvertebrate density among riffle sites significantly varied (p = 0.01, KW = 12.43) with a difference between Riffle 1 and 5; however density did not significantly differ among run sites (p = 0.49, KW = 4.42).

While a general trend was not seen among sites and between sampling times, total macroinvertebrate density from all riffle sites and from all run sites were pooled for comparisons between habitat units. There were significant main effects of habitat (p <

0.0001, F = 63.75, df = 1) and sampling time (p < 0.0001, F = 20.46, df = 1) on macroinvertebrate density when biological data from riffle sites and run sites were pooled. Macroinvertebrate density was significantly greater in riffle habitats (mean ± SE;

1892 ± 200.2, 4018 ± 493.6) than in run habitats (540.3 ± 76.8, 930.1 ± 139.6), in both

! 119! June and September, respectively (Figure 3). Bonferroni post-tests found macroinvertebrate densities were significantly greater in the riffle habitat in September than June (p < 0.01 t = 5.41), whereas density in the run habitat did not differ between the two months (p > 0.01, t = 0.99). Similarly, there were significant main effects of habitat

(p = 0.02, F = 6.055, df = 1) and sampling time (p = 0.03, F = 4.7, df = 1) on Simpson’s

Diversity Index. Diversity was greater in the riffle than run habitat in June (mean ± SE;

0.88 ± 0.03, 0.79 ± 0.06) and by September, diversity increased in both habitats, respectively (1.03 ± 0.05, 0.87 ± 0.06). Further, linear regression found a positive and significant relationship (y = 4097x – 115.1, p < 0.0001) where 49% of variation in macroinvertebrate density was explained by mid-column velocity (Figure 4).

While the relative percent composition of taxa varied between habitat units, seven taxonomic groups comprised 90% of riffle and 89% of run habitats, including

Chironomidae, Heptageniidae, Optioservus sp., Hydropsyche sp., Baetis sp., Caenis sp., and Tricorythodes sp. (Figure 5). The remaining 10 – 11% was made up of rare taxa (<

3% each) including Ephemera sp., Isonychia sp., Plauditus sp., Dubiraphia sp.,

Psephenus sp., Berosus sp., Antocha sp., Simullidae, Cheumatopsyche sp., Hydroptila sp., Goniobasis sp., Corbicula sp., Sphaeriidae sp., Lebertia sp., Sperchon sp.,

Torrenticola sp., Ceratopogenidae, Corydalidae, and Nematoda. Although rare, there were two species, Isonychia sp. and Plauditus sp., which were only found in riffle samples, while Corydalidae were only found in run samples. All other taxa were present in both riffle and run habitats.

Community composition differences between riffle and run habitats were investigated through multivariate analyses, however, a 3-dimensional NMDS ordination

! 120! with high stress (32.41) explained only 0.1% of the variation among communities (Figure

6). Although clear partitioning was not visually obtained on the ordination, significant community composition separation was found between riffle and run habitat communities

(A = 0.069, p = 0.00000) and between June and September sampling times (A = 0.061, p

= 0.00000) through MRPP analysis.

DISCUSSION

As depicted in the hydrograph, mean daily discharge decreased throughout the summer season, which was expected as later summer months are drier and typically experience less precipitation in continental mid-western streams. While total discharge did decrease, the streambed remained wetted at all sites and allowed for sampling. Also expected were the differences in flow velocity and substrate size between riffle and run habitats. Riffle habitats are known to be more hydraulically rough with greater substrate heterogeneity (Pringle et al. 1988), whereas run habitats are hydraulically smooth with smaller, homogeneous sediments.

Higher total densities were found on the cobbled beds of riffles than in the predominantly sand/gravel beds of runs, which were comparable to results found by

Pridmore & Roper (1985). This would be expected because riffle habitats are naturally more oxygenated, readily provide food resources, and are comprised of diverse microhabitats for an array of macroinvertebrate assemblages. Many have observed that benthic invertebrates increase in numbers with increasing particle size from sand through large boulders (Percival & Whitehead 1929, Tarzwell 1936, Pennak & Van Gerpen 1947,

Sprules 1947, Bell 1969, Pridmore & Roper 1985). Hynes (1970) suggested that an increase in mean particle size may be related to an increase in the complexity of the

! 121! substratum, which in turn in essential for an abundant and diverse fauna. Further, the relationship between hydraulic variables and invertebrate densities and community structure has been the topic of many investigations, especially in the development of habitat suitability curves for benthic invertebrates (Quinn & Hickey 1994, Lloyd & Sites

2000, Jowett 2003, Mérigoux & Dolédec 2004, Brooks et al. 2005). Specifically, Froude number has been shown to be closely related to the distribution of some stream insects

(Statzner 1981, Orth & Maughan 1983, Jowett et al. 1991, Jowett 1993).

In this investigation, community composition was qualitatively similar between riffle and run habitats. Only two taxa groups, Isonychia sp. and Plauditus sp., were found exclusively in riffle habitats. Most Ephemeroptera nymphs are collectors or scrapers that feed on a variety of detritus and algae (Merritt, et al. 2008), therefore these findings would be expected. Our results were consistent with Pridmore & Roper (1985) who concluded that a qualitative description of the riffle fauna would likely describe the run fauna as well. However, quantitative differences did exist between riffle and run habitats.

