STREAM FISH RESPONSE TO INTERMITTENCY AND DRYING IN …

STREAM FISH RESPONSE TO INTERMITTENCY AND DRYING IN THE

ICHAWAYNOCHAWAY CREEK BASIN

by

JESSICA L. DAVIS

(Under the Direction of Mary C. Freeman and Stephen W. Golladay)

ABSTRACT

Streamflow alteration from the combined effects of water extraction and climate

change is recognized as a major threat to aquatic ecosystems. The Ichawaynochaway

Creek Basin is a Gulf Coastal Plain stream system in southwestern Georgia, where

streamflows are strongly influenced by agricultural water withdrawals and recent

droughts. This study explores effects of stream intermittency and drying on the

composition of biologically diverse fish communities, and life history traits that may

influence persistence of four closely related cyprinid species. Intermittent stream

communities were found to be a subset of perennial stream communities, with the highest

persistence rates among adults and juveniles of species that commonly occur in

intermittent streams. My results identify life history traits that may be useful for

understanding differences in how closely related species respond to changing

environments, with smaller body size at maturity along with appropriate reproductive

timing promoting greater persistence given more frequent and intense disturbances.

INDEX WORDS: Warmwater Streams, Fish Community Structure, Drought,

Persistence, Colonization, Life History Traits



by

JESSICA DAVIS

B.S., University of North Carolina, Asheville, 2015

A Thesis Submitted to the Graduate Faculty of The University of Georgia in Partial

Fulfillment of the Requirements for the Degree

MASTERS OF SCIENCE

ATHENS, GEORGIA

2017

© 2017

Jessica L. Davis

All Rights Reserved



by

JESSICA L. DAVIS

Major Professor: Mary C. Freeman

Stephen W. Golladay

Committee:

Seth J. Wenger

Robert B. Bringolf

Electronic Version Approved:

Suzanne Barbour

Dean of the Graduate School

The University of Georgia

December 2017

iv

DEDICATION

For pop, the best dad a kiddo could ever have asked for.

v

ACKNOWLEDGEMENTS

I couldn't have made it through this project without the support from my

colleagues, family, and friends. My project would have been little compared to what it is

without the help of Mary Freeman at every turn. From helping write code, to always

making herself available for questions big and small, I couldn't have found a more caring

and supportive advisor. Special thanks to Steve Golladay, my co-advisor, for his support

of both me and my husband, d.w., during our time at the Jones Center. To my committee

members, Seth Wenger and Robert Bringolf, thank you for helping develop my

understanding of statistics and fishes. I would also like to thank the Odum School of

Ecology and the Joseph W. Jones Ecological Research Center for funding me through

this endeavor. The opportunity to live and work in such a magical part of the world is

something I will always look back on fondly.

I would also like to thank those at the Jones Center who helped make this project

possible. Especially, Denzell Cross, Meg Hederman, and Robert Ritger we made it

through the heat, the gnats, the mosquitoes, and the snakes, all while singing songs and

dancing the electrofish dance! Denzell, you were with me from day one, and words can’t

describe how happy I am to see you at Odum in pursuit of your PhD. Chelsea Smith, you

are my live version of stackexchange, thank you for always being there to bounce ideas

off and help me with statistics. Camille Herteux and Cara McElroy, thank you for all of

the laughs and little distractions that helped keep me sane.

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A final thanks to d.w. giddens, my husband and partner in all else, without whom

I would rarely have taken a step back to appreciate all that is wonderful in the Universe. I

give my deepest love and appreciation for the encouragement and sacrifices you gave and

made throughout this project.

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TABLE OF CONTENTS

Page

ACKNOWLEDGEMENTS .................................................................................................v

LIST OF TABLES ........................................................................................................... viii

LIST OF FIGURES .......................................................................................................... xii

CHAPTER

1 LITERATURE REVIEW AND SUMMARY OF OBJECTIVES ...................1

2 STREAM DRYING AND FISH OCCUPANCY DYNAMICS IN THE

ICHAWAYNOCHAWAY CREEK BASIN ..................................................10

3 IDENTIFYING LIFE HISTORY TRAITS THAT PROMOTE FISH

SPECIES PERSISTENCE IN INTERMITTENT STREAMS ......................73

4 CONCLUSIONS AND SUMMARY ...........................................................140

APPENDICES

A SPECIES OCCURRENCE OF TAXA FOUND FOR CHAPTER 2 ..........146

B DETAILED DESCRIPTION OF OCCUPANCY MODEL ........................157

C SPECIES AND AGE CLASS OCCURRENCES OF TAXA FOUND AT

INTERMITTENT STREAMS FOR CHAPTER 2 .....................................163

D R CODE USED FOR DYNAMIC OCCUPANCY MODEL ......................165

E INDICATOR ANALYSIS AND CLASSIFICATION FOR SPECIES

STRATEGISTS ENDPOINTS .....................................................................169

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LIST OF TABLES

Page

Table 2.1: Summary statistics of water quality data obtained in 90 isolated pools

monitored in 12 stream sites in the Ichawaynochaway Creek basin, June through

September 2015 and 2016, followed by their values centered and scaled around

zero by subtracting the mean and dividing by the standard deviation. Scaled

values were used as covariate effects on observed fish occurrence in isolated

pools. Numbers of isolated pools (n), and mean covariate value are shown along

with standard deviation (SD), standard error (SE), minimum (Min) and

Maximum. ..............................................................................................................43

Table 2.2: Effects of covariates on regression coefficients for persistence, colonization,

and detection from multi-taxa, dynamic occupancy models using a time-series

(2015-2017) of detection for adults of 21 species and juveniles of 25 species in

the Ichawaynochway Creek basin. Stream state, sampling method and cool season

use binary coding. Distance is the distance of the study site from the nearest

downstream perennial stream, standardized by subtracting the mean and dividing

by the standard deviation. Effects of indicator-species covariates (Intermittent

Nonindicative species and Perennial species, with Intermittent species as the

baseline) on regression coefficients are shown for persistence during the number

of weeks a site was isolated (Weeks Slack) and for colonization after resumption

of flow (Weeks Flowing). Variance terms are for random effects of site and date

ix

(“surveys”) on intercepts for persistence, colonization, and detection, and on

species-slopes for relations between persistence and Weeks Slack, and between

colonization and Weeks Flowing. All values are on the logit scale, and show the

posterior means and 95% credible intervals (in parentheses) ................................44

Table 2.3: Modeled effects of environmental covariates on probability of observed

occurrence of adult fishes in 90 isolated stream pools in the Ichawaynochway

Creek basin, 2015-2016. Values are the estimated effects on the log-odds of

occurrence (95% confidence intervals) for predictor variables (values were

centered and scaled around zero by subtracting the mean and dividing by the

standard deviation) and the estimated random variance in intercepts attributable to

species, surveys, and pools (nested within repeated survey of a pool), and in

slopes attributable to species ..................................................................................45

Table 2.4: Modeled effects of environmental covariates on probability of observed

occurrence of juvenile fishes in 90 isolated stream pools in the Ichawaynochway

Creek basin, 2015-2017. Values are the estimated effects on the log-odds of

occurrence (95% confidence intervals) for predictor variables (values were

centered and scaled around zero by subtracting the mean and dividing by the

standard deviation) and the estimated random variance in intercepts attributable to

species, surveys, and pools (nested within repeated survey of a pool), and in

slopes attributable to species ..................................................................................46

Table 3.1: Ovary and oocyte stages and descriptions of development based on oocyte

size, coloration, yolk condition, and physical location within the ovum modified

from Heins and Rabito (1986) and Heins and Baker (1987) ...............................106

x

Table 3.2: Standard lengths of males and females of four species assessed for

reproductive development (>25mm) from seven study sites in the

Ichawaynochaway Creek Basin from May 2016- April 2017. Numbers of

individuals (n), and mean lengths are shown along with standard deviation (SD),

standard error (SE), minimum (Min) and maximum (Max) ................................107

Table 3.3: Results from Chi-square tests of significance, which were performed

separately on sexually mature individuals and non-reproductive individuals.

Significant differences are marked with an * (p>.05) between the expected sex

ratio of 1:1 and the observed sex ratio for males and females of a given

species ..................................................................................................................108

Table 3.4: Standard lengths of mature males and females (Mature, Mature Ripening, or

Ripe) of four species assessed for reproductive development from seven study

sites in the Ichawaynochaway Creek Basin from May 2016- April 2017. Numbers

of individuals (n), and mean lengths are shown along with standard deviation

(SD), standard error (SE), minimum (Min) and maximum (Max) ......................109

Table 3.5: Standard lengths of all individuals captured during survey periods for length

distributions at thirteen study sites in the Ichawaynochaway Creek Basin from

May 2016- April 2017. Individuals within the genus Notropis that were not

identifiable in the field were categorized as Notropis sp. Numbers of individuals

(n), and mean lengths are shown along with standard deviation (SD), standard

error (SE), minimum (Min) and maximum (Max) ...............................................110

Table 3.6: Summary statistics for egg size (mm) of mature, mature ripening, and ripe

females assessed for reproductive investment. Each individual had twenty eggs

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measured, where n is the number of individuals assessed per species. Numbers of

individuals (n), and mean lengths are shown along with standard deviation (SD),

standard error (SE), minimum (Min) and maximum (Max) ................................111

Table 3.7: Species strategy weight and assignment for Soft Classification for

opportunistic strategist (OS), periodic strategist (PS), and equilibrium strategist

(ES) strategist end points calculated following Mims et al. (2010) for species

identified in the Ichawaynochaway Creek Basin (June 2015-January 2017).

Species strategy weight was assessed using only the life history traits of the four

cyprinid species ....................................................................................................112

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LIST OF FIGURES

Page

Figure 2.1: Locations of intermittent streams study sites (marked with squares) that were

surveyed to assess shifts in community assemblages, species-specific rates of

persistence and colonization in dynamic occupancy models, and probability of

persistence in isolated pools within the Ichawaynochaway Creek Basin during

2015-2017. Perennial sites (marked with triangles) indicate streams where

published and unpublished data were obtained using similar survey methods, and

were used to assess differences in community assemblages between intermittent

and perennial streams .............................................................................................47

Figure 2.2: Discharge, water temperature, and air temperature at Spring Creek near

Leary, GA (USGS gage 02354475). Periods where discharge is at or near zero

represent timing of intermittency, during which isolation or complete drying

occurred ..................................................................................................................48

Figure 2.3: Changes in stream state used as covariates to estimate persistence and

colonization in intermittent streams, where “flowing” represents stream state

where discharge is >0, “isolated” represents a pool that is isolated from upstream

or downstream movement of fishes (e.g. a small pool), and “isolated-open”

represents an isolated pool that is open to upstream or downstream movement of

fishes (e.g. a big pool) ............................................................................................49

xiii

Figure 2.4(a-c): Non-metric multi-dimensional scaling (NMDS) ordination of stream

samples based on Brays-Curtis dissimilarities in species occurrences. Ellipses

represent centroids and 95% confidence intervals for mean scores for samples

from perennial and intermittent streams. Each graphic represents 2 of the 3

dimensions in two-dimensional space ...................................................................50

Figure 2.5: Time series of changes in stream state for 12 intermittent study sites in the

Ichawaynochaway Creek Basin, June 2015 to January of 2017 ............................52

Figure 2.6: Posterior mean probabilities of taxa-specific detection and 95% confidence

intervals for adults of species found in >5% of surveys averaged over 12 study

sites in the Ichawaynochaway Creek Basin. Values plotted are estimates for each

of the 21 species using a multi-taxa, dynamic occupancy model. Taxa are

identified by the first three letters of their genus and species ................................53


intervals for juveniles of species found in >5% of surveys averaged over 12 study




Figure 2.8: Posterior mean probabilities of taxa-specific persistence and 95% confidence

intervals for adults of species found in >5% of surveys averaged over 12 study




xiv


intervals for juveniles of species found in >5% of surveys averaged over 12 study




Figure 2.10: Posterior mean probabilities of taxa-specific colonization and 95%

confidence intervals for adults of species found in >5% of surveys averaged over

12 study sites in the Ichawaynochaway Creek Basin. Values plotted are estimates

for each of the 21 species using a multi-taxa, dynamic occupancy model. Taxa are



confidence intervals for juveniles of species found in >5% of surveys averaged

over 12 study sites in the Ichawaynochaway Creek Basin. Values plotted are

estimates for each of the 25 species using a multi-taxa, dynamic occupancy

model. Taxa are identified by the first three letters of their genus and species .....58

Figure 2.12: Average mean of probability of persistence for adult fish in isolated pools,

plotted in relation to duration of pool isolation. Probabilities are plotted for 21

species estimated using a multi-taxa, dynamic occupancy model applied to 26

periods of continuous isolation at 12 study sites in the Ichawaynochaway Creek

Basin. Black lines indicate the species-specific means of persistence and red lines

indicate the means for each of the three species types ...........................................59

Figure 2.13: Average mean of probability of persistence for juvenile fish in isolated

pools, plotted in relation to duration of pool isolation. Probabilities are plotted for

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25 species estimated using a multi-taxa, dynamic occupancy model applied to 26




Figure 2.14: Average mean of probability of colonization for adult fish, plotted in relation

to duration of flow since isolation or complete drying. Probabilities are plotted for

21 species estimated using a multi-taxa, dynamic occupancy model applied to 26




Figure 2.15: Average mean of probability of colonization for juvenile fish, plotted in

relation to duration of flow since isolation or complete drying. Probabilities are

plotted for 25 species estimated using a multi-taxa, dynamic occupancy model

applied to 26 periods of continuous isolation at 12 study sites in the

Ichawaynochaway Creek Basin. Black lines indicate the species-specific means of

persistence and red lines indicate the means for each of the three species types ..62

Figure 2.16: Modeled probability of observed occurrence of adults in relation to

maximum total ammonia (ug/L) in 90 isolated pools samples in the

Ichawaynochaway Creek Basin, 2015-2016. Plot shows mean and 95%

confidence intervals ...............................................................................................63

Figure 2.17: Modeled observed occurrence of adults in relation to maximum depth (m) in

90 isolated pools samples in the Ichawaynochaway Creek Basin, 2015-2016. Plot

shows mean and 95% confidence intervals ...........................................................64

xvi

Figure 2.18: Modeled observed occurrence of juveniles in relation to maximum depth (m)

in 90 isolated pools samples in the Ichawaynochaway Creek Basin, 2015-2016.

Plot shows mean and 95% confidence intervals ....................................................65

Figure 2.19: Modeled observed occurrence of juveniles in relation to maximum ammonia

(u/gL) in 90 isolated pools samples in the Ichawaynochaway Creek Basin, 2015-

2016. Plot shows mean and 95% confidence intervals ..........................................66

Figure 2.20: Modeled observed occurrence of juveniles in relation to dissolved oxygen

(mg/L) in 90 isolated pools samples in the Ichawaynochaway Creek Basin, 2015-

2016. Plot shows mean and 95% confidence intervals ..........................................67

Figure 2.21: Species-specific random effects on the intercept and slope of modeled

observed occurrence of juveniles in relation to maximum depth in 90 isolated

pools samples in the Ichawaynochaway Creek Basin, 2015-2016. Plots show

means and 95% confidence intervals .....................................................................68


observed occurrence of juveniles in relation to maximum ammonia in 90 isolated



Figure 2.23: Species-specific random effects on the intercept and slope of a modeled

observed occurrence of juveniles in relation to dissolved oxygen in 90 isolated




observed occurrence of adults in relation to maximum depth in 90 isolated pools

xvii

samples in the Ichawaynochaway Creek Basin, 2015-2016. Plots show means and

95% confidence intervals. ......................................................................................71


observed occurrence of adults in relation to maximum ammonia in 90 isolated



Figure 3.1: Mean probabilities of species-specific persistence for four adult cyprinid

species found intermittent streams using a multi-taxa, dynamic occupancy model

over the weekly duration of isolation. Species-specific rates of persistence were

averaged for each species over 12 study sites and 14 weeks of continuous

isolation in the Ichawaynochaway Creek Basin from 2015-2017 (Chapter 2).

Species-specific persistence rates were used to develop hypotheses for life history

trait differences among N. harperi, N. petersoni, N. texanus, and P.

grandipinnis .........................................................................................................113

Figure 3.2: Locations of thirteen study sites within the Ichawaynochaway Creek Basin

that were used to measure length distributions for four cyprinid species and to

obtain individuals for analyzing diet and reproductive characteristics, May 2016-

April 2017. Apart from Brantley Creek (the most north easterly circle) all survey

streams are intermittent and experienced isolation or complete drying during the

survey period ........................................................................................................114

Figure 3.3: Standard length distribution to the nearest millimeter for all P. grandipinnis

individuals found at thirteen study sites within the Ichawaynochaway Creek Basin

xviii

from May 2016- April 2017, plotted by Julian date. The horizontal line represents

the minimum reported length at maturity (34.82 mm standard length) ...............115

Figure 3.4: Observed GSI for P. grandipinnis females (upper left) and males (upper right)

and standard length for females (bottom left) and males (bottom right) of

individuals assessed for reproductive state from within the Ichawaynochaway

Creek Basin from May 2016- April 2017, plotted by Julian date. For females, the

black symbols for MA, MR, and RE represent reproductively mature individuals

and the grey symbols for LA, EM, and LM represent reproductively latent or

immature individuals. For males, black symbols indicate mature males and the

grey symbols indicate latent or immature individuals. The horizontal line for

standard length represents minimum observed length of reproductively mature

females (34.82) and males (39.32)………………………………………… 116

Figure 3.5: Standard length distribution to the nearest millimeter for all N. harperi

collected at thirteen study sites within the Ichawaynochaway Creek Basin from

May 2016- April 2017, plotted by Julian date. The horizontal line represents the

minimum reported length at maturity (34.82mm standard length) ......................117

Figure 3.6: Observed GSI for N. harperi females (upper left) and males (upper right) and

standard length for females (bottom left) and males (bottom right) of individuals

assessed for reproductive state from within the Ichawaynochaway Creek Basin

from May 2016- April 2017, plotted by Julian date. For females, the black

symbols for MA, MR, and RE represent reproductively mature individuals and

the grey symbols for LA, EM, and LM represent reproductively latent or


xix



females (38.26) and males (32.35) .......................................................................118

Figure 3.7: Standard length distribution to the nearest millimeter for all N. petersoni



minimum reported length at maturity (46.84mm standard length) ......................119

Figure 3.8: Observed GSI for N. petersoni females (upper left) and males (upper right)










Figure 3.9: Standard length distribution to the nearest millimeter for all N. texanus



minimum reported length at maturity (49.2mm standard length) ........................121

Figure 3.10: Observed GSI for N. texanus females (upper left) and males (upper right)


xx







standard length represents minimum observed length reproductively mature


Figure 3.11: Discharge at USGS 02354475 Spring Creek near Leary, GA (left y-axis)

during the survey period. Light gray regions indicate the Palmer Drought Index

for the region (National Integrated Drought Information System, NIDIS;

www.drought.gov). While drought index values were exceptional from October to

December of 2016, values were not exceptional for summer and early fall moths

(July-September) ..................................................................................................123

Figure 3.12: The Tukey adjusted comparison of trends of slopes for reproductive timing

of individuals of four cyprinid species using a ANCOVA. Points indicate the

slope of the probability curves for a given species with error bars indicating the

95% confidence intervals. Results are given on the response scale (the natural log

of a given date), where date 1is January 1st. Means sharing a letter are not

significantly different by Tukey-adjusted mean separations ...............................124

Figure 3.13: Probability curves of presence of mature individuals of a given species over

a year time span. Normal confidence intervals are constructed on the link scale,

http://www.drought.gov/

xxi

and then back-transformed to the response scale. The numeric date of 1 represents

the first day of the calendar year (January 1st) .....................................................125

Figure 3.14: The least square means of the standard length for mature individuals of four

cyprinid species using ANOVA. Points indicate the least square mean of the

standard length by species; error bars indicate the 95% confidence intervals using

Tukey-adjusted comparisons. Means sharing a letter are not significantly different

by Tukey-adjusted mean separations ...................................................................126

Figure 3.15: The simple linear regression of the natural log of eviscerated mass and the

natural log of standard length for all fishes of an individual species combined

were: P. grandipinnis, log(mass)= -12.55+3.40*log(length), F1,150=5846,

p=<.001; N. harperi, log(mass)=-11.69+3.16*log(length), F1,196=3987, p=<.001;

N. petersoni, log(mass)=-12.12+3.26*log(length), F1,86=7519, p=<.001; N.

texanus, log(mass)=-11.83+3.20*log(length), F1,81=2522, p=<.001 .................127

Figure 3.16: The least square means of the eviscerated mass for mature males of four

cyprinid species using ANCOVA. Points indicate the least square mean of the

eviscerated mass of an individual and error bars indicate the 95% confidence

intervals using Tukey-adjusted comparisons. Means sharing a letter are not

significantly different by Tukey-adjusted mean separations ...............................128

Figure 3.17: The least square means of the gonadosomatic index values (GSI) for mature

females of four cyprinid species using an ANCOVA. Points indicate the lease

square mean of the GSI of an individual and error bars indicate the 95%

confidence intervals using Tukey-adjusted comparisons. Means sharing a letter

are not significantly different by Tukey-adjusted mean separations. The

xxii

ANCOVA was fit with a fixed effect of a given species, a covariate of standard

length, and a response variable of the GSI of an individual female fish .............129


males of four cyprinid species using an ANCOVA. Points indicate the lease

square mean of the GSI of an individual and error bars indicate the 95%

confidence intervals using Tukey-adjusted comparisons. Means sharing a letter

are not significantly different by Tukey-adjusted mean separations. The

ANCOVA was fit with a fixed effect of a given species, a covariate of standard

length, and a response variable of the GSI of an individual male fish ................130

Figure 3.19: The least square means of gonad weight for mature females of four cyprinid

species using an ANCOVA. Points indicate the lease square mean of the gonad

weight of an individual and error bars indicate the 95% confidence intervals using


by Tukey-adjusted mean separations. The ANCOVA was fit with a fixed effect of

a given species, a covariate of standard length, and a response variable of the

gonad weight of an individual female fish ...........................................................131

Figure 3.20: The least square means of gonad weight for mature males of four cyprinid

species using an ANCOVA. Points indicate the lease square mean of the gonad

weight of an individual and error bars indicate the 95% confidence intervals using


by Tukey-adjusted mean separations. The ANCOVA was fit with a fixed effect of

a given species, a covariate of standard length, and a response variable of the

gonad weight of an individual male fish ..............................................................132

xxiii

Figure 3.21: The least square means of egg diameter for four cyprinid species using a

nested ANOVA. Points indicate the lease square mean and error bars indicate the

95% confidence intervals using Tukey-adjusted comparisons. Means sharing a

letter are not significantly different by Tukey-adjusted mean separations. The

ANOVA was fit with a fixed effect of species and with egg diameter nested

within the individual fish it was collected ...........................................................133

Figure 3.22: The least square means of four cyprinid species using a nested ANCOVA.

Points indicate the lease square mean and error bars indicate the 95% confidence

intervals using Tukey-adjusted comparisons. Means sharing a letter are not

significantly different by Tukey-adjusted mean separations. The ANCOVA was

fit with a fixed effect of species, a covariate of species length, with egg diameter

nested within the individual fish it was collected ................................................134

Figure 3.23: Ternary plot illustrating trilateral life history trade-offs in traits among

commonly occurring species within the Ichawaynochaway Creek basin. Axis

scores indicate degree of species affiliation with opportunistic, periodic, or

equilibrium strategists. Species points are represented by which stream type they

are associated with. The target species (P. grandipinnis, N. harperi, N. petersoni,

and N. texanus), represented by cross symbols, score highest on the opportunistic

axis when evaluated in the context of this assemblage ........................................135

Figure 3.24: Ternary plot illustrating trilateral life history trade-offs in traits among four

cyprinid species, where axis scores indicate degree of species affiliation with

opportunistic, periodic, or equilibrium strategists ...............................................136

xxiv

Figure 3.25: Index of relative importance for samples of individual assessed for diet

during flowing states and isolated states in the Ichawaynochaway Creek Basin

(May 2016- July 2016). Each of the twenty categories represents the total percent

of the IRI for a given sample where the number of individuals per sample ranged

from one to nine. Prey categories were assigned to family or to the lowest known

taxonomic level ....................................................................................................137

Figure 3.26: Index of relative importance for subsamples of individuals assessed for diet

during flowing states and isolated states in the Ichawaynochaway Creek Basin

(May 2016- July 2016). Each of the twenty categories represents the total percent

of the IRI for a given subsample where the number of individuals per subset

ranged from one to five. Categories were assigned based on whether diet taxa

identified were aquatic, terrestrial, or an unknown category of “other” (e.g.

detritus, eggs, and oligochaetes) ..........................................................................138

Figure 3.27: Non-metric multi-dimensional scaling (NMDS) ordination of %IRI for diet

categories of all individuals assessed. Grouping is based by species and the stream

state when species were captured. Hollow symbols represent diet components for

an individual within a subsample for P. grandipinnis in parametric space and

solid symbols represent diet components for an individual N. harperi. Shapes of

symbols represent the stream state when an individual was captured, with

triangles representing periods of flowing and circles are periods of isolation.

Ellipses represent centroids and 95% confidence intervals for scores from

grouping of species and stream state ...................................................................139

1

CHAPTER 1

LITERATURE REVIEW AND SUMMARY OF OBJECTIVES

Literature Review

Water abstraction for irrigation affects hydrology by lowering stream flows, leading to

significant changes in rivers worldwide (Palmer et al. 2008a, Arthington et al. 2014, Walker and

Adams 2016). Decreased groundwater levels alter the quantity and quality of surface waters,

cause changes in riparian communities, and have negative consequences for the persistence of

many aquatic species (Falke et al. 2011, Rugel et al. 2012). Cases of human-caused stream

drying are increasing in frequency, and are characterized by abrupt changes from perennial to

intermittent flow regimes (Larned et al. 2010). There has been growing interest in understanding

the connectivity between surface water and groundwater, and how it affects the biology and

hydrology of flowing waters (Rugel et al. 2012). Flow intermittence can lead to fishery declines,

loss of migratory pathways, altered nutrient cycles, and reductions or losses of other ecosystem

services (Jackson et al. 2001, Larson et al. 2009). In regions where groundwater both supports

stream baseflow and is a major water resource, the careful management of groundwater is crucial

to the protection of flow regimes (Woessner 2000).

Groundwater Use and River Flow

Worldwide, 2.5 billion people depend solely on groundwater resources to satisfy their

daily water needs, and hundreds of millions of farmers rely on groundwater to sustain their

2

livelihoods (UNESCO 2009). Groundwater levels are declining in several of the world’s most

intensely cultivated agricultural areas and around numerous mega-cities (UNESCO 2015). As

climate change alters rainfall patterns, and in many areas increases the frequency and duration of

droughts, the amount of water necessary for human use will inevitably exceed water availability.

On average, the southeastern US has experienced drought conditions every 5-10 years since

1895. In Georgia, localized droughts have occurred even more frequently, approximately every

2-3 years (Baker 2000). Climate models predict an increased frequency of precipitation

extremes, and a shift in rainfall from the growing season to the winter and early spring (Ingram

2013).

The Flint River Basin (FRB), located in southwestern Georgia, has experienced an

increased demand on water resources resulting from population expansion in the upper basin,

and irrigation expansion in the lower basin (Golladay and Hicks 2013, Golladay et al. 2016).

Irrigated farmlands in southwestern Georgia have increased from 0.13 million acres in 1976

(Pollard et al. 1978), to almost 1.2 million acres in 2014 (USDA 2014). The 2012 Census of

Agriculture indicated that of the total irrigated acreage in Georgia, groundwater and surface

waters contributed 80% and 20% respectively (USDA 2014). Groundwater withdrawals in

Georgia for agricultural use have increased more than 3070% between 1970 and 1990 following

the introduction of center pivot irrigation (Marella et al. 1993). The amount of irrigation water

needed to support agriculture varies from year to year depending on rainfall during the growing

season. Consequently, trends in agricultural irrigation will affect Georgia's future efforts to

manage its water resources (Harrison 2001). Long-term climate data show no change in average

annual rainfall in the lower FRB; however, minimum flows in USGS stream gage records show

substantial declines since the development of irrigation (Rugel et al. 2012). Current rates of

3

human water use are likely unsustainable, causing an increase in severity and duration of low

flows during droughts throughout the FRB, and are likely to pose a significant threat to stream

health and biological diversity (Golladay and Hicks 2013).

Streamflow Shifts to Intermittency

Climate-driven flow intermittence has increased in some regions of the US within the last

century (Palmer et al. 2008a, Falke et al. 2011), in particular the Coastal Plain of Georgia, and is

projected to continue in the near future (Larned et al. 2010, Golladay and Hicks 2013). Flow

intermittence caused by climate change is likely to occur more gradually than intermittence

caused by groundwater pumping, and is in phase with regional drying trends (Larned et al.

2010). During periods of water scarcity, streams are easily fragmented due to their linear and

hierarchical structure (Fagan 2002). As flow ceases, connectivity is quickly lost and remaining

wetted habitat becomes increasingly isolated (Bunn and Arthington 2002). Intermittent streams

may partially or completely dry for weeks or months during the year, on a roughly predictable

basis (Arthington et al. 2014). Intermittent streams are a natural part of the landscape, but some

streams are experiencing longer periods of isolation or complete drying. Shifts in intermittency

may be, in part, driven by drought conditions, but groundwater pumping is likely contributing to

longer periods of isolation.

The Ichawaynochaway Creek Basin, located in the lower FRB, southwestern GA, is

dominated by irrigated agriculture. Over 35,000 ha of the land area is irrigated, with 59%

irrigated with groundwater and the remainder irrigated from surface water (Couch and

McDowell 2006). For planning purposes, water from center pivot irrigation is considered 100%

consumptive, with no water return to surface waters or aquifers. Increases in water use have had

4

negligible effect on average annual streamflow in Ichawaynochaway Creek, but have been

associated with substantial reduction in summer baseflows, increasingly so during drought or

low precipitation years (Rugel et al. 2016).

Stream Fish and Their Responses to Increasing Intermittency

The southeastern US is noted for its aquatic faunal diversity, having the most diverse

freshwater fish fauna in North America (Burr and Mayden 1993). The American Fisheries

Society lists approximately 662 native freshwater fishes present in drainages spanning Virginia

to Texas, with roughly 28% of species deemed of conservation concern (Warren et al. 2000). As

alterations to the hydrologic cycle increase, there is a growing need to understand how drought

and groundwater withdrawals affect freshwater biodiversity and biotic integrity in streams.

Environmental variability is a natural part of aquatic ecosystems and influences the structure of

aquatic communities (Resh et al. 1988, Poff and Allan 1995). Non-sustainable water withdrawals

from aquifers and streams cause drastic alterations to the biota of aquatic ecosystems (Magoulick

and Kobza 2003, Falke et al. 2011, Skoulikidis et al. 2011). Freshwater fishes are one of the most

threatened faunal groups and are expected to be among the most severely affected by climate

change (Palmer et al. 2008a, Beatty et al. 2014). Non-game fishes in particular have historically

been under-studied and overlooked by natural resource managers, with many species becoming

imperiled before conservation efforts are focused on them (Cooke et al. 2005).

Human-caused declines in fish populations have been attributed primarily to habitat loss,

stream impoundments, channelization, increased sedimentation, introduced species, and

pollution. Increased water withdrawals due to expansions in population, urbanization, irrigated

agricultural acreage, and industrialization have also been associated with species decline in

5

Georgia (Tabit and Johnson 2002). Human-caused changes in streamflow may contribute to

flows outside the range of historical variability and could have substantial consequences for river

ecosystems and human welfare (Palmer et al. 2007). Stream fish assemblages may also change in

response to streamflow alteration. Groundwater withdrawal is expected to alter stream fish

assemblages because of increased severity and duration of low-flow and no-flow events, as has

occurred throughout the lower FRB (Rugel et al. 2012).

The availability of suitable refuge habitat for stream fishes may fluctuate dramatically

during stream drying, resulting in spatial and temporal variability of species occurrences (Palmer

et al. 2007). Long-term exposure to non-lethal high temperatures can make fish more susceptible

to sources of mortality such as disease and predation, and ultimately reduce population

persistence (Bevelhimer and Bennett 2000). Drought selects for species that demonstrate

resistance or resilience to effects of low flow and stream drying (Resh et al. 1988, Lake 2011).

Species-specific responses to periods of intermittency, resumption of flow, and the availability of

refugia should determine resulting community composition and rate of recovery. Additionally,

life history traits may be useful in understanding species persistence during intermittency.

Project Objectives

This study was designed to assess the effects of intermittency on fishes within the

Ichawaynochaway Creek Basin (ICB), a major tributary to the lower Flint River. I examined

assemblage variation across a gradient of flow permanence, isolation, and reach position within

the ICB to quantify species-specific responses to changes in abiotic conditions. By monitoring

site level hydrologic effects within streams that periodically cease flowing or dry completely, I

estimated rates of species-specific occurrence, persistence, and colonization. I also tested the

6

effects of environmental variables on species occurrence in isolated pools. Finally, I analyzed

life history traits of four common cyprinid species, each differing in their ability to persist in

intermittent streams, and identified traits most closely correlated with species persistence. The

goal of this work was to develop an analytical basis for understanding and predicting fish faunal

changes to increasing flow intermittency in the ICB, with potential applications for other systems

having similar faunal and flow characteristics.

7

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ecological quality and reconciliation ecology. River Research and Applications 30:1209-

1215.

Baker, T. L. 2000. Survival, habitat use, movement patterns, and thermal refuge selection of

adult striped bass in Lake Blackshear, GA. University of Georgia, Masters Thesis.

Beatty, S. J., D. L. Morgan, and A. J. Lymbery. 2014. Implications of climate change for

potamodromous fishes. Global Change Biology 20:1794-1807.

Bevelhimer, M., and W. Bennett. 2000. Assessing cumulative thermal stress in fish during

chronic intermittent exposure to high temperatures. Environmental Science and Policy

3:211-216.

Bunn, S. E., and A. H. Arthington. 2002. Basic principles and ecological consequences of altered

flow regimes for aquatic biodiversity. Environmental Management 30:492-507.

Burr, B. M., and R. L. Mayden. 1993. Phylogenetics and North American freshwater fishes.

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Cooke, S. J., C. M. Bunt, S. J. Hamilton, C. A. Jennings, M. P. Pearson, M. S. Cooperman, and

D. F. Markle. 2005. Threats, conservation strategies, and prognosis for suckers

(Catostomidae) in North America: insights from regional case studies of a diverse family

of non-game fishes. Biological Conservation 121:317-331.

Couch, C. A., and R. J. McDowell. 2006. Flint River Basin regional water development and

conservation plan. Georgia Department of Natural Resources-Environmental Protection

Division.

Fagan, W. F. 2002. Connectivity, fragmentation, and extinction risk in dendritic

metapopulations. Ecology 83:3243-3249.

Falke, J. A., K. D. Fausch, R. Magelky, A. Aldred, D. S. Durnford, L. K. Riley, and R. Oad.

2011. The role of groundwater pumping and drought in shaping ecological futures for

stream fishes in a dryland river basin of the western Great Plains, USA. Ecohydrology

4:682-697.

Golladay, S. W., and D. W. Hicks. 2013. Indicators of long term hydrologic change in the Flint

River. Proceedings of the 2013 Georgia Water Resources Conference. University of

Georgia. Athens, GA.

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Golladay, S. W., K. L. Martin, J. M. Vose, D. N. Wear, A. P. Covich, R. J. Hobbs, K. D.

Klepzig, G. E. Likens, R. J. Naiman, and A. W. Shearer. 2016. Review and synthesis:

Achievable future conditions as a framework for guiding forest conservation and

management. Forest Ecology and Management 360:80-96.

Harrison, K. A. 2001. Agricultural irrigation trends in Georgia. Proceedings of the 2001 Georgia

Water Resources Conference. Institute of Ecology. The University of Georgia. Athens,

Georgia.

Ingram, K. T. 2013. Climate of the southeast United States: variability, change, impacts, and

vulnerability. NCA Regional Input Reports, Washington, DC.

Jackson, R. B., S. R. Carpenter, C. N. Dahm, D. M. McKnight, R. J. Naiman, and S. L. Postel.

2001. Water in a changing world. Ecological Applications 11:1027-1045.

Lake, P. S. 2011. Drought and Aquatic Ecosystems: Effects and Responses. John Wiley & Sons.

Larned, S. T., T. Datry, D. B. Arscott, and K. Tockner. 2010. Emerging concepts in temporary-

river ecology. Freshwater Biology 55:717-738.

Larson, E. R., D. D. Magoulick, C. Turner, and K. H. Laycock. 2009. Disturbance and species

displacement: different tolerances to stream drying and desiccation in a native and an

invasive crayfish. Freshwater Biology 54:1899-1908.

Magoulick, D. D., and R. M. Kobza. 2003. The role of refugia for fishes during drought: a

review and synthesis. Freshwater Biology 48:1186-1198.

Marella, R. L., J. L. Fanning, and W. S. Mooty. 1993. Estimated use of water in the

Apalachicola-Chattahoochee-Flint River Basin during 1990, with state summaries from

1970 to 1990. US Department of the Interior, US Geological Survey.

Palmer, M. A., Dennis Lettenmaier, N. L. Poff, S. Postel, B. Richter, and R. Warner. 2007.

Adaptation options for climate-sensitive ecosystems and resources: wild and scenic

rivers. Washington, DC: US Climate Change Science Program.

Palmer, M. A., C. A. R. Liermann, C. Nilsson, M. Floerke, J. Alcamo, P. S. Lake, and N. Bond.

2008. Climate change and the world's river basins: anticipating management options.

Frontiers in Ecology and the Environment 6:81-89.

Poff, N. L., and J. D. Allan. 1995. Functional-organization of stream fish assemblages in relation

to hydrological variability. Ecology 76:606-627.

Pollard, L. D., R. G. Grantham, and J. H. E. Blanchard. 1978. A preliminary appraisal of the

impact of agriculture on ground-water availability in southwest Georgia. U.S. Geological

Survey Water-Resources Investigations Report 79:21.

9

Resh, V. H., A. V. Brown, A. P. Covich, M. E. Gurtz, H. W. Li, G. W. Minshall, S. R. Reice, A.

L. Sheldon, J. B. Wallace, and R. C. Wissmar. 1988. The role of disturbance in stream

ecology. Journal of the North American Benthological Society 7:433-455.

Rugel, K., S. W. Golladay, C. R. Jackson, and T. C. Rasmussen. 2016. Delineating

groundwater/surface water interaction in a karst watershed: Lower Flint River Basin,

southwestern Georgia, USA. Journal of Hydrology: Regional Studies 5:1-19.

Rugel, K., C. R. Jackson, J. J. Romeis, S. W. Golladay, D. W. Hicks, and J. F. Dowd. 2012.

Effects of irrigation withdrawals on streamflows in a karst environment: lower Flint

River Basin, Georgia, USA. Hydrological Processes 26:523-534.

Skoulikidis, N., L. Vardakas, I. Karaouzas, A. Economou, E. Dimitriou, and S. Zogaris. 2011.

Assessing water stress in Mediterranean lotic systems: insights from an artificially

intermittent river in Greece. Aquatic Sciences 73:581-597.

Tabit, C. R., and G. M. Johnson. 2002. Influence of urbanization on the distribution of fishes in a

southeastern upper piedmont drainage. Southeastern Naturalist 1:253-268.

UNESCO. 2009. Water in a Changing World. Routledge, Paris.

UNESCO. 2015. Water for a Sustainable World. Routledge, Paris.

Walker, R. H., and G. L. Adams. 2016. Ecological factors influencing movement of creek chub

in an intermittent stream of the Ozark Mountains, Arkansas. Ecology of Freshwater Fish

25:190-202.

Warren, M. L., B. M. Burr, S. J. Walsh, H. L. Bart, R. C. Cashner, D. A. Etnier, B. J. Freeman,

B. R. Kuhajda, R. L. Mayden, H. W. Robison, S. T. Ross, and W. C. Starnes. 2000.

Diversity, distribution, and conservation status of the native freshwater fishes of the

southern United States. Fisheries 25:7-31.

Woessner, W. W. 2000. Stream and fluvial plain ground water interactions; rescaling

hydrogeologic thought. Ground Water 38:423-429.

1Davis, J. L., M. C. Freeman, S. W. Golladay. To be submitted to Freshwater Biology

CHAPTER 2

STREAM DRYING AND FISH OCCUPANCY DYNAMICS IN THE


10

11

Abstract

Changes in climate and water demands can shift hydrologic regimes in streams and

consequently change aquatic faunal communities. Stream drying is natural process, with species

having a natural ability to respond. The point at which a disturbance, like stream drying, exceeds

the ability of a community to recover or causes a shift in assemblages is not well understood.

This study explores effects of stream intermittency and drying on the composition of biologically

diverse fish communities in the Ichawaynochaway Creek basin, southwest GA. I tested whether

faunal composition differed between perennial and intermittent streams, and which species were

strongly associated with each stream type. I used data for fish species collected in intermittent

stream surveys to analyze occupancy dynamics of adults and juveniles of commonly occurring

fishes, while accounting for incomplete species detection. I explored species-specific covariates

of changes in stream state, rates of persistence during isolation, and how quickly individuals

recolonize following the resumption of flow. I then tested the probability of occurrence of

individuals in isolated pools in response to environmental characteristics. Intermittent stream

communities were found to be a subset of perennial stream communities, with all species

identified found in perennial streams, but not in intermittent streams. Species with the lowest

persistence rates during isolation among adults and juveniles were species that more commonly

occur in perennial than intermittent streams. Colonization after the resumption of flow did not

significantly differ among species associated with perennial or intermittent streams. I found

support for the hypothesis that high concentrations of ammonia and low water depth decrease the

probability of fish occurrence in isolated pools. The incorporation of a species-specific rates

approach, via dynamic occupancy modeling, to stream intermittency is relatively novel, and can

12

help advance the mechanistic understanding of flow-ecology relationships, while also informing

environmental flow standards.