Therefore, when quantitative information is required, as in productivity studies or resource assessment surveys, the two habitats should be sampled separately (Pridmore &

Roper 1985).

Hydrodynamic variables may be the most important, yet least understood environmental factors affecting the ecology of benthic organisms (Hart et al. 1996). This may be supported by our positive and significant relationship obtained between mid- column velocity and invertebrate density. The hydraulic stream ecology approach suggested by Statzner et al. (1988) links the metabolism, feeding and behavior of lotic organisms to the physical environment through the consideration of hydraulic

! 122! characteristics. Such an approach, utilizing simple and more complex hydraulic variables, would better our understanding of the distribution of macroinvertebrate communities.

In conclusion, our hypotheses were supported and basic knowledge of riffle and run habitats was further gained. Our results call for the need of future analyses and studies that focus on the relationship between flow and community dynamics, using simple and complex hydraulic variables in an attempt to determine which more accurately predict the distribution of invertebrate communities. Other field studies have obtained mixed results and it is still unclear which are better predictors of benthic invertebrate abundance and habitat suitability (Jowett et al. 1991, Quinn & Hickey 1994,

Gore 1996, Jowett 2003, Mérigoux & Dolédec 2004).

! 123!

100

60 ) s / 3

m 40 (

e g r 20 a h c s i 15 D

0 May June July August September October

Figure 1: Hydrograph depicting mean daily discharge (m3/s) for the Little Miami River (USGS gage 03240000), near Oldtown, Ohio from May through September 2008. Vertical dotted lines indicate sampling events. Data were retrieved from www.usgs.gov.

! 124!

100 t n e )

m 75 i % ( d

e n o S i

t e

u 50 v i b i t r a t l s u i D

m 25 u C 0 1 2 4 8 16 32 64 128 256 512 1024 2048

Particle Size Category (mm Log2)

Figure 2: Riffle and run habitat substrate particle size frequency distribution. Horizontal dashed line indicates the particle size at which 50% of the substrate was equal to or smaller than for each habitat. Square data points represent the run habitat and circle data points represent the riffle habitat.

! 125!

s / 1.0 m

) * E

S 0.8 (

y t i

c 0.6 o l e V

n 0.4 m u l

o 0.2 c - d i

M 0.0 Run Riffle

Figure 3: Mean (SE) flow velocity measured at each benthic sample in the riffle and run habitats. The * represents a significant difference at p < 0.0001 level.

! 126!

5000

y c t i s n e ) 4000 2 D

m / e t s

a 3000 r m b s

i b e t n r a

e 2000 g v r

n a o i (

o a r 1000 c a M 0 Run Riffle Run Riffle June September

Figure 4: Mean (SE) macroinvertebrate density for habitat and sampling time. Different letters indicate significant differences between comparisons (p < 0.05).

! 127!

8000 y t i s n e ) 2

D 6000

m / e t s a r m b s i

e 4000 t n r a e g v r n o i (

o 2000 r c a M 0 0.0 0.5 1.0 1.5 Mid-Column Flow Velocity (m/s)

Figure 5: Linear regression relationship between mid-column velocity (m/s) and macroinvertebrate density. Data obtained from riffle and run habitats were pooled. y = 4097x - 115.1 p < 0.0001 2 R = 0.49

! 128!

100

80 n o i t i s o

p 60 m o c

a x a t

n 40 Hydropsyche sp. e c r Optioservus sp. e

P Tricorythodes sp. Heptageniidae 20 Caenis sp. Baetis sp. Chironomidae

0 Rare (taxa<3%) Run Riffle

Figure 6: Community composition of riffle and run habitats for all sites pooled. Taxa comprising less than 3% of the total community were pooled and defined as rare.

! 129!

Figure 7: NMDS ordination with habitat overlay. Total stress = 32.41. Only 0.1% of the variation among samples was explained by the 3-dimensional solution. However, MRPP analysis demonstrated significant separation in community composition among habitats (A = 0.0691, p = 0.00000) and among sampling times (A = 0.2461, p = 0.00000).

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FUTURE DIRECTIONS

In summary, the projects involved in this dissertation investigated the effects of hydrology on habitat template and macroinvertebrate communities within lotic ecosystems. The completion of these projects has provided deeper insight into the ecological organization of streams and rivers, generated information that can be used to predict how flow alterations caused by human activities affect these vital ecosystems, and has the potential to guide management and restoration efforts.

While the conclusions drawn from each project will contribute substantially to fundamental knowledge of Stream Ecology, further analyses can provide a more comprehensive understanding of the relationship between hydrology and macoinvertebrate communities. Advanced hydraulic modeling, along with the calculation of simple and complex hydraulic variables could assist in predicting the spatial and temporal distribution of invertebrate communities. Data gathered from these investigations could be utilized for such purposes at a later time, and insight gained has generated additional questions and future project plans.

! 135!