13

Introduction

Streamflow alteration due to the combined effects of water extraction and climate change

is recognized as a major threat to aquatic ecosystems. Evidence suggests that streamflow

intermittence has increased in the southeastern US (Palmer et al. 2008b, Falke et al. 2011),

including the Coastal Plain of Georgia, and is projected to continue increasing in the near future

(Larned et al. 2010, Golladay and Hicks 2013). The southeastern US is noteworthy for its

abundance and diversity of freshwater fishes. While various biotic and abiotic factors determine

fish community structure (Power et al. 1988), streamflow alteration can reduce suitability for

native fauna (Pringle et al. 2000). Natural resource managers face the challenge of understanding

projected increases in intermittency when working towards conserving biological integrity of

freshwater systems. Creating models that predict responses of fishes to extended low flows

requires an understanding of the relationships between stream flow, fish populations, and

community dynamics (Poff et al. 2010). This study focuses on fishes in a Gulf Coastal Plain

stream basin in southwestern Georgia, where streamflows are strongly influenced by agricultural

water withdrawals and droughts, to explore the effects of stream intermittency on the

composition of biologically diverse fish communities.

Increases in irrigated agriculture and domestic water consumption have generated

concerns for the sustainability of aquatic ecosystems (Dudgeon et al. 2006). Water resource

development affects the pattern of flow variability, including the timing, frequency, and

magnitude of flow events, which can act as important drivers of ecological processes in stream

ecosystems. Comparison of ecological patterns between natural and hydrologically-altered

streams yields flow-ecological response relationships, which can inform environmental flow

standards (Arthington et al. 2006) aimed at sustaining the quantity, quality, and timing of water

14

flows required by freshwater ecosystems (Poff et al. 2010). Estimates of flow effects on

demographic rates, including both persistence during periods of intermittency, and recolonization

following local extirpation, facilitates effective management through temporal projections of

biotic responses to flow alterations (Wheeler et al. 2017). Measured fish occupancy responses to

flow alteration can ultimately be used to improve water resource decision-making (Peterson and

Freeman 2016).

Streams are especially vulnerable to habitat fragmentation due to their linear and

hierarchical structure (Fagan 2002). Lowered streamflow can reduce sediment sorting, alter

stream temperature, reduce nutrient loading to downstream communities, and cause habitat

fragmentation and loss (Magoulick and Kobza 2003, Falke et al. 2012, Golladay and Hicks

2013). As flow diminishes, upstream-downstream connectivity may be quickly lost, while

channel drying can isolate remaining patches of inundated habitat (Bunn and Arthington 2002).

Drought and stream drying can negatively affect fish movement and survival in inundated

patches, and can decrease population persistence through local extirpation (Scheurer et al. 2003,

Falke et al. 2012). Generally, larger individuals are more susceptible to low-flow events

(McCargo and Peterson 2010) as predation pressure increases in shallow pools (Harvey and

Stewart 1991). Extended or unusually low flow can have negative effects on reproductive

success during summer months (Peterson and Shea 2014) and during the rearing period (Craven

et al. 2010). However, small flow pulses during drought have been found to increase young-of-

year survival (Katz and Freeman 2015).

Freshwater fishes are a globally imperiled faunal group and are expected to be among the

most severely affected by climate change (Palmer et al. 2008a, Beatty et al. 2014). Drought and

stream drying, through their effects on habitat quality and availability, alter fish population

15

dynamics (Magoulick and Kobza 2003, Hodges and Magoulick 2011, Hoch et al. 2015). For

example, summer water temperatures in Coastal Plain streams of the southeastern US may

exceed 31°C during July through August in low-discharge years (DeVries 2006). As water

temperatures increase, dissolved oxygen (DO) concentration decreases. The lowering of DO

often combines with other sublethal stressors including increased metabolic demand, and

decreased growth rates and activity. Fish response to such stressors depends on both the duration

of exposure and life history stage.

As inundated habitat contracts during drying, movement of fish is restricted. At this

point, net immigration into remaining wetted areas occurs, with some fish populations becoming

trapped in pools (Larned et al. 2010). This creates a metapopulation structure in which fishes

persist or become locally extirpated in isolated refugia, subsequently dispersing and recolonizing

reaches when flow resumes. Additionally, with flow resumption, recovering populations are

influenced by colonization from adjacent refugia or perennial reaches. Metapopulation theory

has increasingly been used to assess stream dwelling organisms, including mussels (Vaughn

2012, Shea et al. 2013), shrimps (Snyder et al. 2016), and fishes (Dunham and Rieman 1999,

Gotelli and Taylor 1999, Fagan 2002, Slack et al. 2004, Shea et al. 2015), including fishes within

southeastern streams (Freeman et al. 2013, Peterson and Shea 2014).

In this study, metapopulation dynamics provided a framework for assessing effects of

stream intermittency on biota, in this case, small-bodied fishes with limited mobility, that

compose species-rich assemblages. The first objective of this study was to use species

occurrence to model differences in fish community structure between a set of perennial and

intermittent streams in the southeastern Coastal Plain, and to identify species strongly associated

with each stream type. I hypothesized that a distinct subset of fishes populating perennial streams

16

would be found in streams known to experience periodic channel drying. The second objective

was to use repeated surveys to evaluate the species-specific and age-specific (i.e., adults

compared to juveniles) responses of individuals within intermittent streams to transitions

between flowing and isolated conditions. Specifically, I tested whether species strongly

associated with intermittent or perennial streams responded differently, and whether juveniles,

because of their smaller body size, would be less affected by streamflow reduction. For the

second objective, I hypothesized that (i) species common to intermittent streams would have a

higher persistence rate during isolation than species more common in perennial streams; (ii)

juveniles would have a higher persistence rate than adults, with juveniles of species common to

intermittent streams having the highest persistence; (iii) species common to intermittent streams

would recolonize reaches more quickly following resumption of flow than other species. The

third objective was to test environmental characteristics that may affect responses using species

and age-class occurrence in isolated pools. For the third objective, I hypothesized that (i) low

DO, elevated temperatures, high ammonia levels, and decreased maximum depth would reduce

fish occurrence; (ii) juveniles would have higher occurrence probabilities than adults when DO

was low, temperature and ammonia levels were high, and maximum depth was shallow.

Methods

Study Area

I used existing data and collected new observations on fish species occurrence and

metapopulation dynamics in the Ichawaynochaway Creek Basin (ICB), located in the lower Flint

River Basin (FRB), southwestern GA. The channels of major tributary streams within the lower

FRB, including Ichawaynochaway Creek, are incised into limestone bearing the upper Floridian

17

aquifer and tend to be perennial. Smaller streams, with channels perched above the aquifer, tend

to be intermittent (Hicks et al. 1987). The ICB contains the Chickasawhatchee Swamp, a

palustrine wetland located in southwest Georgia (Golladay and Battle 2001). The study area has

low topographic relief, and porous, sandy soils, which results in low stream drainage density.

During typical winters streamflow increases in response to extended storms (Hicks et al. 1987,

Albanese et al. 2007) and lower temperature and evapotranspiration rates (Torak and Painter

2006). Rainfall is evenly distributed throughout the year, but during the summer most

precipitation is lost through evapotranspiration, causing water table decline as groundwater

recharge is minimal. This results in riparian areas drying and streams decreasing to seasonal low-

flows (Golladay and Battle 2001) or periods of intermittency.

The Flint River Basin has experienced an increased demand on water resources resulting

from population expansion in the upper basin and irrigation expansion in the lower basin

(Golladay and Hicks 2013). Over the last four decades, the lower FRB has experienced

increasing water withdrawals from groundwater and surface waters. As a result, some streams

are shifting from historically perennial to intermittent. In particular, streams crossing the

Dougherty Plain, a recharge area for the upper Floridan aquifer region in the lower ICB, are

prone to drying during periods of low rainfall and high groundwater withdrawal (Opsahl et al.

2007). In contrast, streams in the upper ICB tend to be perennial. This mix of perennial and now-

intermittent streams provides a framework for assessing differences in fish assemblages

associated with shifts from perennial to intermittency, as well as to compare occupancy

dynamics between species and age-classes as streams shift between flowing and non-flowing

states.

18

Survey Methods

To measure species occurrence in intermittent streams, I surveyed twelve sites on eight

streams in the lower ICB over two years (June 2015- January 2017) during flowing and

intermittent periods. Study sites were located within the Chickasawhatchee Wildlife

Management Area, the Albany Nursery Wildlife Management Area, and at streams accessible at

bridge crossings. Sites were selected at differing distances from the nearest perennial stream, but

were otherwise similar in stream size, with second or third Strahler stream order. An initial

survey of eight sites on four streams was conducted in the summer and fall of 2015. Each stream

was surveyed at a downstream site near the confluence of the next adjoining stream, and at a site

located at least two river kilometers upstream. An additional four sites on four streams were

surveyed beginning in the spring of 2016 and continuing until after flow resumed in January of

2017 (Figure 2.1). At each site, two temperature loggers (HOBO UA-001-08 Pendant

Temperature Data Loggers, Onset Computer Corp., Bourne, Massachusetts) monitored air

temperature and water temperature at 30-minute intervals. Periods of isolation, drying, and

resumption of flow were assessed using a combination of USGS stream gage data (02354475),

visual monitoring, and diel changes in temperature (Figure 2.2).

I sampled fishes using a combination of backpack electrofishing and seining (2.4 m X 1.8

m; 3 mm mesh) at intervals ranging from every six weeks (unless a site became unwadeable)

during winter and early spring, to every one to three weeks when streams ceased to flow and

dried to isolated pools. Survey frequency increased during periods of stream drying to track

species persistence in isolated pools. To provide samples for estimating the probability of

detecting a species during a given survey, I sampled two adjacent stream reaches at each site that

I assumed contained the same species assemblage. In 2015, when streams were flowing, each

19

survey comprised multiple seine-sets in two 25-meter reaches, where two persons held the seine

in flowing water with the lead-line on the substrate, while one person disturbed water and bed

sediment while backpack electrofishing. Each reach was sampled with two passes, the first

upstream and the second downstream. For each pass, fish were removed, kept in aerated,

frequently exchanged water, and released at the end of the survey period. In 2016 and 2017, I

employed a single upstream pass for each survey reach in 80% of the samples, with the

remaining 20% randomly selected for two passes. This allowed me to account for the effect of

differing effort (1 vs. 2 passes) on species-specific detection. When streams dried to isolated

pools, I sampled using only seining to minimize fish stress caused by electrofishing. I seined

isolated pools until no new species were found in five consecutive seine hauls. On every

sampling date, fish were identified to species, counted, and measured. I assigned individuals to

either adult or juvenile (including young-of-year) age classes based on published minimum

lengths at maturity. Live fish were released at the end of the sampling within the reach where

they were captured. Any mortalities or unidentifiable individuals were collected and preserved in

10% formalin.

Community Assemblage Differences Between Intermittent and Perennial Streams

To assess differences in assemblage structure between intermittent and perennial streams,

I combined my data with other similarly collected data from perennial streams in the ICB

(McCargo 2004, McPherson 2005, M. C. Freeman, USGS, unpublished). McCargo (2004)

collected individuals in the ICB from 6 perennial sites from 2001 to 2003, with surveys

occurring in winter, spring, and summer. McPherson (2005) collected individuals in the ICB

from three perennial sites and one intermittent site from 2003-2004, with surveys occurring in

20

winter, spring, and summer. M. C. Freeman (USGS, unpublished) collected individuals from

seven perennial sites and three intermittent sites from 2011-2016 during summer and fall, though

not all sites were surveyed during each period. Individuals previously reported as Pteronotropis

hypselopterus were assigned to Pteronotropis grandipinnis; Gambusia holbrooki and Gambusia

affinis were assigned to Gambusia sp.; Erimyzon sucetta and Erimyzon oblongus were assigned

to Erimyzon sp.; Fundulus dispar and Fundulus escambiae were assigned to Fundulus sp.;

Lepomis punctatus and Lepomis miniatus were assigned to Lepomis punctatus X miniatus

(Appendix A). A total of 52 species were identified in published and unpublished data at 12

intermittent stream study sites and 12 perennial stream study sites (Appendix A). Sixteen species

never occurred at intermittent sites, with twelve considered rare (<5% of perennial surveys) and

removed from analysis. A total of 168 surveys in the intermittent sites and 56 surveys in the

perennial sites were used to assess assemblage structure after surveys with fewer than two

species detected were removed.

I performed a multivariate ordination of species occurrence data (as presence/absence)

for the 24 sites (Figure 2.1) using nonmetric multidimensional scaling (NMDS). The NMDS

used pairwise Brays Curtis dissimilarity measures to estimate distances between samples and to

test for differences between stream types. NMDS was performed with six and descending to

three dimensions using a random starting configuration and convergence determined through

Procrustes analysis. Stress was calculated for each convergent solution and the lowest number of

axes with the final stress of less than 0.2 was considered ecologically interpretable (Clarke

1993). I created 95% confidence ellipses around each centroid for intermittent and perennial

study sites. Permutational multivariate analysis of variance (PERMANOVA) was used to

examine differences in a priori defined reach types. Indicator species analysis was then

21

performed to identify taxa strongly associated with reach type (De Cáceres 2010). I classified

taxa significantly associated with a reach type as “intermittent species” or “perennial species”,

and taxa that were weakly associated with reach type as “nonindicative species”. All analyses

were performed in R version 3.4.1 (R Core Team 2014) using the package ‘vegan’ (Oksanen et

al. 2013).

Species and Age-class Occupancy Dynamics in Intermittent Streams

I used multispecies dynamic occupancy models to assess the effects of flow condition on

metapopulation dynamics of “intermittent species”, “perennial species”, and “nonindicative

species” in intermittent streams of the ICB. Specifically, I used species detections in replicated

samples on multiple dates to estimate fish persistence (the probability that a species that was

present at a site on a given date was still present at that site on the next sampling date) and

colonization (the probability that a species that was absent from a site on a given date was

present on the next sampling date) in relation to changes in flow condition (Figure 2.3), for fishes

characteristic of each stream type (intermittent, perennial, or nonindicative), while accounting for

incomplete species detection (Royle and Marc 2007, MacKenzie et al. 2009, Peterson and Shea

2014). Complete details of the model can be found in Appendix B. I modeled occupancy

dynamics for adults and juveniles separately to evaluate evidence that younger fish had higher

persistence or colonization rates than adults. For each analysis, I included all taxa that occurred

in at least 5% of samples (21 species for adults; 25 species for juveniles, Appendix C). The two

data matrices (one each for adults and juveniles) contained species-specific detections in one or

two reaches at each of the 12 sites for 82 weekly samples spanning June 2015 to January 2017.

Detection data were coded as “NA” for weeks lacking samples at a given site. I fit models with a

22

Bayesian framework implemented with the Markov chain Monte Carlo (MCMC) software JAGS

version 4.3.0 (Plummer 2003), run using the R package “jagsUI” (Kellner 2015), in R version

3.4.1 (R Core Team 2014). I used diffuse priors for parameter coefficients and I assessed

convergence using the Brooks-Gelman-Rubin statistic, R-hat (Brooks and Gelman 1998). I

assessed model fit with a Bayesian p-value based on the discrepancy (Freeman-Tukey statistic)

between the observed and (model-based) expected number of species detected in each survey,

and the same statistic calculated for a replicate data set simulated using persistence, colonization,

and detection estimates at each MCMC iteration (Freeman et al. 2017). A value of less than 0.05

or greater than 0.95 would indicate substantial model lack-of-fit (Schaub and Kéry 2012). I

considered a covariate informative if 95% confidence intervals did not cross zero for adults or

juveniles. The full model code is provided in Appendix D.

Abiotic Effects on Observed Species and Age-class Occurrence in Isolated Pools

To test a priori hypotheses of the effects of changes in abiotic variables on fish

persistence in isolated pools, I measured water quality in pools from June until September in

2015 and 2016. In contrast to the dynamic occupancy model, which used detection or non-

detection at the site level, this analysis used species and age-class detections in individual,

isolated pools, for which I also monitored environmental conditions. I deployed temperature

loggers (HOBO UA-001-08 Pendant Temperature Data Loggers, Onset Computer Corp., Bourne,

Massachusetts) at each pool during the sampling period to measure the maximum temperature

between survey periods at 30-minute intervals. I used dissolved oxygen (DO) levels measured

during each fish survey using a YSI handheld WQM, model 55. To test the influence of total

ammonia, the total amount of ionized (NH4+) and un-ionized (NH3) ammonia in solution

23

(hereafter referred to as ammonia), I collected water samples during periods of isolation from

each pool in 2016 on each fish survey date. Water samples were preserved using phenol and

analyzed within 30 days of collection (Solorzano 1969) using a Lachat QuickChem +8500 Series

2 FIA System flow injection analyzer (Hach Company, Loveland, CO, USA). I also measured

wetted length and width of each pool on each sampling date, at a minimum of three transects,

with depth measured at 25%, 50% and 75% of the wetted width.

To test how these water quality parameters affected the probability that a species

occurred and was detected in an isolated pool (i.e., observed occurrence), I used logistic

regression in general linear mixed models with a repeated nested design. I restricted the analysis

to the 35 species found to occur during intermittency, using adults and juveniles for the same 21

and 25 species, respectively, included in occupancy analyses (Appendix C). For each isolated

pool, I only included detections and non-detections of species that had at least one known

previous occurrence at the site where the pool was located. That is, if a species had never been

detected at a given site, that species was coded as “NA” for all pools at that site. I separately

estimated effects of each covariate (maximum temperature, DO, maximum pool depth, and

maximum measured ammonia level) on the probability of observing adults and (in separate

models) juveniles in isolated pools. I included three random effects on the model intercepts to

account for variation in observed occurrence among species, pools, and repeated surveys (nested

within pools). Additionally, I included a random effect for species on the estimated slope for

each tested covariate. I modeled adults and juveniles separately because I expected that the age

classes would respond differently to environmental conditions. I used binomial regression within

the GLMM framework in the program R version 3.4.1 (R Core Team 2014) with the package

“lme4” (Bates et al. 2014). Prior to analysis, I centered and scaled values around zero by

24

subtracting the mean and dividing by the standard deviation for each covariate (Table 2.1). I

considered parameters as informative if their 95% confidence intervals (CI) did not overlap zero.

Results

Community Assemblage Differences Between Intermittent and Perennial Streams

The differences in community assemblage at intermittent and perennial sites in the ICB

are represented by a convergent three-dimensional ordination using the NMDS analysis (stress

=0.15). PERMANOVA detected significant differences among a priori reach types

(F1,216=56.719, p<0.0001). The stream types associated with NMDS axes 1 and 2, and NMDS

axes 1 and 3 shows intermittent streams and perennial streams occupying different regions of

space, reflecting different, non-overlapping locations of functional centroids in three-

dimensional space (Figure 2.4a-b). Ellipses of credible intervals indicate that perennial streams

have less variation, likely because community composition is similar at all perennial sites and

these are represented by fewer samples, whereas intermittent streams have wide variation in

species occurrence (Figure 2.4c). Of the species included in NMDS analysis, all species were

found in perennial sites, whereas four species were never found at intermittent sites. These four

taxa, Notropis longirostris, Notropis chalybaeus, Ichthyomyzon gagei, and Etheostoma

parvipinne, are all stream dwelling fishes (and not known from lentic environments) present in

5% to 37% of samples at perennial sites. Indicator species analysis showed that perennial reaches

had 23 significant indicator species (“perennial species”), whereas intermittent reaches only had

five significant indicator species (“intermittent species”). The remaining species (“nonindicative

species”) were not significantly associated with stream type (Appendix E). These species-

specific designations were used as species covariates in the dynamic occupancy analysis.

25

Species and Age-class Occupancy Dynamics in Intermittent Streams

Streamflow varied substantially during the study period, with ten of the twelve study sites

experiencing complete drying in the fall of 2015, and eleven experiencing complete drying in the

fall of 2016 (Figure 2.5). During the study, I completed 134 surveys where fish were present and

five surveys in the winter of 2015 where study sites had resumed flow but no fish were captured.

A total of 77 surveys were conducted when study sites were flowing and 66 surveys were

conducted during periods of isolation (i.e., when pools were isolated completely within the study

reach or when a pool within the reach was connected to a larger pool extending downstream or

upstream of the reach boundaries). When fishes were present, observed richness varied from one

to 19 species, with the most species occurring during isolated events. The best models for

dynamic occupancy (Appendix C) fit the observed data with a Bayesian p-value of 0.72 for

adults and 0.65 for juveniles, suggesting model fit was adequate in both cases (Gelman et al.

1996).

Based on the dynamic occupancy models, probability of detection for adults varied from

0.09 for Centrarchus macropterus to 0.82 for Gambusia sp. (Figure 2.6) and for juveniles varied

from 0.08 for L. gulosus to 0.85 for N. harperi (Figure 2.7). Detection increased during

during “isolated” and “isolated-open” events (Figure 2.3) but not significantly (Table 2.2). For

both adults and juveniles, an upstream and a downstream pass during sampling did not have a

significant effect on detection, with covariate effects broadly centered around zero, indicating

that sampling method did not affect detection (Table 2.2). Of the species assessed, twenty

species had both juveniles and adults, with thirteen of twenty juveniles having higher mean

persistence than adults (Figure 2.8, Figure 2.9). During periods of intermittency, there was a

positive effect but not significant on persistence when streams were isolated, but still open to

26

upstream and downstream movement (“isolated-open”, Table 2.2). Colonization rates also varied

among species, with fifteen out of twenty adults having higher colonization rates than juveniles

(Figure 2.10, Figure 2.11). Colonization was negatively (but not significantly) affected for both

adults and juveniles as distance from perennial stream increased (range of 1.8-18.6 km), and

during periods when streams were isolated. Juveniles showed a significantly negative effect of

cool season on colonization (Table 2.2).

Species Persistence in Isolated Pools

Periods of isolation, when a stream was “isolated” or “isolated-open” (Figure 2.3), lasted

up to 13.4 weeks, at which time streams dried or resumed flow. I monitored a total of 26 periods

of isolation over the duration of the study (June 2015-January 2017), with the number of weeks

isolated calculated from the most recent date when a stream had flow. My initial hypotheses for

species differences in persistence in isolated pools, where intermittent species were predicted to

have higher persistence rates than perennial species, were generally supported for both adults and

juveniles. Perennial adults had the lowest rates of persistence as the number of weeks a site was

isolated increased (Figure 2.12), whereas intermittent and nonindicative adults persisted similarly

through weeks of isolation (Table 2.2). Most juveniles had higher rates of persistence in isolated

pools compared to adults (Figure 2.12, Figure 2.13). Intermittent juveniles had the highest rates

of persistence during isolated periods, with mean probability of persistence reaching above 0.9

for all intermittent species within a few weeks of isolation (Figure 2.13). Nonindicative juveniles

had a similar persistence during isolation as intermittent juveniles. Perennial juveniles had, on

average, somewhat higher rates of persistence during isolation than adults, with the effect of

weeks isolated lower than intermittent species (although credible intervals included 0; Table 2.2).

27

Species Colonization After Resumption of Flow

Continuous flow resumed at all study sites between mid-October to December during the

fall of 2015 and lasted from eighteen to fifty weeks. In 2016, flow resumed at all sites in early

December after over two months with no rainfall; my final survey occurred after about five

weeks of continuous flow. Hypotheses for intermittent species colonizing reaches more quickly

were based on the concept that if species can persist in local refugia, whether at study sites or in

pools located near sampled reaches, then those fish would colonize reaches sooner than species

recolonizing from more distant, perennial reaches. In fact, adults (Figure 2.10) but not juveniles

(Figure 2.11) of species indicative of intermittent streams had higher point estimates of

colonization rates than most other species. However, the effect of weeks since resumption of

flow (“weeks flowing”) was not significantly different among species types for adults (Figure

2.14) or juveniles (Figure 2.15). Overall, the mean effect of weeks flowing on colonization was

positive, but credible intervals included 0, representing the possibility that colonization

probability did not depend on how long a site had been flowing (Table 2.2).

Observed Differences Among Species and Age-class Occurrence in Isolated Pools

Of the four variables that I expected to influence the occurrence of fishes in isolated

pools, I only found support for ammonia and maximum depth, for both adults and juveniles. In

the case of adults, elevated levels of ammonia had a significant negative effect (Figure 2.16) and

maximum depth had a significant positive effect on observed occurrence (Table 2.3, Figure

2.17). Juveniles had a significant positive effect of increased maximum depth (Figure 2.18) and

significantly lower probabilities of observed occurrence in relation to elevated levels of ammonia

(Figure 2.19), and dissolved oxygen (Table 2.4, Figure 2.20). Although ammonia and maximum

28

depth were found to be significant for both adults and juveniles, these variables were correlated

(Pearson correlation, r=-0.30), and therefore were not used together in a model. I found no

support for my a priori hypothesis that maximum temperature would have a negative effect on

species occurrence within isolated pools, even though maximum temperature ranged from 15.60

- 32.91°C (Table 2.1).

Random effects on model intercepts showed that, among juveniles, Gambusia sp., A.

sayanus, L. macrochirus, E. zonatum, N. harperi and N. crysoleucas all had higher probabilities

of occurring in isolated pools relative to other species. Probability of observed occurrence was

lower for juvenile E. swaini, E. sucetta, N. texanus and P. grandipinnis across all covariate

models (Figure 2.21, Figure 2.22). Random slopes indicated that juvenile E. swaini and E.

sucetta were the most strongly affected by changes in DO (Figure 2.23). Among adults,

Gambusia sp., E. zonatum, L. macrochirus, L. microlophus, and N. harperi all had higher

probabilities of occurring in isolated pools relative to other species. Additionally, A. sayanus, E.

swaini, and E. americanus had the lowest probability of occurrence in isolated pools (Figure

2.24, Figure 2.25). For maximum depth, adult P. nigrofasticata also showed a lower probability

of occurrence, with L. macrochirus having the most positive effect of increasing depth on

occurrence (Figure 2.24). As ammonia concentration increased and maximum depth decreased,

adult A. sayanus was among three species with lower probability of occurring in isolated pools

(Figure 2.24, Figure 2.25). Conversely, juvenile A. sayanus was among those with the highest

probability of occurrence (Figure 2.21-Figure 2.23).

Discussion

I found that fish community structure differs between intermittent and perennial streams,

with intermittent streams having a subset of species that also occur in perennial streams.

29

Commonly occurring species in intermittent streams of the Ichawaynochaway Creek Basin (ICB)

may occur more frequently because of several differing, non-exclusive processes. Survival

during periods of isolation is one of the drivers of shifts in community composition between

intermittent and perennial streams. Species that are strongly associated with intermittent streams

(intermittent species), and some species that are weakly associated with either stream type

(nonindicative species), generally have higher persistence rates than species strongly associated

with perennial streams (perennial species) when steams become isolated. Juveniles also tend to

have higher persistence rates than their adult counterparts for all species types. This may indicate

that juvenile survivorship of perennial species during isolation may be the driver of later adult

colonization. Intermittent species, as well as some nonindicative species, are also generally the

species with the highest probability of occurring once isolation arises and environmental

conditions become harsh. Colonization rates among species types are similar, with a small effect

size of weeks flowing. This may indicate that species are able to persist in other localized

refugia along stream reaches that experience drying, allowing them to recolonize previously dry

reaches almost immediately. Below, I discuss potential causes of community changes, and

possible implications of increasing intermittency in the ICB.

Increased withdrawals from surface and groundwaters, coupled with climate change, has

altered stream hydrology in southwestern Georgia (Golladay et al. 2016). Shifts in community

composition of stream fishes are common in other systems during low flow events (Walters

2016). My research supports this, as fishes occurring in intermittent streams proved to be a

subset of species found in perennial reaches. A total of twelve rare (<5% of surveys) and four

more common species were never found in intermittent streams, while all species observed in

intermittent streams were also found in perennial streams. The absence of these species in

30

samples from intermittent streams indicates that some species could become scarcer, or be lost

completely, if intermittency becomes more common or widespread. The wide variation in

community structure in intermittent streams, evident in ordination analysis, reflected the changes

in assemblages due to surveys spanning periods of isolation (where a few species commonly

occurred), and long durations of continuous flow (where many species were able to colonize

reaches that had previously dried). Intermittent communities were likely driven by the presence

of the five intermittent indicator species and three nonindicative species (Lepomis macrochirus,

Aphredoderus sayanus and Ameiurus natalis) all of which are found most frequently during

isolated events. Higher occurrence of intermittent species may indicate ecological release during

periods of isolation that does not otherwise occur in perennial streams. Ecological release may be

due to a lack of larger piscivorous fishes, or simply because some species are able to persist

longer during periods of intermittency when other species become extirpated. The three

nonindicative species commonly occur in both perennial and intermittent reaches, as they tend to

have relatively high persistence rates, particularly as juveniles.

Dynamic occupancy modeling allows one to generate projections of population or

community responses to flow sequences that represent hydrologic conditions favoring greater

intermittency in the region. Occupancy studies exploring the effect of low-flows on fish species

in the lower FRB during drought years showed that local extinction was strongly related to short-

term (10-day) low flows (Peterson and Shea 2014). Small-bodied fishes with generalized life-

history characteristics (i.e., high tolerance) were more resilient to flow variability (Peterson and

Shea 2014) than large bodied species with low tolerance to anthropogenic effects including low-

flows (McCargo and Peterson 2010). While I did not specifically test for body size, all

intermittent species were small bodied, and as both juveniles and adults they generally had the

31

highest rates of persistence within isolated pools. Peterson and Shea (2014) and McCargo and

Peterson (2010) found support for generalized traits (i.e., high or low tolerance) in relation to

low-flows used in extinction models, with tolerance based on species trait accounts (Boschung

and Mayden 2004) and index of biotic integrity designations (GADNR 2005). Many tolerance

assignments were accurate within the context of persistence during intermittency, especially

among well-studied species. However, I found that some species classified as low tolerance to

low-flows (e.g., N. harperi) had the highest persistence rates in isolated pools, while some

species considered to have high tolerance (e.g., P. grandipinnis and P. nigrofasciata) were

among those with the lowest persistence rates in isolated pools. Similarly, some perennial

species were found to have high persistence rates during periods of isolation as juveniles, but not

as adults. This implies that for some species existing classifications may not accurately describe

ecological responses, and that actual tolerance to low-flows may vary among life stages.

Species persistence dynamics are controlled by the physiographic characteristics of

isolated refugia. I found that species were more likely to persist in pools that were connected to

upstream and downstream reaches (although the estimated effect was uncertain), similar to a

spatially-explicit model for Brassy minnow (Scheurer et al. 2003). Potentially higher persistence

rates in isolated pools connected to upstream or downstream reaches (“isolated-open”) may

primarily be explained by the influence of maximum depth. Increases in maximum depth are

indicative of increased habitat availability, as individuals are better able to escape adverse

conditions (i.e., predation or decreased water quality). Conversely, I found that the probability of

species occurrence significantly decreased with the loss of water depth in isolated pools.

Many species found in intermittent streams have high persistence rates within isolated

pools. Specifically, adults and juveniles strongly associated with intermittent streams generally

32

have the highest persistence rates. In intermittent tallgrass prairie streams, common species also

possessed lower extinction rates than rarer species, though unlike our results, common species

had higher colonization rates (Whitney et al. 2016). Karst streams across the Dougherty Plain

exchange water with underlying springs, fractures, and porous stream beds (Albertson and Torak

2002, Rugel et al. 2012), often creating multiple deep pools along stream reaches that may act as

refugia during periods of intermittency. The lack of a strong effect on colonization of number of

weeks flowing, and of differences among species types, suggests that isolated refugia between

study reaches and perennial streams may serve as nearby sources of colonizers. Intermittent

species, nonindicative species, and perennial juveniles that were found in intermittent streams are

likely able to survive, even within the harshest of isolated pools. However, those species that

rarely or never occurred in intermittent streams may not be able to survive periods of isolation

during any life history stage.

Davey and Kelly (2007) showed that the rate of recolonization of a reach in intermittent

streams declined strongly with increasing distance to refugia. While I was not able to test

specifically how distances to isolated refugia influenced colonization rates, I found distance to a

perennial stream had a negative effect on colonization rates. This indicates that although some

source populations may exist along streams, colonization rates were also influenced by

populations from perennial reaches. Species more common to perennial reaches may be able to

quickly colonize and persist in now intermittent streams if flow continues for an extended period

of time. If the intensity and duration of stream drying increases, it may become more difficult for

even resilient species to colonize upper stream reaches.

Fish community structure can fluctuate seasonally, with change generally low during

winter due to decreased activity (Schlosser 1991, Peterson and Rabeni 1996) and a need for

33

energy conservation (Schlosser 1991). At low temperatures, fish metabolic rates are reduced, and

fish feeding, movement, and growth are low (Winberg 1960), generally promoting stable

community structure (i.e. population persistence). During spring, high-flow events trigger large-

scale upstream migrations of some adult and larger juvenile fish in warm water streams, as they

move to fulfill life history requirements (Peterson and Rabeni 1996, 2001). The relationship

between cold season and colonization rates for adults and juveniles are likely due to a

combination of these effects. Most of the species in this study are small-bodied individuals that

reproduce around age 1. The negligible effect of cool season on adults indicates that dispersal

likely occurs once flow resumes during fall and winter months, as individuals seek access to

newly inundated reaches. The significant negative effect of cool season on juvenile colonization

is likely due to the fact that spawning for many individuals does not occur until streams begin to

warm.

The lack of evidence for temperature effects on species occurrence in isolated pools is not

surprising for warm water stream fish. It has been well documented that fish living in streams

that experience frequent drying can tolerate temperatures above 34°C (Welcomme 1964,

Matthews and Heins 1987, Smale and Rabeni 1995, Ostrand and Wilde 2001, Matthews 2012).

Temperature variation does not necessarily affect all life stages equally, with early life history

stages (e.g., embryos and larvae), often the most vulnerable because of their sensitivity to

temperature variation (Rombough 1997). While there was no effect of temperature on adults or

juveniles, my sampling method did not include assessing the earliest life stages, when sensitivity

to prolonged elevated temperatures is increased. Further, without flow and continued mixing of

waters, embryos that sink to the bottom of isolated pools are subject to the highest concentration

of ammonia, and lowest concentration of dissolved oxygen (DO).

34

Elevated temperatures and low DO are often coupled. Most species can tolerate short-

term exposure to hypoxia, but only a few are adapted to persist for extended periods under such

conditions (Matthews and Heins 1987, Matthews and Marsh-Matthews 2003). Streams with

seasonal low DO concentrations often contain fewer, more tolerant species relative to streams

with higher DO (Smale and Rabeni 1995). Average DO concentrations in isolated pools were

close to 1.6 mg/L, which can be a lethal level for many stream fishes (Smale and Rabeni 1995).

This is consistent with the significantly positive effect of DO on the occurrence of juvenile

fishes, though there was no significant effect on adult occurrence. Evidence generally suggests

that bigger fishes are better equipped than smaller fishes to tolerate periods of suboptimal oxygen

conditions (Urbina and Glover 2013). This is attributed to larger stores of glycogen available for

anaerobic metabolism, and greater reservoirs for the accumulation of toxic anaerobic end

products (Almeida-Val et al. 2000, Nilsson and Östlund-Nilsson 2004, Everett and Crawford

2009). Analysis indicated that depth and ammonia levels were correlated, possibly indicating that

as crowding increased, ammonia concentrations may have become toxic for some species.

Increased connectivity facilitates the movement of fish from areas with higher temperatures,

greater ammonia concentrations, and lower DO levels, while increased maximum depth allows

larger bodied individuals to avoid predation.

Intermittent fish communities in the ICB are a subset of fishes found in perennial reaches,

with the fishes that are indicators of perennial streams often having the lowest persistence rates. I

found evidence that low levels of DO and high concentrations of ammonia decrease the

probability of fish occurrence in isolated pools, though many species found in intermittent

streams had high survivorship during periods of isolation. In a region where groundwater

strongly influences the baseflows of streams, it is essential to understand the consequences that

35

reduced flows can have on fish assemblages. While persistence rates were high among species

that commonly occurred in intermittent streams, many species of the ICB were never found in

intermittent reaches. My results are similar to many observations that responses to environmental

factors are species-specific. The incorporation of a species-specific rates approach, via dynamic

occupancy, to stream intermittency is relatively novel, and can help advance the mechanistic

understanding of flow-ecology relationship, while also informing environmental flow standards.

Many of the species assessed in this study are found throughout the Coastal Plain. One thing

missing from this analysis is the identification of species characteristics that characterize ability

to maintain populations in intermittent streams. This is explored in the next chapter.

36

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Table 2.1: Summary statistics of water quality data obtained in 90 isolated pools monitored in 12

stream sites in the Ichawaynochaway Creek basin, June through September 2015 and 2016,

followed by their values centered and scaled around zero by subtracting the mean and dividing

by the standard deviation. Scaled values were used as covariate effects on observed fish

occurrence in isolated pools. Numbers of isolated pools (n), and mean covariate value are shown

along with standard deviation (SD), standard error (SE), minimum (Min) and Maximum (Max).

n Mean SD SE Min Max

Dissolved Oxygen (mg/L) 90 1.56 1.12 0.12 0.08 4.81

Maximum Water Temperature (°C) 90 26.10 2.98 0.31 15.60 32.91

Maximum depth (m) 90 0.41 0.25 0.03 0.03 1.58

Maximum ammonia (ug/L) 69 258.54 295.14 35.53 6.19 1640

44

Table 2.2: Effects of covariates on regression coefficients for persistence, colonization, and

detection from multi-taxa, dynamic occupancy models using a time-series (2015-2017) of

detection for adults of 21 species and juveniles of 25 species in the Ichawaynochway Creek

basin. Stream state, sampling method and cool season use binary coding. Distance is the distance

of the study site from the nearest downstream perennial stream, standardized by subtracting the

mean and dividing by the standard deviation. Effects of indicator-species covariates (Intermittent

Nonindicative species and Perennial species, with Intermittent species as the baseline) on

regression coefficients are shown for persistence during the number of weeks a site was isolated

(Weeks Slack) and for colonization after resumption of flow (Weeks Flowing). Variance terms

are for random effects of site and date (“surveys”) on intercepts for persistence, colonization, and

detection, and on species-slopes for relations between persistence and Weeks Slack, and between

colonization and Weeks Flowing. All values are on the logit scale, and show the posterior means

and 95% credible intervals (in parentheses).

Covariate Effect Adults Juveniles

Persistence

Isolated-Open 1.15 (-0.72, 3.64) 1.43 (-0.12, 3.46)

Variance among surveys 0.66 (0.00, 0.99) 0.42 (0.002, 0.97)

Weeks Slack, Intercept 0.54 (-0.11, 1.74) 1.12 (0.30, 2.27)

Weeks Slack, Nonindicative species 0.14 (-1.19, 3.10) 0.08 (-1.23, 1.78)

Weeks Slack, Perennial species -0.89 (-2.05, -0.05) -0.81 (-2.00, 0.26)

Weeks Slack, Variance among species 0.27 (0.00, 0.91) 0.29 (0.001, 0.92)

Colonization

Cool Season -0.13 (-1.39, 1.28) -2.48 (-3.23, -1.77)

Isolated/Isolated-Open -1.95 (-3.15, -0.9) -1.22 (-2.02, -0.50)

Distance to Perennial Reach -0.34 (-0.74, 0.07) -0.30 (-0.62, 0.01)


Weeks Flowing, Intercept 0.05 (-0.06, 0.33) 0.06 (-0.04, 0.20)

Weeks Flowing, Nonindicative species 0.02 (-0.20, 0.31) -0.03 (-0.17, 0.11)

Weeks Flowing, Perennial species -0.02 (-0.25, 0.10) -0.02 (-0.16, 0.11)

Weeks Flowing, Variance among species 0.01 (0.00, 0.04) 0.01 (0.001, 0.03)

Detection

Isolated 0.19 (-0.55, 0.95) -0.03 (-0.55, 0.5)

Isolated Open 0.36 (-0.59, 1.35) 0.00 (-0.81, 0.84)

Survey Method (1 vs 2 passes) -0.38 (-0.93, 0.17) -0.44 (-0.99, 0.12)


45

Table 2.3: Modeled effects of environmental covariates on probability of observed occurrence of adult fishes in 90 isolated stream

pools in the Ichawaynochway Creek basin, 2015-2016. Values are the estimated effects on the log-odds of occurrence (95%

confidence intervals) for predictor variables (values were centered and scaled around zero by subtracting the mean and dividing by the

standard deviation) and the estimated random variance in intercepts attributable to species, surveys, and pools (nested within repeated

survey of a pool), and in slopes attributable to species.

Model Parameter

Maximum temperature Maximum Depth Dissolved Oxygen Maximum Ammonia

Intercept -2.11 (-2.81, -1.21) -2.06 (-2.74, -1.39) -2.08 (-2.95, -1.44) -2.33 (-3.35, -1.37)

Covariate -0.08 (-0.38, 0.23) 0.47 (0.18, 0.81) 0.19 (-0.23, 0.48) -0.54 (-1.28, -0.10)

Random Effects

Surveys (intercept) 0.41 (0.09, 0.96) 0.36 (0.00, 0.91) 0.38 (0.04, 1.03) 0.13 (0.00, 0.82)

Pools (intercept) 1.06 (0.43, 1.40) 0.75 (0.39, 1.19) 1.04 (0.52, 1.33) 1.20 (0.72, 1.50)

Species (intercept) 1.80 (0.80, 1.62) 1.78 (0.87, 1.82) 1.75 (0.90, 1.77) 1.48 (0.90, 2.07)

Species (slope) 0.07 (0.05, 0.43) 0.11 (0.04, 0.54) 0.03 (0.01, 0.38) 0.42 (0.20, 1.27)

46

Table 2.4: Modeled effects of environmental covariates on probability of observed occurrence of juvenile fishes in 90 isolated stream

pools in the Ichawaynochway Creek basin, 2015-2017. Values are the estimated effects on the log-odds of occurrence (95%

confidence intervals) for predictor variables (values were centered and scaled around zero by subtracting the mean and dividing by the

standard deviation) and the estimated random variance in intercepts attributable to species, surveys, and pools (nested within repeated

survey of a pool), and in slopes attributable to species.

Model Parameter Maximum temperature Maximum Depth Dissolved Oxygen Maximum

Ammonia

Intercept -1.30 (-1.94, -0.65) -1.35 (-2.13, -0.48) -1.36 (-2.00, -0.78) -1.27 (-2.06, -0.45)

Covariate -0.05 (-0.34, 0.23) 0.56 (0.34, 0.95) 0.33 (0.06, 0.69) -0.42 (-0.89, -0.13)

Random Effects

Surveys (intercept) 0.61 (0.41, 1.02) 0.38 (0.21, 0.89) 0.55 (0.40, 1.08) 0.49 (0.26, 1.05)

Pools (intercept) 0.89 (0.49, 1.24) 0.75 (0.47, 1.15) 0.95 (0.49, 1.29) 0.74 (0.09, 1.23)

Species (intercept) 1.95 (0.97, 1.82) 2.03 (0.90, 1.92) 1.46 (0.99, 1.91) 1.62 (1.09, 2.08)

Species (slope) 0.03 (-1.00, 1.00) 0.04 (0.01, 0.40) 0.15 (0.04, 0.54) 0.12 (0.10, 0.73)

47

Figure 2.1: Locations of intermittent streams study sites (marked with squares) that were

surveyed to assess shifts in community assemblages, species-specific rates of persistence and

colonization in dynamic occupancy models, and probability of persistence in isolated pools

within the Ichawaynochaway Creek Basin during 2015-2017. Perennial sites (marked with

triangles) indicate streams where published and unpublished data were obtained using similar

survey methods, and were used to assess differences in community assemblages between

intermittent and perennial streams.

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Figure 2.2: Discharge, water temperature, and air temperature at Spring Creek near Leary, GA (USGS gage 02354475). Periods where

discharge is at or near zero represent timing of intermittency, during which isolation or complete drying occurred.

49

Figure 2.3: Changes in stream state used as covariates to estimate persistence and

colonization in intermittent streams, where “flowing” represents stream state where

discharge is >0, “isolated” represents a pool that is isolated from upstream or downstream

movement of fishes (e.g., a small pool), and “isolated-open” represents an isolated pool

that is open to upstream or downstream movement of fishes (e.g., a big pool).

50

(a)

(b)

51

(c)

Figure 2.4(a-c): Non-metric multi-dimensional scaling (NMDS) ordination of stream

samples based on Brays-Curtis dissimilarities in species occurrences. Ellipses represent

centroids and 95% confidence intervals for mean scores for samples from perennial and

intermittent streams. Each graphic represents 2 of the 3 dimensions in two-dimensional

space.

52

Figure 2.5: Time series of changes in stream state for 12 intermittent study sites in the Ichawaynochaway Creek Basin, June 2015 to

January of 2017.

53


intervals for adults of species found in >5% of surveys averaged over 12 study sites in the

Ichawaynochaway Creek Basin. Values plotted are estimates for each of the 21 species

using a multi-taxa, dynamic occupancy model. Taxa are identified by the first three

letters of their genus and species.

54


intervals for juveniles of species found in >5% of surveys averaged over 12 study sites in

the Ichawaynochaway Creek Basin. Values plotted are estimates for each of the 25

species using a multi-taxa, dynamic occupancy model. Taxa are identified by the first

three letters of their genus and species.

55


intervals for adults of species found in >5% of surveys averaged over 12 study sites in the

Ichawaynochaway Creek Basin. Values plotted are estimates for each of the 21 species

using a multi-taxa, dynamic occupancy model. Taxa are identified by the first three

letters of their genus and species.

56


intervals for juveniles of species found in >5% of surveys averaged over 12 study sites in

the Ichawaynochaway Creek Basin. Values plotted are estimates for each of the 25

species using a multi-taxa, dynamic occupancy model. Taxa are identified by the first

three letters of their genus and species.

57


confidence intervals for adults of species found in >5% of surveys averaged over 12

study sites in the Ichawaynochaway Creek Basin. Values plotted are estimates for each of

the 21 species using a multi-taxa, dynamic occupancy model. Taxa are identified by the

first three letters of their genus and species.

58


confidence intervals for juveniles of species found in >5% of surveys averaged over 12

study sites in the Ichawaynochaway Creek Basin. Values plotted are estimates for each of

the 25 species using a multi-taxa, dynamic occupancy model. Taxa are identified by the

first three letters of their genus and species.

59

Figure 2.12: Average mean of probability of persistence for adult fish in isolated pools, plotted in relation to duration of pool isolation.

Probabilities are plotted for 21 species estimated using a multi-taxa, dynamic occupancy model applied to 26 periods of continuous

isolation at 12 study sites in the Ichawaynochaway Creek Basin. Black lines indicate the species-specific means of persistence and red

lines indicate the means for each of the three species types.

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Figure 2.13: Average mean of probability of persistence for juvenile fish in isolated pools, plotted in relation to duration of pool

isolation. Probabilities are plotted for 25 species estimated using a multi-taxa, dynamic occupancy model applied to 26 periods of

continuous isolation at 12 study sites in the Ichawaynochaway Creek Basin. Black lines indicate the species-specific means of

persistence and red lines indicate the means for each of the three species types.

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Figure 2.14: Average mean of probability of colonization for adult fish, plotted in relation to duration of flow since isolation or

complete drying. Probabilities are plotted for 21 species estimated using a multi-taxa, dynamic occupancy model applied to 26 periods

of continuous isolation at 12 study sites in the Ichawaynochaway Creek Basin. Black lines indicate the species-specific means of


62

Figure 2.15: Average mean of probability of colonization for juvenile fish, plotted in relation to duration of flow since isolation or

complete drying. Probabilities are plotted for 25 species estimated using a multi-taxa, dynamic occupancy model applied to 26 periods

of continuous isolation at 12 study sites in the Ichawaynochaway Creek Basin. Black lines indicate the species-specific means of


63

Figure 2.16: Modeled probability of observed occurrence of adults in relation to maximum total

ammonia (ug/L) in 90 isolated pools samples in the Ichawaynochaway Creek Basin, 2015-2016.

Plot shows mean and 95% confidence intervals.

64

Figure 2.17: Modeled observed occurrence of adults in relation to maximum depth (m) in 90

isolated pools samples in the Ichawaynochaway Creek Basin, 2015-2016. Plot shows mean and

95% confidence intervals.

65

Figure 2.18: Modeled observed occurrence of juveniles in relation to maximum depth (m) in 90

isolated pools samples in the Ichawaynochaway Creek Basin, 2015-2016. Plot shows mean and

95% confidence intervals.

66

Figure 2.19: Modeled observed occurrence of juveniles in relation to maximum ammonia (u/gL)

in 90 isolated pools samples in the Ichawaynochaway Creek Basin, 2015-2016. Plot shows mean

and 95% confidence intervals.

67

Figure 2.20: Modeled observed occurrence of juveniles in relation to dissolved oxygen (mg/L) in

90 isolated pools samples in the Ichawaynochaway Creek Basin, 2015-2016. Plot shows mean

and 95% confidence intervals.

68

Figure 2.21: Species-specific random effects on the intercept and slope of modeled observed

occurrence of juveniles in relation to maximum depth in 90 isolated pools samples in the

Ichawaynochaway Creek Basin, 2015-2016. Plots show means and 95% confidence intervals.

69


occurrence of juveniles in relation to maximum ammonia in 90 isolated pools samples in the


70

Figure 2.23: Species-specific random effects on the intercept and slope of a modeled observed

occurrence of juveniles in relation to dissolved oxygen in 90 isolated pools samples in the


71


occurrence of adults in relation to maximum depth in 90 isolated pools samples in the


72


occurrence of adults in relation to maximum ammonia in 90 isolated pools samples in the


1Davis, J. L., M. C. Freeman, S. W. Golladay. To be submitted to Freshwater Biology

CHAPTER 3

IDENTIFYING LIFE HISTORY TRAITS THAT PROMOTE FISH SPECIES PERSISTENCE

IN INTERMITTENT STREAMS1

73

74

Abstract

Life history traits of stream fishes partly reflect adaptations to natural flow regimes,

which in turn shape assemblage composition via environmental filtering on species persistence.

Thus, trait-based approaches, including the trilateral life history model, have been useful for

understanding species responses to streamflow alteration. In this study, I focused on life history

traits of four cyprinid species in a Coastal Plain stream system of southwestern GA that is

shifting from historically perennial streamflow to intermittency. Native fishes, including these

four species, vary in occurrence, and tolerance to intermittency. I evaluated differences among

the four cyprinids in reproductive timing (based on ovary and oocyte development), sex ratio,

body size at maturity, and reproductive investment (gonadosomatic index (GSI), gonad weight

and egg diameter), traits hypothesized to influence the ability of species to persist in intermittent

streams. I periodically sampled individuals in 14 streams over the duration of a year (May 2016-

April 2017). I found that for Notropis harperi, a species with high persistence rates, reproductive

timing did not overlap with typical seasonal stream drying. N. harperi also had the significantly

smallest minimum length at maturation, greatest GSI and gonad weight, and a tendency towards

larger average egg diameter. Species with low persistence rates in isolated pools (Notropis

petersoni, Notropis texanus, and Pteronotropis grandipinnis), had at least a portion of their

reproductive timing overlapping with times when streams were likely to dry, and had

significantly lower GSI and relative gonad weight than N. harperi. All four species would be

considered opportunistic, rather than periodic or equilibrium, strategists. Our results suggest

however, that some life history traits used to define the trilateral life history model may be useful

for understanding differences in how even closely related species respond to changing

75

environments, with smaller body size at maturity along with appropriate reproductive timing

promoting greater persistence given more frequent and intense periods of drying.

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Introduction

Streamflow defines the physical template of river ecosystems (Poff 1997) and acts as a

selective force and an ecological filter for survival strategies of aquatic organisms (Townsend

and Hildrew 1994, Lytle and Poff 2004). It shapes the distribution and character of riverine

habitats and, in turn, the distribution and abundance of lotic organisms (Power et al. 1995, Bunn

and Arthington 2002). Species adaptations to flow regime occur as a response to the interaction

between frequency, magnitude, and predictability of mortality-causing events (Lytle and Poff

2004). Streamflow in lotic systems has been altered by humans for many reasons, including

extraction for water supply, impoundments for flood control and hydropower, and to support

irrigated agriculture. Human freshwater needs and actions have altered historical hydrologic

regimes, reducing effectiveness of some biotic adaptations, and decreasing stream suitability for

native fauna (Pringle et al. 2000).

Increasingly, species traits are being used to study flow-ecology relationships across

diverse species-assemblages and broad geographic scales (Poff et al. 2006, Frimpong and

Angermeier 2010). Life-history theory predicts that the magnitude, frequency, and predictability

of hydrologic events, such as floods or droughts, can affect evolutionary processes (Iwasa and

Levin 1995, Lytle and Poff 2004). Convergence in the suites of traits characterizing dominant

species along hydrologic gradients has been demonstrated for freshwater fishes (Lamouroux et

al. 2002, Logez et al. 2010), and studies testing predictions from life-history theory support these

relationships (Tedesco et al. 2008, Kennard et al. 2010, Carlisle et al. 2011). This developing

body of work has provided insights into environmental influences on community assemblages in

fresh waters (Poff and Allan 1995, McManamay and Frimpong 2015) and provides useful

77

frameworks for developing flow-ecology hypotheses and environmental flow standards

(McManamay et al. 2015).

Environmental variability is a natural part of aquatic ecosystems and influences the

structure of aquatic communities (Resh et al. 1988, Poff and Allan 1995). From an evolutionary

perspective, floods and droughts that are relatively predictable in their frequency, duration, and

intensity can exert selective pressures that filter certain life history traits, while on ecological

time scales, flow regime shapes assemblage composition by altering population numbers and

species persistence (Poff 1997, Naiman et al. 2008). In systems with a high degree of flow

variability (i.e., hydrologically ‘flashy’) including intermittence, fish assemblages are controlled

by abiotic factors (Echelle et al. 1972, Taylor 1997, Matthews and Marsh-Matthews 2003). As a

result, the predominant fish species in these systems are especially tolerant of variable

environmental conditions ( Winemiller 1989, Fausch and Bramblett 1991).

While drought and stream intermittency are normal processes, low-flow and no-flow

events are increasing in frequency in many areas due to anthropogenic alterations of streamflow

regimes through dams, water diversion, and climate shifts (Brown et al. 2013). The ecological

effects of low-flows are expected to accumulate with increasing frequency and duration (Lake

2003). Evidence suggests that climate-driven streamflow intermittence has increased in the

southeastern US (Palmer et al. 2008b, Falke et al. 2011), including the Coastal Plain of Georgia,

where trends in declining seasonal flows are projected to continue (Larned et al. 2010, Golladay

and Hicks 2013). Natural resource managers face the challenge of understanding the effects of

both water withdrawals, and projected increases in intermittency, when working towards

conserving biological integrity of freshwater systems. Management can be improved through

models of fish responses to low-flow or isolated events that account for life history traits.

78

Life history theory proposes that populations are regulated by trade-offs along

demographic axes of age at maturity, juvenile survival, and fecundity (Stearns 1977). The

trilateral life history model proposed by Winemiller and Rose (1992) identifies three life history

strategies for fishes using tradeoffs among basic demographic parameters. The endpoints of the

trilateral life history model represent strategies that are optimal under certain environmental

conditions (Winemiller and Rose 1992, Winemiller 2005). Opportunistic Strategists (OS) are

predicted to be associated with habitats defined by frequent and intense disturbances. OS have

short generation time, high fecundity, and low juvenile survivorship. Periodic Strategists (PS) are

favored under predictable yet seasonally fluctuating environments. PS maximize fecundity via

delayed reproduction and have larger maturation size. Equilibrium Strategists (ES) are favored

under stable environmental conditions. ES maximize juvenile survival through low fecundity per

spawning event, have larger eggs, and prolonged parental care. Studies have found utility in fish

life-history trait ordination along these three axes across North America (Kennard et al. 2010,

Mims et al. 2010, Mims and Olden 2013, Perkin et al. 2017) for predicting responses to natural

and altered flow regimes on fish assemblages. For example, a study of two cyprinid and one

percid species shows PS traits (high fecundity) favored at sites with greater flow seasonality and

low variability, while ES traits (large eggs) were prevalent in stable flow conditions (Bennett et

al. 2016). Nevertheless, intraspecific trait variation may contradict the trilateral life history

model (Bennett et al. 2016). Additionally, species placement within the trilateral life history

model is largely dependent on the species assessed within an assemblage.

This study focused on four commonly occurring, yet relatively understudied cyprinid

species, Pteronotropis grandipinnis, Notropis harperi, Notropis petersoni, and Notropis texanus,

in a southeastern Coastal Plain system. I previously found that mature individuals of these

79

cyprinid taxa vary substantially in rates of persistence during periods of flow intermittence, when

fishes are restricted to isolated pools, yet the underlying life history mechanisms remain

unknown (Chapter two of this thesis, Figure 3.1). Regional streamflows in my study area are

strongly influenced by agricultural water withdrawals and climate variability, resulting in

increases in the duration and intensity of intermittency. The primary objective of this study was

to explore life history strategies that enable cyprinid species to persist within intermittent streams

and to determine if those life history traits actually coincide with high persistence rates. I

explored differences in reproductive timing, sex ratio, reproductive size, reproductive investment

(via gonadosomatic index (GSI), gonad weight, egg diameter), and food habits during flowing

and intermittent periods. A secondary objective was to determine if predictions of the trilateral

life history model apply to morphologically similar species belonging to a single family. I

hypothesized that (i) species with low persistence probabilities reproduce during summer months

(e.g., when periods of stream drying to isolated pools occur); (ii) species with high persistence

probabilities have increased reproductive investment as assessed via GSI, gonad weight, and egg

diameter; (iii) higher probabilities of persistence (N. harperi) coincides with OS, low

probabilities of persistence (N. texanus and P. grandipinnis) coincides with ES, and intermediate

probabilities of persistence (N. petersoni) coincides with PS; and (iv) species with higher

persistence will have a greater shift in diet items between flowing and intermittent periods,

reflecting intentional feeding patterns while streams flow and more opportunistic feeding

patterns during intermittence.

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Methods

Study sites

I collected and used data on four cyprinid species found in the Ichawaynochaway Creek

basin (ICB), a major tributary of the lower Flint River Basin (FRB), southwestern GA, to study

life history patterns in intermittent streams (Figure 3.2). The channels of major tributary streams

within the lower Flint River and the Ichawaynochaway Creek are incised into the Upper

Floridian aquifer and tend to be perennial. Small streams in the northwestern portion of the ICB

are in the Fall Line Hills physiographic district, and also tend to be perennial. Small streams in

the remainder of the basin, within the Dougherty Plain physiographic district, have channels

perched above the aquifer, and tend to have periods of intermittence The study area has low

topographic relief, and porous, sandy soils, which results in low stream drainage density. During

typical winters streamflow increases in response to extended storms (Hicks et al. 1987, Albanese

et al. 2007), and lower temperature and evapotranspiration rates (Torak and Painter 2006).

Rainfall is evenly distributed throughout the year, but during the summer most precipitation is

lost through evapotranspiration, causing water table declines as groundwater recharge is

minimal. This results in riparian areas drying and streams decreasing to seasonal low-flows

(Golladay and Battle 2001) or periods of intermittency.

The FRB has experienced an increased demand on water resources resulting from

population expansion in the upper basin, and irrigation expansion in the lower basin (Golladay

and Hicks 2013). Over the last four decades, the FRB has experienced warming temperatures,

more frequent growing season and multiyear droughts, and increased water withdrawal from

groundwater and surface waters for agriculture. As a result, some streams are shifting from

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perennial to intermittent. Streams crossing the Dougherty Plain in the southern portion of the

ICB are increasingly likely to dry during periods of low rainfall and high groundwater use,

during which the upper Floridan aquifer levels drop below stream channels (Opsahl et al. 2007).

Flow in the smaller streams of ICB is typically lowest (and most likely to cease) during summer

and early fall, usually June through October, then recovers from November through May. The

shift from historically perennial to intermittency in the ICB, combined with the predictable

timing of intermittency, provides a useful framework for assessing fish life history patterns.

Survey and Collection Methods

To study cyprinid life-history patterns, I collected individuals of the four target species

from thirteen study sites in the ICB over a period of one year (May 2016-April 2017). Initial

surveys included twelve sites, all experiencing periods of flow cessation to isolated pools

(“isolation”) or complete drying, with some target species becoming locally extirpated when

isolation occurred. Mills Creek was the only site that maintained inundated habitat and was thus

sampled for the duration of the study period. One perennial stream, Brantley Creek, within the

upper ICB in the Fall Line Hills district, was sampled from October of 2016 to April of 2017 to

continue collection after target individuals became locally extirpated at the original twelve

survey sites within the Dougherty Plain (Figure 3.2).

Target species were collected using a combination of backpack electrofishing and seining

(2.4 m X 1.8 m; 3 mm mesh) at study sites on a three week to monthly basis, with some targets

not found on each date. Streams were surveyed to collect ten individuals of four cyprinid species,

Pteronotropis grandipinnis, Notropis harperi, Notropis petersoni, and Notropis texanus. When

streams were flowing, surveys comprised multiple seine-sets in a minimum of a 50-meter reach,

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where two persons held the seine in flowing water, with the lead-line on the substrate, while one

person disturbed water and bed sediment while backpack electrofishing. In isolated pools, I

collected with only a seine using multiple passes until no additional target species were found.

On every survey date, fish were identified to species, counted, and up to 50 individuals of each

species were measured for standard length to the nearest millimeter. A total of 2725 individuals

were assessed for length at the thirteen survey sites, but 177 were too small to confidently

identify to species. A target of ten individuals were collected during each survey and euthanized

using an overdose (100 mg/L) of buffered MS-222, immediately preserved in 10% formalin, and

transported back to the laboratory for dissection (Heins and Baker 1999). Individuals occurring

at seven study sites were used for reproductive assessment, and individuals occurring at four

study sites were used for diet assessment.

Laboratory Methods and Gonad Assessment

Prior to dissection, all specimens were measured for standard length (SL) with digital

calipers to the nearest 0.01mm, blotted dry, and weighed to the nearest 0.0001g. Specimens were

then cut longitudinally and their gonadal tissue and GI tract removed (esophagus to anus) with

the aid of a stereo microscope (Olympus SZX7). Gonads were blotted dry and weighed to the

nearest 0.0001g using an analytical balance (Mettler AE240). Sex was determined at the time of

excision. Individuals <25 mm SL were classified as immature and sex was not determined.

Males were categorized as either latent (testes were small, thin, and transparent or translucent) or

mature (testes were long, highly vascularized, opaque, smooth, and visible along the length of

the body cavity).

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Ovaries were classified according to staging methods of Heins (1986), which categorizes

oocytes into maturity stages based on oocyte size, coloration, yolk condition, and physical

location within the ovum (Heins and Baker 1993). Females were classified into one of six stages:

latent (LA), early maturing (EM), late maturing (LM), mature (MA), mature-ripening (MR), or

ripe (RE). Female classifications were part of the seasonal clutch production cycle, and were

based on coloration, size, and position of oocytes and mature ova in ovaries, following published

literature for the closest related taxa (Table 3.1).

Statistical analyses

Several life history variables were measured and evaluated to test differences among

species for reproductive timing, sex ratio, reproductive size, and reproductive investment via

GSI, gonad weight, and egg diameter. Boundaries of the reproductive season were determined by

the presence of the first and last reproductively mature males and clutch-bearing females, along

with presence of individuals less than 15 mm SL (Hughey et al. 2012) across collections made

between May of 2016 and April of 2017 . Species-specific trends in reproductive timing were

tested with ANCOVA, followed by post-hoc Tukey adjustment for multiple comparisons of

slopes among species. The response variable was reproductive state, either mature (MA, MR,

and RE for females) or latent (LA, EM, LM for females), coded as binary, with the predictor

variable as an interaction of species and the natural log of “day” in a logistic regression. For the

predictor variable “day”, Julian dates were used with January 1st as “day 1”.

For each species standard length of the smallest mature male and smallest clutch-bearing

female was considered the minimum size of reproductive maturity (Table 3.2). Observed sex

ratio compared to an expected ratio of 1:1 was examined using chi-square tests applied

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separately to adults and immature individuals (above or below minimum length of fecundity) for

each species. Average reproductive size was calculated from the arithmetic mean of standard

length for all MA, MR, and RE females and for all mature males of each species. Pair-wise

comparisons using Tukey-adjusted least square (LS) means following an ANOVA were used to

test differences among species in size of mature individuals. I used simple linear regression of

the natural log of weight and length to determine growth rates among species. Standard length

was used in ANCOVA followed by pair-wise comparison using Tukey-adjusted LS means to

determine if body mass (total and eviscerated mass) was related to standard length, and if mass

varied among species and genders.

Reproductive investment was examined using gonad weight and gonadosomatic index

values (GSI) for males and females:

GSI = (Gonadal mass (g) / (gross body mass (g) – gonadal mass (g))) * 100 (1)

Gonad weight and GSI in relation to SL were compared among species using ANCOVA,

followed by a pair-wise comparison using Tukey-adjusted LS means. To further examine

reproductive investment variation among species, a sample of 20 eggs from each mature,

ripening, and ripe female were measured for egg diameter to the nearest 0.001mm using a stereo

microscope. Each egg had its length and width measured, with its mean used as a response

variable. Egg diameters were compared using a nested ANOVA, and their relation to SL using a

nested ANCOVA, with a random effect for eggs collected from the same individual. Standard

length was included as a covariate for analyses of reproductive investment (egg diameter, gonad

weight, and GSI) because reproductive investment has been found to be positively related to

female body size (Smith and Fretwell 1974). For each model listed above, residuals were

checked for normality by visual inspection of residual plots. Analysis was performed in R

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version 3.4.1 (R Core Team 2014) using the packages “nlme” (Pinheiro et al. 2014) “car” (Fox et

al. 2017) and “lsmeans” (Lenth 2016).

Life History Strategies and the Trilateral Life History Model

Life history variables were evaluated according to strategies defined in the trilateral life

history model of Winemiller & Rose (1992). I used an existing life history trait database (Mims

et al. 2010), along with data obtained from this study, to classify fishes commonly found in

intermittent and perennial streams within the ICB (Appendix A) according to their life history

strategy. Previous analyses (Chapter 2) identified species that could be used as indicators for

intermittent streams (intermittent species), perennial streams (perennial species), and species that

were not strongly correlated with stream type (nonindicative species) (Appendix E). For each

species, I “soft” classified species according to the closest (least distance) relative affinity to

each strategy (Kennard et al. 2010) by calculating Euclidian distance in trivariate life history

space between species position and a single life history strategy endpoint. OS have minimum

fecundity, minimum length at maturation, and minimum relative investment per offspring. PS

have maximum fecundity, maximum length at maturation, and minimum relative investment per

offspring. ES have median fecundity, maximum length at maturation, and maximum relative

investment per offspring. Fecundity was defined as the number of eggs per spawning event.

Length at maturation was defined as the minimum standard length (mm) at maturity for a

species. Egg size was defined as the mean egg diameter (mm) (Mims et al. 2010), and parental

care was determined following Winemiller (1989). I used the statistical mean of life history traits

for species that were combined for analysis (e.g., Gambusia affinnis and Gambusia holbrooki). I

normalized values between 0 and 1 obtained from: ln(fecundity), ln(length at maturation) and

ln(average egg size + 1) + ln (parental care), where a 1 represented the maximum observed value

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for the species assessed. Euclidian distances to each endpoint strategy were calculated and then

normalized between 0 and 1, with the inverse of these values used to provide the affinity to

opportunistic, periodic, or equilibrium life history strategies (Mims and Olden 2013). I then

separately analyzed the study species using length at maturation, egg size, fecundity, and

parental care obtained from Mims et al. (2010) . Euclidian distances were calculated using

distance matrix computation in R version 3.4.1 (R Core Team 2014).

Diet Assessment

To examine diets and test for possible shifts between periods of flowing and isolation, I

examined gut contents of the two target species for which I was able to collect sufficient

individuals from isolated pools (P. grandipinnis and N. harperi). Specimens were cut

longitudinally and their gastrointestinal (GI) tract removed with the aid of a stereo microscope

(Olympus SZX7) and stored in 10% formalin until contents were examined. The range of diet

items used by target species were assessed using the entire GI tract (esophagus to anus). I

quantified diets of consumer species by calculating the percent Relative Importance (%IRI) for

each prey category as follows:

%𝐼𝑅𝐼 =100𝐼𝑅𝐼𝑖

∑ 𝐼𝑅𝐼𝑖′𝑛

𝑖=1

(2)

𝐼𝑅𝐼 = %𝐹𝑂𝑖(%𝑉𝑖 + %𝑁𝑖) (3)

%𝐹𝑂 =100𝐹𝑂𝑖

∑ 𝐹𝑂𝑖′𝑛

𝑖=1

(4)

%𝑉 =100𝑉𝑖

∑ 𝑉𝑖′𝑛

𝑖=1

(5)

%𝑁 =100𝑁𝑖

∑ 𝑁𝑖′𝑛

𝑖=1

(6)

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where n is the number of diet components identified for a sub-population (individuals per survey

assessed), Vi and Ni are the volume and number of individuals of prey i in a sub-population,

respectively, FOi is the number of stomachs containing prey i in a sub-population divided by the

total number of individuals sampled in a sub-population. Volume was estimated by a points

method (Pinkas et al. 1971, Grover 1998) in which a diet component with the highest volume

was given 16 points, and every other component is given 16, 8, 4, 2, 1, and 0 depending on the

volume relative to the component with the highest volume. Percent volume for an individual is

calculated as:

%𝑣 =Number of points allovated to component v

Total points allocated to subsample∗ 100 (7)

I assessed shifts in diet by performing a multivariate ordination of individuals based on

gut contents expressed as %IRI from a subset of 19 surveys, with the number of individuals with

items in their GI tract ranging from one to nine per subset. I performed Non-Metric Multi-

Dimensional Scaling (NMDS) with Brays Curtis dissimilarity measures on diet data. NMDS was

performed with six and descending to two dimensions using a random starting configuration and

convergence determined through Procrustes analysis. Stress was calculated for each convergent

solution and the lowest number of axes with the final stress of less than 0.2 was considered

ecologically interpretable (Clarke 1993). Indicator species analysis was then performed to

identify diet components strongly associated with stream state (isolated or flowing) and species

(N. harperi or P. grandipinnis). All analysis were performed in R version 3.4.1 (R Core Team

2014) using the packages ‘vegan’ (Oksanen et al. 2013) and ‘indicspecies’ (De Cáceres 2010).

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Results

Reproductive Timing and Streamflow

The target species displayed evidence of reproductive activity that extended over 5 to 9

months. Presence of individuals <15 mm SL, mature males, and MA, MR, or RE females

indicated a conservative estimate of reproductive timing of March through November for

Pteronotropis grandipinnis (Figure 3.3, Figure 3.4), January through June for Notropis harperi

(Figure 3.5, Figure 3.6), April through September for Notropis petersoni (Figure 3.7, Figure 3.8),

and March through July for Notropis texanus (Figure 3.9, Figure 3.10). Low flows and seasonal

cessation of flow were seen in early summer and early fall (July through October), with flow

resumption and high flows seen in winter (November through May, Figure 3.11). I used this to

provide a framework in which to link optimal reproductive timing for lotic fishes in these

streams.

Comparison of reproductive timing based on ANCOVA of occurrence of mature

individuals in relation to Julian date showed no significant difference in slopes between N.

harperi and N. texanus, or between P. grandipinnis and N. petersoni, yet the pairs were

significantly different from one another (Figure 3.12). Simple probability curves show presence

of mature N. harperi earliest, beginning in January, but the steeper negative slope indicates a

shorter spawning period than any other species (Figure 3.13). The absence of small individuals

(<15 mm SL) in collections after June (Figure 3.5) indicates that spawning likely ceased before

periods of pool isolation began. Results were similar for N. texanus, indicating an earlier

spawning period. Because I did not capture individuals in August and September (Figure 3.9), I

cannot conclude with certainty that spawning did not continue through months when periods of

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isolation typically occur. The less steep slopes for P. grandipinnis and N. petersoni indicate a

longer period of spawning (Figure 3.13). The presence of small P. grandipinnis from September

through November of 2016 indicates that while reproductively mature individuals were not

captured, spawning likely occurred in spring and fall. P. grandipinnis appears to have the longest

spawning duration of the four species, spanning typical periods of low and high flows (Figure

3.3). N. petersoni was the only species we found with reproductively mature females in August

(Figure 3.8), and no small individuals found during high flow periods, indicating that spawning

likely occurred during summer when flows are typically low and streams are prone to drying

(Figure 3.7).

Sex Ratio and Body Size of Individuals

Overall sex ratio did not differ significantly from the expected 1:1 ratio in sexually

mature females and males, however the ratio of 74% females to 26% males for non-reproducing

N. harperi, and 61% females to 29% males for non-reproducing P. grandipinnis was

significantly different from 1:1 (Table 3.3). The smallest reproductively mature females and

males were, respectively: P. grandipinnis,34.85 and 39.32 mm SL; N. harperi, 38.26 and 32.35

mm SL; N. petersoni , 46.84 and 49.2 mm SL; and N. texanus, 49.47 and 49.26 mm SL (Table

3.4). Pair-wise comparison of LS means for mature individuals were not significantly different

between N. texanus and N. petersoni, but both were significantly different than N. harperi and P.

grandipinnis. N. harperi had the smallest size at maturation and N. texanus the largest (Figure

3.14). Size of identifiable individuals assessed for length at the thirteen sites surveyed ranged

from: 13 to 68 mm SL (39.1 + 0.52; mean + SE) for P. grandipinnis; 9 to 55mm SL (26.49 +

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0.16) for N. harperi; 13 to 70 mm SL (40.06 + 1.05) for N. petersoni; 21 to 70 mm SL (55+ .93)

for N. texanus; and 9 to 21 mm SL (12.74 + 0.17) for unidentified Notropis sp. (Table 3.5).

Taxa did not significantly differ in their relationship between length and mass (Figure

3.15). After controlling for differences in standard length, eviscerated mass for males compared

to females were similar within a species; P. grandipinnis (F1,149=3.00, p=0.09), N. harperi

(F1,117= 0.06, p=0.44), N. petersoni (F1,85= 0.02, p=0.89), N. texanus (F1,80= 1.14, p=0.29). Pair-

wise comparison of eviscerated mass, controlling for standard length, among mature females of

the four species was not significant (F3, 93=1.0791, p=0.36), but eviscerated mass among mature

males was significant (F3, 93= 2.93, p=0.04). Males of P. grandipinnis had significantly greater

mass, adjusted for standard length, than males of N. petersoni; N. harperi and N. texanus were

intermediate. (Figure 3.16). The mean length of mature male N. texanus was 34% larger than the

mean length N. harperi, and female N. texanus were 49% larger than the mean length of N.

harperi (Table 3.4).

Reproductive Investment

Among mature females and males, standard length was not significantly related to GSI

(F93,1=1.7, p=0.19, F74,1=3.08, p=0.08). Both mature females and males of N. harperi had GSI

significantly greater than other species, with no differences among the LS means for P.

grandipinnis, N. petersoni, and N. texanus (Figure 3.17, Figure 3.18). Standard length was

positively and significantly related to total gonad mass for mature females (F=1,93= 42.86,

p=3.18E-19) and mature males (F=1,74, p=1.25E-4) for all taxa. Like GSI, gonad mass was

significantly greater for N. harperi than all other species, regardless of whether length was

adjusted for or not, with no significant differences among the LS means for the others (Figure

3.19, Figure 3.20).

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N. harperi and N. texanus had larger eggs than N. petersoni and P. grandipinnis (F3, 95=

12.356, p <0.001). The diameter of N. harperi and N. texanus showed no significant difference,

nor was there a significant difference in the diameter between N. petersoni and P. grandipinnis

(Figure 3.21). When female size was included as a covariate, egg diameter appeared related to

female size, but not significantly (F3, 94, p=0.197). After controlling for female size, N. harperi

had the largest egg diameter, but was not significantly different than N. texanus. N. petersoni had

the smallest diameter, and was not significantly different than P. grandipinnis (Figure 3.22).

When not accounting for standard length, there was a significant difference between N. texanus

and P. grandipinnis, but while accounting for it there was not (Table 3.6).

Fishes of the Ichawaynochaway Creek Basin and the Trilateral Life History Model

When assessing the target species, N. harperi, the species with the highest persistence

rate during isolation, was considered an OS; N. texanus, the species with the lowest persistence

rate, was an ES; and N. petersoni, the species with intermediate persistence was a PS. I predicted

P. grandipinnis, the species with the second lowest persistence rate to also be an ES, but soft

classification categorized it as a PS (Figure 3.25, Table 3.7). When assessed on an assemblage

level, the target species were classified as OS. Classifications for species shifts are due to

strategist endpoints being fully dependent on what the highest and lowest values are for each

parameter assessed. Life history trait classifications for fishes identified in the ICB included a

continuum of soft classifications in which most species were nearest to the OS endpoint. This

included 19 OS, 11 PS, 6 ES, and three species that were removed from this analysis (Figure

3.23). All species identified as indicators of intermittent streams were classified as OS (60%) or

PS (40%). Species identified as indicators of perennial streams were dominated by OS (65%),

followed by PS (20%) and ES (15%). Species weakly associated with stream type were PS

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(45.5%), ES (27.25%) and OS (27.25%). This indicates that there is no discernable trend in

strategist endpoints and species indicative of stream types (Appendix E).

Diet of Two Fishes in Intermittent Streams

A total of 23 categories of prey taxa were identified in diets of P. grandipinnis and N.

harperi after grouping prey items to family or the lowest possible taxonomic category (Figure

3.25). Percent index of relative importance (Barnett et al.) is based on grouping individuals of a

species collected during a single survey. Groups consisted of one to nine individuals with food

items in their GI tracts, with empty GI tracts not used for analysis. Diets of P. grandipinnis and

N. harperi differed during isolated events. Differences may be due to P. grandipinnis primarily

feeding on terrestrial items, whereas N. harperi primarily fed primarily on aquatic items (Figure

3.26). N. harperi tended to be more omnivorous, with freshwater sponges (Porifera) and midge

larvae (Chironomidae) dominating their diets. P. grandipinnis diet primarily consisted of

terrestrial insects, including ants (Formicidae) and terrestrial Dipterans, likely adult midges

(Figure 3.25). The differences in diet composition between P. grandipinnis and N. harperi

during periods of intermittency and flow were represented by a convergent two-dimensional

ordination using NMDS (stress=.15). Overlapping ellipses representing 95% confidence

indicated that N. harperi diet did not differ between isolated states and flowing states, while P.

grandipinnis diet differed only slightly (Figure 3.27).

Discussion

My results suggest that reproductive timing during periods of flow, and some life history

traits used to define the trilateral life history model, may be useful for understanding differences

in how even closely related species respond to changing environments. This study furthers the

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concept that life history traits, including smaller body size at maturity, lower fecundity, and

appropriate reproductive timing promotes greater persistence during more frequent and intense

changes in flow regime. A shift from flowing to isolation decreases the availability of terrestrial

prey, leaving species that depend on drifting prey (terrestrial input) to forage at an energetic

disadvantage, possibly leading to starvation. Known persistence rates, combined with life history

traits, may assist in identifying species not likely to survive in streams that shift from perennial

to intermittent. These relationships can be helpful to ecologists striving to understand flow-

ecological responses.

Studying closely related species allowed me to assess how life history traits may be

useful in explaining differing rates of persistence when faced with adverse conditions. Consistent

with my initial hypotheses, I found that Notropis harperi, the species with the greatest

persistence following drying, reproduced prior to the season when streams typically dry.

Conversely, Notropis petersoni, Notropis texanus, and Pteronotropis grandipinnis had at least a

portion of their reproductive stage during months of typical intermittent periods. The highest

GSI, gonad weight, and largest egg diameters, relative to body size, also belonged to N. harperi.

Since species life histories are closely tied to flow regime, some species may not be able

to find suitable flow environments for reproduction or survival because they lack adaptations to

intermittency. Alterations to environmental temperature regimes, often associated with flow

stagnation and climate change, can have a dramatic effect on the development, fitness, and

lifespan of fish (Wood and McDonald 1997, Clusella-Trullas et al. 2011, Andrews and

Schwarzkopf 2012). Temperature variation does not necessarily affect all life stages equally,

with early life history stages (e.g., embryos and larvae) often the most vulnerable because of

their sensitivity to temperature variation (Rombough 1997). In this system, elevated

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temperatures may have had the greatest effect on the earliest stages of development. Thus,

reproductive timing not concurrent with periods of intermittency (as is the case for N. harperi)

may allow for embryos and larvae to pass through vulnerable stages and reach a size where

temperature is less critical for persistence.

Previous studies of prairie stream cyprinids have shown that reproductive potential and

output are greatest during the earlier portion of the reproductive season (Bonner 2000, Durham

and Wilde 2005, Durham 2007), and that growth of early season juveniles exceeds that of

juveniles spawned later in the season (Durham and Wilde 2005). The earlier reproductive timing

of N. harperi, even by as little as two months, may be critical for maximizing the benefit of

flowing water necessary for reproduction. The shift to earlier reproductive timing may, in part,

be due to phenotypic plasticity. The ability of an organism to change its phenotype in response to

environmental changes is usually thought of as an adaptive strategy for dealing with differing

environments (Miner et al. 2005). N. harperi is most commonly associated with spring runs

having relatively low water temperatures. The resumption of flow in the intermittent streams of

the Ichawaynochaway Creek Basin (ICB) often occurs during the late fall and early winter, when

water temperatures are cool, which may act as a cue to initiate spawning. While the reproductive

schedule of N. harperi may in part be a phenotypic response induced by environmental cues, a

life history study of this species conducted by Marshal (1947) also found no reproductive

females between mid-July and November. This suggests that timing may be an evolutionarily

developed trait that has now become an advantage for persisting in intermittent streams.

Target species in this study are broadcast spawners, but only N. petersoni is known to

have pelagic eggs (Coburn 1986). Although N. petersoni was not classified as an intermittent

species, the indicator analysis suggests that it is more closely associated with intermittent streams

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than perennial streams (p=0.07, Appendix E). Perkin and Gido (2010) identified human

fragmentation in streams as an overarching cause of declining prairie stream fishes that produce

semi-buoyant eggs. With increases in stream intermittency in the lower Flint River Basin over

the last few decades, N. petersoni might be expected to decline given the reductions of flow for

dispersal of developing embryos. Without flow and continued mixing of waters, developing

embryos may sink to the bottom of isolated pools, where they are subjected to siltation, high

concentration of ammonia, and low dissolved oxygen (DO) levels. While the other study species

have non-buoyant/adhesive eggs, reproducing during intermittent events may still be detrimental

to young fishes if eggs are spawned in areas that become dry.

Parental investment via GSI and gonad weight were much greater in N. harperi,

indicating a substantial investment per individual offspring compared to the other target species.

In an Oklahoma prairie stream, Spranza and Stanley (2000) found that fish occupying areas of

the stream with greater environmental fluctuations were not at an energetic or reproductive

disadvantage, and had equal or greater GSI scores compared to fish occupying more stable areas.

This may indicate that species persisting in variable areas can tolerate apparently adverse

conditions by maximizing other benefits (e.g., low predation or competition during isolation). N.

harperi exploit periods of intermittency to their advantage, allowing for this greater investment

in offspring.

Egg diameter tended to be greater for N. harperi, but not significantly so. It has been

demonstrated that egg diameter can be related to standard length among fishes at different sites

(Heins and Baker 1987, Casten and Johnston 2008), which is consistent with my results. I found

a difference in egg diameter between N. texanus and P. grandipinnis, but not after accounting for

standard length. This indicates that the egg diameter difference between the species could be

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driven by female body size. In contrast, the significant difference in egg diameter between N.

harperi and P. grandipinnis, and the similarity between N. harperi and N. texanus, were

consistent and independent of female body size, suggesting differences among species will be

maintained within or among survey sites. Differences in egg diameter may be explained by

environmental variations affecting investment, whereby phenotypic variation and environmental

fluctuation may alter parental investment. Due to extirpation of species early on at intermittent

sites, this could not be examined more fully, therefore differences within and among these

species are still largely unknown.

Life-history theory predicts that the magnitude, frequency, and predictability of flow

regime events, including floods and droughts, can affect how organisms evolve (Iwasa and Levin

1995, Lytle and Poff 2004). When aquatic systems are altered from their natural flow regime, the

traits that allow species to persist over generations may become less suitable. It has been

demonstrated that flow alteration caused by dams can foster increases in non-native fishes

(Pringle et al. 2000). As flow becomes more stable, equilibrium strategists (ES) increase, while

periodic strategists (PS) and opportunistic strategists OS decrease (Mims and Olden 2013, Perkin

et al. 2017). Classification of strategy endpoints within the trilateral life history theory is

primarily dependent on what species are included in assessing distances to strategy endpoints.

On a basin scale, my results are similar to previous life history studies that follow Mims et al.

(2010) classification of species on the trilateral life history model, where fish assemblages in

intermittent stream classes tend to affiliate towards opportunistic endpoints (McManamay and

Frimpong 2015). However, species strongly associated with intermittent streams were classified

as both OS and PS. There were also no discernable differences among species commonly

associated with perennial streams, or weakly associated with either stream type. Within the

97

context of the species assemblage in the ICB (with rare species removed), all of the target

species were classified as OS. Based on persistence rates during isolated events (Figure 2.12,

Figure 2.13), we know that each of these species responds differently to intermittency.

When examining only target species against the trilateral history model, N. harperi was

classified as OS. This classification was driven primarily by its small length at maturation and

low fecundity. GSI and gonad weight is highest for N. harperi, indicating that although fecundity

was low, investment per egg was high. High investment per offspring, combined with

reproductive timing during periods when low-flows or intermittency do not occur, may be

indicative of why this species commonly occurs in intermittent streams rather than a small body

size at maturity and low fecundity. ES are predicted to be favored under stable conditions, and

this is consistent for N. texanus, the species with the lowest persistence rates. While predictions

coincided with persistence rates, spawning timing coinciding with intermittency may hinder N.

texanus from being successful in intermittent streams more so than a larger body size at maturity.

N. petersoni and P. grandipinnis are classified as PS, and are associated with predictable but

seasonally fluctuating environments. Consistent with findings that persistence is low during

isolation for these species, McManamay et al. (2015) found that PS species tended to decrease

with increasing intermittency; however, within their analysis all target species would have been

classified as OS rather than PS. It is because of this that life history models may be suited to

explaining general shifts in species trends, but may not be useful for exploring species-specific

changes to altered stream flow. Some aspects of the trilateral life history model, including length

at fecundity, as well as reproductive timing and diet can help explain why some species persist

better than others during intermittency.

98

Cyprinids are among the most abundant and diverse vertebrates in streams of the

southeastern US, and can therefore have potentially large effects on trophic dynamics and

ecological functions (Wheeler et al. 2017). Local extirpation of individuals in isolated pools did

not allow for assessing diet of all the target species. However, I did find a shift in feeding

patterns for P. grandipinnis, but not for N. harperi. For P. grandipinnis, this shift was primarily

driven by the lack of terrestrial diet items during isolated periods. Decreases in terrestrial

invertebrate input are known to lower fish biomass in headwater streams (Kawaguchi et al.

2003), as flow cessation lowers the availability of terrestrial prey items. While not evident in the

percent index of relative importance, guts of P. grandipinnis were frequently empty (personal

observation), which may suggest limited food availability. Contrary to hypotheses, shifts in diet

were not driven by a more advantageous opportunistic feeding pattern during isolation, but by

the lack of terrestrial prey items found. Given that the diet of N. harperi consists primarily of

aquatic items, the fact their feeding patterns did not shift likely means this species had an

energetic advantage for feeding during periods of isolation.

Life history traits of stream fishes partly reflect adaptations to natural flow regimes, and

trait-based approaches, including the trilateral life history model, have been useful for

understanding species responses to streamflow alteration. Native fishes in Coastal Plain stream

systems of southwestern GA, including the four study species, vary in occurrence and tolerance

to intermittency. I found that high persistence of one species corresponded to reproduction before

streams dry down, a small size at maturity, a greater reproductive investment, and a diet not

dependent on a terrestrial prey. All four species examined would be considered opportunistic

strategists within the trilateral life history model framework. However, my results suggest that

life history models, along with reproductive timing and diet, can help explain why even closely

99

related species differ in ability to persist in intermittent streams. The identification of species

differences in persistence probabilities in intermittent streams allows for the influence of life

history traits to be systematically explored. Understanding which life history traits are associated

with species persistence can help advance the understanding of flow ecology relationships.

100

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106

Table 3.1 Ovary and oocyte stages and descriptions of development based on oocyte size,

coloration, yolk condition, and physical location within the ovum modified from Heins and

Rabito (1986) and Heins and Baker (1987).

Stage Description

Latent (LA) Ovaries transparent-translucent, thin, small in

diameter; maturing oocytes present are

without visible yolk or with nuclei still

visible.

Early Maturing (EM) Ovaries translucent to opaque and small-

moderate sized inhabiting a larger portion of

the abdominal cavity; maturing oocytes are

varying in size, translucent to opaque, and

with nuclei obscured by yolk

Late Maturing (LM) Ovaries white to cream and small to greatly

enlarged; maturing oocytes are moderate to

large and white-cream or yellow-orange

Mature (MA) Ovaries cream to yellow and moderate to

greatly enlarged; multiple stages of oocyte

development including small oocytes

(translucent to opaque) and larger ripening

oocytes are opaque and cream to yellow but

without vitelline membrane separated from

yolk

Ripening (MR) Ovaries cream to yellow-orange and

moderately to greatly enlarged; two distinct

groups of oocytes present including a group

of small oocytes and a group of larger oocytes

that are translucent or transparent with the

chorions obviously separated from yolk

Ripe (RE) Ovaries cream to yellow and moderately sized

to greatly enlarged; two groups of relatively

large oocytes present including a group of

white-cream moderate to large maturing

oocytes and a group of translucent to

transparent ripe oocytes concentrated in the

posterior lumen of the ovary with the chorions

separated from yolk

107

Table 3.2: Standard lengths of males and females of four species assessed for reproductive

development (>25mm) from seven study sites in the Ichawaynochaway Creek Basin from May

2016- April 2017. Numbers of individuals (n), and mean lengths are shown along with standard

deviation (SD), standard error (SE), minimum (Min) and maximum (Max).

Gender n Mean SD SE Min Max

P. grandipinnis Male 60 45.42 7.43 0.96 25.74 59.12

N. harperi Male 38 32.75 5.88 0.95 26.22 51.08

N. petersoni Male 50 46.06 10.30 1.46 27.62 61.09

N. texanus Male 32 52.67 7.82 1.38 30.75 60.34

P. grandipinnis Female 92 39.59 6.31 0.66 26.02 54.44

N. harperi Female 82 32.98 6.23 0.69 25.76 57.32

N. petersoni Female 41 47.97 10.91 1.70 26.99 60.79

N. texanus Female 52 56.89 8.24 1.14 36.33 67.72

108

Table 3.3: Results from Chi-square tests of significance, which were performed separately on

sexually mature individuals and non-reproductive individuals. Significant differences are marked

with an * (p>.05) between the expected sex ratio of 1:1 and the observed sex ratio for males and

females of a given species.

Class chi-square df p.value

P. grandipinnis Reproductive 2.92 1 0.09

N. harperi Reproductive 0.17 1 0.68

N. petersoni Reproductive 0.10 1 0.75

N. texanus Reproductive 2.50 1 0.11

P. grandipinnis Non-reproductive 3.85 1 *0.05

N. harperi Non-reproductive 22.04 1 *2.6E-06

N. petersoni Non-reproductive 0.96 1 0.33

N. texanus Non-reproductive 2.27 1 0.13

109

Table 3.4: Standard lengths of mature males and females (Mature, Mature Ripening, or Ripe) of

four species assessed for reproductive development from seven study sites in the

Ichawaynochaway Creek Basin from May 2016- April 2017. Numbers of individuals (n), and

mean lengths are shown along with standard deviation (SD), standard error (SE), minimum

(Min) and maximum (Max).

Gender n Mean SD SE Min Max

P. grandipinnis Male 31 49.33 5.15 0.92 39.32 59.12

N. harperi Male 14 38.09 4.24 1.13 32.35 51.08

N. petersoni Male 21 55.32 3.69 0.81 49.2 61.09

N. texanus Male 15 57.05 3.12 0.81 49.26 60.34

P. grandipinnis Female 46 43.15 4.70 0.69 34.82 54.44

N. harperi Female 11 43.72 5.38 1.62 38.26 57.32

N. petersoni Female 19 55.15 3.48 0.80 46.84 60.79

N. texanus Female 25 58.48 4.71 0.94 49.47 67.72

110

Table 3.5: Standard lengths of all individuals captured during survey periods for length

distributions at thirteen study sites in the Ichawaynochaway Creek Basin from May 2016- April

2017. Individuals within the genus Notropis that were not identifiable in the field were

categorized as Notropis sp. Numbers of individuals (n), and mean lengths are shown along with

standard deviation (SD), standard error (SE), minimum (Min) and maximum (Max).


P. grandipinnis 439 39.10 10.90 0.52 13 68

N. harperi 1793 26.49 6.74 0.16 9 55

N. petersoni 201 40.06 14.86 1.05 13 70

N. texanus 114 55.00 9.94 0.93 21 70

Notropis sp. 178 12.76 2.29 0.17 9 21

111

Table 3.6: Summary statistics for egg size (mm) of mature, mature ripening, and ripe females

assessed for reproductive investment. Each individual had twenty eggs measured, where n is the

number of individuals assessed per species. Numbers of individuals (n), and mean lengths are

shown along with standard deviation (SD), standard error (SE), minimum (Min) and maximum

(Max).


P. grandipinnis 47 0.87 0.08 0.004 0.65 1.2

N. harperi 9 0.99 0.16 0.017 0.75 1.45

N. petersoni 18 0.88 0.10 0.008 0.65 1.15

N. texanus 25 0.98 0.12 0.007 0.75 1.3

112

Table 3.7: Species strategy weight and assignment for Soft Classification for opportunistic

strategist (OS), periodic strategist (PS), and equilibrium strategist (ES) strategist end points

calculated following Mims et al. (2010) for species identified in the Ichawaynochaway Creek

Basin (June 2015-January 2017). Species strategy weight was assessed using only the life history

traits of the four cyprinid species.

Stream Type OS PS EQ Soft Classification

N. harperi Intermittent 0.41 0.00 0.20 Opportunistic

N. petersoni Nonindicative 0.27 0.88 0.45 Periodic

N. texanus Perennial 0.00 0.47 0.92 Equilibrium

P. grandipinnis Perennial 0.43 0.52 0.23 Periodic

113

Figure 3.1: Mean probabilities of species-specific persistence for four adult cyprinid species found intermittent streams using a multi-

taxa, dynamic occupancy model over the weekly duration of isolation. Species-specific rates of persistence were averaged for each

species over 12 study sites and 14 weeks of continuous isolation in the Ichawaynochaway Creek Basin from 2015-2017 (Chapter 2).

Species-specific persistence rates were used to develop hypotheses for life history trait differences among N. harperi, N. petersoni, N.

texanus, and P. grandipinnis.

114

Figure 3.2: Locations of thirteen study sites within the Ichawaynochaway Creek Basin that were used to measure length distributions

for four cyprinid species and to obtain individuals for analyzing diet and reproductive characteristics, May 2016- April 2017. Apart

from Brantley Creek (the most north easterly circle) all survey streams are intermittent and experienced isolation or complete drying

during the survey period.

115

Figure 3.3: Standard length distribution to the nearest millimeter for all P. grandipinnis individuals found at thirteen study sites within

the Ichawaynochaway Creek Basin from May 2016- April 2017, plotted by Julian date. The horizontal line represents the minimum

reported length at maturity (34.82 mm standard length).

116

Figure 3.4: Observed GSI for P. grandipinnis females (upper left) and males (upper right) and standard length for females (bottom

left) and males (bottom right) of individuals assessed for reproductive state from within the Ichawaynochaway Creek Basin from May

2016- April 2017, plotted by Julian date. For females, the black symbols for MA, MR, and RE represent reproductively mature

individuals and the grey symbols for LA, EM, and LM represent reproductively latent or immature individuals. For males, black

symbols indicate mature males and the grey symbols indicate latent or immature individuals. The horizontal line for standard length

represents minimum observed length of reproductively mature females (34.82) and males (39.32).

117

Figure 3.5: Standard length distribution to the nearest millimeter for all N. harperi collected at thirteen study sites within the

Ichawaynochaway Creek Basin from May 2016- April 2017, plotted by Julian date. The horizontal line represents the minimum

reported length at maturity (34.82mm standard length).

118

Figure 3.6: Observed GSI for N. harperi females (upper left) and males (upper right) and standard length for females (bottom left) and

males (bottom right) of individuals assessed for reproductive state from within the Ichawaynochaway Creek Basin from May 2016-

April 2017, plotted by Julian date. For females, the black symbols for MA, MR, and RE represent reproductively mature individuals

and the grey symbols for LA, EM, and LM represent reproductively latent or immature individuals. For males, black symbols indicate

mature males and the grey symbols indicate latent or immature individuals. The horizontal line for standard length represents

minimum observed length of reproductively mature females (38.26) and males (32.35).

119

Figure 3.7 Standard length distribution to the nearest millimeter for all N. petersoni collected at thirteen study sites within the



120

Figure 3.8: Observed GSI for N. petersoni females (upper left) and males (upper right) and standard length for females (bottom left)

and males (bottom right) of individuals assessed for reproductive state from within the Ichawaynochaway Creek Basin from May




represents minimum observed length of reproductively mature females (46.84) and males (49.20).

121

Figure 3.9: Standard length distribution to the nearest millimeter for all N. texanus collected at thirteen study sites within the



122

Figure 3.10: Observed GSI for N. texanus females (upper left) and males (upper right) and standard length for females (bottom left)

and males (bottom right) of individuals assessed for reproductive state from within the Ichawaynochaway Creek Basin from May




represents minimum observed length reproductively mature females (49.47) and males (49.26).

123

Figure 3.11: Discharge at USGS 02354475 Spring Creek near Leary, GA (left y-axis) during the survey period. Light gray regions

indicate the Palmer Drought Index for the region (National Integrated Drought Information System, NIDIS; www.drought.gov). While

drought index values were exceptional from October to December of 2016, values were not exceptional for summer and early fall

moths (July-September).

http://www.drought.gov/

124

Figure 3.12: The Tukey adjusted comparison of trends of slopes for reproductive timing of

individuals of four cyprinid species using a ANCOVA. Points indicate the slope of the

probability curves for a given species with error bars indicating the 95% confidence intervals.

Results are given on the response scale (the natural log of a given date), where date 1is January

1st. Means sharing a letter are not significantly different by Tukey-adjusted mean separations.

125

Figure 3.13: Probability curves of presence of mature individuals of a given species over a year

time span. Normal confidence intervals are constructed on the link scale, and then back-

transformed to the response scale. The numeric date of 1 represents the first day of the calendar

year (January 1st).

126

Figure 3.14: The least square means of the standard length for mature individuals of four

cyprinid species using ANOVA. Points indicate the least square mean of the standard length by

species; error bars indicate the 95% confidence intervals using Tukey-adjusted comparisons.

Means sharing a letter are not significantly different by Tukey-adjusted mean separations.

127

Figure 3.15: The simple linear regression of the natural log of eviscerated mass and the natural log of standard length for all fishes of

an individual species combined were: P. grandipinnis, log(mass)= -12.55+3.40*log(length), F1,150=5846, p=<.001; N. harperi,

log(mass)=-11.69+3.16*log(length), F1,196=3987, p=<.001; N. petersoni, log(mass)=-12.12+3.26*log(length), F1,86=7519, p=<.001;

N. texanus, log(mass)=-11.83+3.20*log(length), F1,81=2522, p=<.001.

128

Figure 3.16: The least square means of the eviscerated mass for mature males of four

cyprinid species using ANCOVA. Points indicate the least square mean of the eviscerated

mass of an individual and error bars indicate the 95% confidence intervals using Tukey-

adjusted comparisons. Means sharing a letter are not significantly different by Tukey-

adjusted mean separations.

129


females of four cyprinid species using an ANCOVA. Points indicate the lease square

mean of the GSI of an individual and error bars indicate the 95% confidence intervals

using Tukey-adjusted comparisons. Means sharing a letter are not significantly different

by Tukey-adjusted mean separations. The ANCOVA was fit with a fixed effect of a given

species, a covariate of standard length, and a response variable of the GSI of an

individual female fish.

130


males of four cyprinid species using an ANCOVA. Points indicate the lease square mean

of the GSI of an individual and error bars indicate the 95% confidence intervals using

Tukey-adjusted comparisons. Means sharing a letter are not significantly different by

Tukey-adjusted mean separations. The ANCOVA was fit with a fixed effect of a given

species, a covariate of standard length, and a response variable of the GSI of an

individual male fish.

131

Figure 3.19: The least square means of gonad weight for mature females of four cyprinid

species using an ANCOVA. Points indicate the lease square mean of the gonad weight of

an individual and error bars indicate the 95% confidence intervals using Tukey-adjusted

comparisons. Means sharing a letter are not significantly different by Tukey-adjusted

mean separations. The ANCOVA was fit with a fixed effect of a given species, a

covariate of standard length, and a response variable of the gonad weight of an individual

female fish.

132

Figure 3.20: The least square means of gonad weight for mature males of four cyprinid

species using an ANCOVA. Points indicate the lease square mean of the gonad weight of

an individual and error bars indicate the 95% confidence intervals using Tukey-adjusted

comparisons. Means sharing a letter are not significantly different by Tukey-adjusted

mean separations. The ANCOVA was fit with a fixed effect of a given species, a

covariate of standard length, and a response variable of the gonad weight of an individual

male fish.

133

Figure 3.21: The least square means of egg diameter for four cyprinid species using a

nested ANOVA. Points indicate the lease square mean and error bars indicate the 95%

confidence intervals using Tukey-adjusted comparisons. Means sharing a letter are not

significantly different by Tukey-adjusted mean separations. The ANOVA was fit with a

fixed effect of species and with egg diameter nested within the individual fish it was

collected.

134

Figure 3.22: The least square means of four cyprinid species using a nested ANCOVA.

Points indicate the lease square mean and error bars indicate the 95% confidence

intervals using Tukey-adjusted comparisons. Means sharing a letter are not significantly

different by Tukey-adjusted mean separations. The ANCOVA was fit with a fixed effect

of species, a covariate of species length, with egg diameter nested within the individual

fish it was collected.

135

Figure 3.23: Ternary plot illustrating trilateral life history trade-offs in traits among commonly occurring species within the

Ichawaynochaway Creek basin. Axis scores indicate degree of species affiliation with opportunistic, periodic, or equilibrium

strategists. Species points are represented by which stream type they are associated with. The target species (P. grandipinnis, N.

harperi, N. petersoni, and N. texanus), represented by cross symbols, score highest on the opportunistic axis when evaluated in the

context of this assemblage.

136

Figure 3.24: Ternary plot illustrating trilateral life history trade-offs in traits among four cyprinid species, where axis scores indicate

degree of species affiliation with opportunistic, periodic, or equilibrium strategists.

137

Figure 3.25: Index of relative importance for samples of individual assessed for diet during flowing states and isolated states in the

Ichawaynochaway Creek Basin (May 2016- July 2016). Each of the twenty categories represents the total percent of the IRI for a

given sample where the number of individuals per sample ranged from one to nine. Prey categories were assigned to family or to the

lowest known taxonomic level.

138

Figure 3.26: Index of relative importance for subsamples of individuals assessed for diet during flowing states and isolated states in

the Ichawaynochaway Creek Basin (May 2016- July 2016). Each of the twenty categories represents the total percent of the IRI for a

given subsample where the number of individuals per subset ranged from one to five. Categories were assigned based on whether diet

taxa identified were aquatic, terrestrial, or an unknown category of “other” (e.g., detritus, eggs, and oligochaetes).

139

Figure 3.27: Non-metric multi-dimensional scaling (NMDS) ordination of %IRI for diet categories of all individuals assessed.

Grouping is based by species and the stream state when species were captured. Hollow symbols represent diet components for an

individual within a subsample for P. grandipinnis in parametric space and solid symbols represent diet components for an individual

N. harperi. Shapes of symbols represent the stream state when an individual was captured, with triangles representing periods of

flowing and circles are periods of isolation. Ellipses represent centroids and 95% confidence intervals for scores from grouping of

species and stream state.

140

CHAPTER 4

CONCLUSIONS AND SUMMARY

The development of the Ecological Limits of Hydrologic Alteration (ELOHA)

framework (Poff et al. 2010), along with placement of streams into hydrological classes,

provides a context for generalizing hydrologic disturbances, assembling and testing hypotheses

regarding ecological responses to hydrological disturbances, and developing environmental flow

standards (McManamay et al. 2015). By comparing ecological patterns between natural and

hydrologically altered streams, ecologists can develop flow-ecological response relationships

that can guide the creation of environmental flow standards (Arthington et al. 2006), whereby the

quantity, timing, and quality of water flows required to sustain freshwater ecosystems and human

livelihood are achieved (Poff et al. 2010). Understanding the biotic consequences of human

streamflow alteration is critical for successful environmental flow management.

Increased withdrawals from surface and groundwaters, coupled with climate change,

have altered stream hydrology in southwestern Georgia (Golladay et al. 2016). Recent multi-year

droughts have highlighted the need to apply flow-ecology relationships within the lower Flint

River basin, as the combined interaction of droughts and groundwater withdrawals have resulted

in many previously perennial streams ceasing to flow (Rugel et al. 2012). In particular, streams

crossing the Dougherty Plain, a recharge area for the upper Floridan aquifer region in the

southern portion of the lower Flint River basin, are prone to drying during periods of low rainfall

and high groundwater withdrawal (Opsahl et al. 2007). To understand the impacts of stream

drying and intermittency on biota within a tributary watershed of the lower Flint River,

141

Ichawaynochaway Creek, I examined fish assemblage variation across a gradient of flow

permanence, isolation, and reach position to quantify species-specific responses to changes in

abiotic conditions. I also explored life history traits to identify those most closely correlated with

species persistence in four cyprinids.

Fish community structure differs between intermittent and perennial streams, with the

former having a subset of species also occurring in perennial streams. I estimated rates of

species-specific occurrence, persistence, and colonization within streams that periodically ceased

flowing, and then determined the probability of occurrence in relation to environmental

characteristics. Indicator species of intermittent streams, as well as species that most commonly

occur, have the highest persistence rates as stream drying proceeds. Species that occurred in

isolated pools experienced greater rates of mortality as dissolved oxygen levels decreased and

ammonia concentrations increased. However, the ultimate driver of local extirpation was isolated

pools drying down to levels that exacerbated these effects. Like other studies of intermittent

streams, common species possessed both higher colonization and lower extinction rates than

rarer species (Whitney et al. 2016). Colonization of reaches after the resumption of flow

indicated that communities were able to recover, although ongoing intermittency decreased the

likelihood of recovery. This suggests that with the increasing frequency of low-flows and

intermittency within the ICB, indicator species of intermittent streams will become more

common, while indicator species of perennial streams will likely decline.

Four commonly occurring cyprinid species differed in rates of persistence in intermittent

streams during periods of isolation. For Notropis haperi, a species with high persistence rates,

reproductive timing did not overlap with typical seasonal stream drying, while species with low

persistence rates in isolated pools (Notropis petersoni, Notropis texanus, and Pteronotropis

142

grandipinnis) had at least a portion of their reproductive timing overlap with times when streams

were likely to dry. I demonstrated that some of the life history traits used to define the trilateral

life history model proposed by Kirk O. Winemiller (1989), including smaller body size at

maturity and low fecundity, may be useful for understanding how species respond to changing

environments. Studies have largely supported fish life history trait ordination along the three

axes of the trilateral life history model across North America (Kennard et al. 2010, Mims et al.

2010, Mims and Olden 2013, Perkin et al. 2017) as a means for predicting responses to natural

and altered flow regimes on fish assemblages through flow stabilization. Incorporating broadly

established models, feeding habits, and reproductive timing, may help in understanding

differences in how even closely related species respond to changing environments, while also

highlighting those traits most likely to promote greater persistence given more frequent and

intense disturbances.

Recent studies have indicated a need to quantify biotic response to changing flow

conditions through quantitative modeling (Arthington et al. 2006, Poff et al. 2010). Within the

lower Flint, water use during low-flow periods will result in both increased periods of flow

cessation in previously perennial streams, and increased dry periods in intermittent streams. The

southeastern US is noted for its aquatic faunal diversity, having the most diverse freshwater fish

fauna in North America (Burr and Mayden 1993). Reduction of diversity within streams due to

shifts in flow regime over time may result in the loss of sensitive species and an increase species

with high persistence.

143

References:

Arthington, A. H., S. E. Bunn, N. L. Poff, and R. J. Naiman. 2006. The challenge of providing

environmental flow rules to sustain river ecosystems. Ecological Applications 16:1311-

1318.

Burr, B. M., and R. L. Mayden. 1993. Phylogenetics and North American freshwater fishes.

Stanford University Press, Stanford, California.



achievable future conditions as a framework for guiding forest conservation and


Growns, I. 2004. A numerical classification of reproductive guilds of the freshwater fishes of

southeastern Australia and their application to river management. Fisheries management

and Ecology 11:369-377.

Kennard, M. J., S. J. Mackay, B. J. Pusey, J. D. Olden, and N. Marsh. 2010. Quantifying

uncertainty in estimation of hydrologic metrics for ecohydrological studies. River

Research and Applications 26:137-156.

McManamay, R. A., M. S. Bevelhimer, and E. A. Frimpong. 2015. Associations among

hydrologic classifications and fish traits to support environmental flow standards.

Ecohydrology 8:460-479.

Mims, M. C., and J. D. Olden. 2013. Fish assemblages respond to altered flow regimes via

ecological filtering of life history strategies. Freshwater Biology 58:50-62.

Mims, M. C., J. D. Olden, Z. R. Shattuck, and N. L. Poff. 2010. Life history trait diversity of

native freshwater fishes in North America. Ecology of Freshwater Fish 19:390-400.




Perkin, J. S., N. E. Knorp, T. C. Boersig, A. E. Gebhard, L. A. Hix, and T. C. Johnson. 2017.

Life history theory predicts long-term fish assemblage response to stream impoundment.

Canadian Journal of Fisheries & Aquatic Sciences 74:228-239.

Poff, N. L., B. D. Richter, A. H. Arthington, S. E. Bunn, R. J. Naiman, E. Kendy, M. Acreman,

C. Apse, B. P. Bledsoe, M. C. Freeman, J. Henriksen, R. B. Jacobson, J. G. Kennen, D.

M. Merritt, J. H. O'Keeffe, J. D. Olden, K. Rogers, R. E. Tharme, and A. Warner. 2010.

The ecological limits of hydrologic alteration (ELOHA): a new framework for

developing regional environmental flow standards. Freshwater Biology 55:147-170.

144



River basin, Georgia, USA. Hydrological Processes 26:523-534.

Vila-Gispert, A., R. Moreno-Amich, and E. García-Berthou. 2002. Gradients of life-history

variation: an intercontinental comparison of fishes. Reviews in Fish Biology & Fisheries

12:417.


approaches to developing flow-ecology relationships. Freshwater Biology. 2017 in press.

Whitney, J. E., K. B. Gido, E. C. Martin, and K. J. Hase. 2016. The first to arrive and the last to

leave: colonisation and extinction dynamics of common and rare fishes in intermittent

prairie streams. Freshwater Biology 61:1321-1334.

Winemiller, K. O. 1989. Patterns of Variation in Life History among South American Fishes in

Seasonal Environments. Oecologia 81(2):225-241.

145

APPENDICES

146

APPENDIX A: Species occurrence of taxa found for Chapter 2. Occurrence of taxa is for all individuals found in intermittent and

perennial stream sites in the Ichawaynochaway Creek basin. Collector names indicate the source of occurrence data: Davis (this

thesis), M. C. Freeman (USGS, unpublished), McPherson (2005), and McCargo (2004). Total occurrence is the number of samples a

species occurred in either intermittent streams (total intermittent samples, n=168) or perennial streams (total perennial samples, n=56).

Species removed from an analysis appeared in less than 5% of perennial samples or 5% of intermittent samples.

Collector Davis Davis Davis Davis Davis

Stream Name

Alligator

Big

Cypress 1

Big

Cypress 2 Keel 1 Keel 2

Stream Type Intermittent Perennial Intermittent Intermittent Intermittent Intermittent Intermittent

Species Name Species Code Total

Occurrence

Total

Occurrence

Ameiurus melas Ame.mel 0 1

Ameiurus natalis Ame.nat 27 8 X X X X

Ameiurus

nebulosus Ame.neb 0 2

Ameiurus

serracanthus Ame.ser 0 2

Amia calva Ami.cal 11 5 X

X X

Aphredoderus

sayanus Aph.say 99 39 X X X X X

Centrarchus

macropterus Cen.mac 45 4 X X X X

Cyprinella

venusta Cyp.ven 9 18 X X X

Cyprinus carpio Cyp.car 0 1

Elassoma

zonatum Ela.zon 108 7 X X X X X

147

Erimyzon sp. Eri.sp. 16 2 X X

Esox americanus Eso.ame 51 38 X X X

Esox niger Eso.nig 20 4 X X X X

Etheostoma

edwini Eth.edw 72 50 X X X X

Etheostoma

fusiforme Eth.fus 31 12 X X

Etheostoma

parvipinne Eth.par 0 3

Etheostoma

swaini Eth.swa 22 20 X X X X

Fundulus sp. Fun.sp. 3 1 X

Gambusia sp. Gam.sp. 140 34 X X X X X

Heterandria

formosa Het.for 3 5 X

Hybopsis sp. cf.

H. winchelli Hyb.win 2 18 X

Ichtalurus

punctatus Ich.pun 0 1

Ichthyomyzon

gagei Ich.gag 0 7

Labidesthes

sicculus Lab.sic 30 30 X X

Lepisosteus

oculatus Lep.ocu 1 8 X

Lepomis auritus Lep.aur 52 47 X X X X X

Lepomis cyanellus Lep.cya 11 2 X X

Lepomis gulosus Lep.gul 26 30 X X X X

Lepomis

macrochirus Lep.mac 102 41 X X X X X

Lepomis

marginatus Lep.mar 2 12 X X

148

Lepomis

microlophus Lep.mic 36 15 X X

Lepomis

punctatus x

miniatus

Lep.pXm 49 52 X X X X

Micropterus

cataractae Mic.cat 0 2

Micropterus

punctulatus Mic.pun 0 1

Micropterus

salmoides Mic.sal 30 30 X X

Minytrema

melanops Min.mel 24 12 X X X

Moxostoma

gammarion Mox.gam 0 1

Moxostoma

lachneri Mox.lac 0 2

Notemigonus

crysoleucas Not.cry 60 9 X X X X X

Notropis

chalybaeus Not.cha 0 5

Notropis harperi Not.har 120 17 X X X X X

Notropis

longirostris Not.lon 0 22

Notropis

maculatus Not.mac 1 4

Notropis

petersoni Not.pet 35 4 X X X X

Notropis texanus Not.tex 34 55 X X X X

Noturus

leptacanthus Not.lep 5 42

Opsopoeodus

emiliae Ops.emi 4 16 X X

149

Percina

nigrofasciata Per.nig 39 51 X X X X

Pteronotropis

welaka Pte.wel 31 54

Pteronotropis

grandipinnis Pte.gra 0 1 X

X

Pylodictus

olivaris Pyl.oli 0 2

Semotilus sp. Sem.sp. 0 2

150

Collector Davis Davis Davis Davis Davis Davis Davis Freeman

Stream Name Kiokee 1 Kiokee 2 Mill Neals Spring 1 Spring 2 Tallahassee Brantley

Stream Type Intermittent Intermittent Intermittent Intermittent Intermittent Intermittent Intermittent Perennial

Species Code

Ame.mel

Ame.nat X

X X X

Ame.neb

Ame.ser

Ami.cal X X

Aph.say X X X X X X X X

Cen.mac X X

X X X

Cyp.ven

X

X

Cyp.car

Ela.zon X X X X X X X

Eri.sp. X X X X X

Eso.ame X X X X X X

X

Eso.nig X X

X X

Eth.edw X X X X X X X X

Eth.fus X X

X X X

Eth.par

Eth.swa X X X X X X

X

Fun.sp.

X

Gam.sp. X X X X X X X X

Het.for

Hyb.win

Ich.pun

Ich.gag

Lab.sic X

X X X

X

Lep.ocu

X

151

Lep.aur X X X X X X X

Lep.cya X X X X

Lep.gul X X X X X X

Lep.mac X X X X X X X X

Lep.mar X

Lep.mic X X X X X

Lep.pXm X X X X X X X

Mic.cat

Mic.pun

Mic.sal X X X X X X

Min.mel X X X X X X

Mox.gam

Mox.lac

Not.cry X X X X X

Not.cha

Not.har X X X X X X X X

Not.lon

Not.mac

Not.pet X X X X X

Not.tex X X X X X

Not.lep X X X

Ops.emi X X

Per.nig X X X X X X X

Pte.wel

Pte.gra X X X X X X X

Pyl.oli

Sem.sp.

152

Collector Freeman Freeman Freeman Freeman Freeman Freeman Freeman Mcpherson Mcpherson

Stream Name Carter Chickasa-

watchee 3 Falling

Ichaway-

nochaway Herod Turkey Kiokee U Pachitla Spring U

Stream Type Perennial Perennial Perennial Perennial Perennial Perennial Intermittent Perennial Intermittent

Species Code

Ame.mel

Ame.nat X X X

Ame.neb

Ame.ser

Ami.cal X


Cen.mac X X

Cyp.ven X X X X

Cyp.car

Ela.zon X X X

Eri.sp. X

Eso.ame X X X X X X X X X

Eso.nig X X

Eth.edw X X X X X X X X X

Eth.fus X X X X

Eth.par

Eth.swa X X X

Fun.sp. X

Gam.sp. X X X X X X X X

Het.for X X

Hyb.win X X X

Ich.pun

Ich.gag X

Lab.sic X X X X X X X

Lep.ocu

153

Lep.aur X X X X X X X X X

Lep.cya X

Lep.gul X X X X X X X

Lep.mac X X X X X X X X X

Lep.mar X X X

Lep.mic X X X X X

Lep.pXm X X X X X X X X X

Mic.cat

Mic.pun

Mic.sal X X X X X X X X

Min.mel X X X X

Mox.gam

Mox.lac

Not.cry X X X X

Not.cha

Not.har X X X X X

Not.lon X X X X

Not.mac X X

Not.pet X

Not.tex X X X X X X X X

Not.lep X X X X X X X X

Ops.emi X X X X X

Per.nig X X X X X X X

Pte.wel X

Pte.gra X X X X X X X X

Pyl.oli

Sem.sp. X

154

Collector Mcpherson Mcpherson McCargo McCargo McCargo McCargo McCargo McCargo

Stream Name Wolf Carter Chickasa-

watchee 1

Chickasa-

whatchee 2

Chickasa-

whatchee 3

Ichaway-

nochaway 1

Ichaway-

nochaway 2

Ichaway-

nochaway 3

Stream Type Perennial Perennial Perennial Perennial Perennial Perennial Perennial Perennial

Species Code

Ame.mel X

Ame.nat X X X X X X

Ame.neb X X

Ame.ser X X

Ami.cal X X X X X


Cen.mac X X

Cyp.ven X X X X X X

Cyp.car X

Ela.zon X X

Eri.sp. X

Eso.ame X X X X X X X

Eso.nig X X

Eth.edw X X X X X X

Eth.fus X X X X X X

Eth.par X X X

Eth.swa X X X X X

Fun.sp.

Gam.sp. X X X X X X

Het.for

Hyb.win X X X X X X X

Ich.pun X

Ich.gag X

Lab.sic X X X X X X X X

155

Lep.ocu X X X X X X

Lep.aur X X X X X X X X

Lep.cya X X

Lep.gul X X X X X X X X

Lep.mac X X X X X X X X

Lep.mar X X X X X

Lep.mic X X X X X X X

Lep.pXm X X X X X X X X

Mic.cat X X

Mic.pun X

Mic.sal X X X X X X X X

Min.mel X

Mox.gam X

Mox.lac X X

Not.cry X X X X

Not.cha X X X X X

Not.har X X X X

Not.lon X X X

Not.mac

Not.pet X X

Not.tex X X X X X X X X

Not.lep X X X X X X X X

Ops.emi X X X X

Per.nig X X X X X X X X

Pte.wel

Pte.gra X X X X X X X X

Pyl.oli X X

Sem.sp.

156

References:

McCargo, J. W. 2004. Influence of drought of seasonal fish assemblages and habitat in the lower

Flint River Basin, Georgia. University of Georgia, Masters Thesis.

McPherson, R. D., Jr. 2005. An assessment of fish community structure and seasonal habitat use

of headwater confined channels and headwater wetlands in the lower Flint River Basin,

southwest Georgia. University of Georgia, Masters Thesis.

157

APPENDIX B: Detailed description of dynamic occupancy model. The model was used to

estimate species occurrence, colonization, and persistence in intermittent streams.

I used a Bayesian occupancy model approach to estimate true occupancy at a given site at

a given weekly interval given imperfect detection. Occupancy models consist of two

hierarchically coupled sub-models. The state model, the dynamic occupancy portion of the

model, governs the true change in species-specific occurrence at a site during a given sample.

The second model, the observation model, governs the probability of detecting a species given

that it is present or absent based on the state model (Appendix B.1). Following others (Freeman

et al. 2017), I treated the detection (1) or non-detection (0) of each taxon at each site on

successive dates as the result of week-to-week changes in occupancy coupled with the sampling

process (during which a species that was present may have been undetected). For the first date, I

assumed each species had its own probability of occurrence across study sites that had wetted

habitat (i.e., were either flowing or had isolated pools, but were not dry channels):

zmi1 ~ Bernoulli (ψm1) (1)

Here, zmi1 represents the true, yet unknown, presence or absence of species m at site i in the first

week, and ψm1 is the probability that species m is present at any site having wetted habitat (e.g.,

isolated pool or flowing) in the first week. For each subsequent week, I modeled species-specific

occupancy at each site as a function of persistence and colonization, provided that the site was

not dry:

zmik.| zmik-1 ~ Bernoulli (zmik-1 * Φmik-1 + (1- zmik-1)*γmik-1) (2)

where Φmik and γmik are species-, site- and week-specific probabilities of persistence and

colonization, respectively. For occasions when a site was dry, I set zmik to 0 for all taxa. I

158

modeled the actual detection (1) or non-detection (0) of each of the m species in the jth replicate

reach at site i in week k as:

ymijk | zmik ~ Bernoulli (zmik * pmijk) (3)

where pmik is the species- and sample-specific probability of detecting species m, given that

species m was present (zmik = 1).

I used this model to evaluate the evidence that species more commonly found in

intermittent streams in fact had either higher persistence when streams dried to isolated pools, or

recolonized more quickly when streams resumed flowing, than species that were more

commonly found in perennial streams (or than species that were not characteristic of either

stream type). For these analyses, I characterized flow condition during each interval between

sampling dates, at each site, as: (1) flowing on both dates (flowing to flowing); (2) flowing to

isolated (or vise versa); (3) isolated on both dates (isolated to isolated); (4) and if an isolated pool

is open to upstream or downstream movement (“isolated-open” to isolated-open” or “flowing to

isolated-open”) (Figure 2.3).

I then used site- and interval-specific flow condition, and species type (“intermittent”,

“perennial”, or “other”) as interacting covariates on fish persistence and colonization. I fit

covariates using a logit-link. The model for persistence was:

logit(Φmik ) = mean.phim + beta.phi.weeks.slackm*weeks.slackik + (4)

beta.phi.close.openm*isolated.openik + epsilon.phi.site.dateik

where, isolated.openik is equal to 1 if site i during interval k experienced conditions of “isolated

to isolated-open” or “flowing to isolated-open” and 0 otherwise. The term weeks.slackik was set

to 0 unless site i during interval k was dried to isolated pools, in which case weeks.slackik was the

number of successive weeks that I observed the site to have been in that state. I assumed that the

159

species could respond differently to each of these conditions, depending on whether the species

were assigned as “intermittent”, “perennial” or “other”. Thus, I defined each slope parameter

corresponding to these flow conditions as an additive combination of a mean effect (for

“intermittent” species) plus an effect of a species being either “perennial” or “other”. For

example, I set

beta.phi.weeks.slackm = mean.beta.phi.weeks.slack + (6)

beta.phi.cease.perennial * perennialm+ beta.phi. weeks.slack.other * otherm

where perennialm and otherm were set to 1 if species m was assigned as “perennial” or “other”,

and 0 otherwise. When sites were flowing, species-specific persistence (logit scale) reduced to

mean.phim plus a random effect (epsilon.phi.site.dateik, included for all conditions, assumed to be

normally distributed with a mean of 0) that represented otherwise unmodeled variation in

persistence among sites and weeks.

I similarly modeled effects of hydrologic state (flow condition) and species type

(intermittent, perennial, other) on colonization using a logit-link model. The colonization model

included three flow conditions: flowing (the baseline), isolated ((“isolated-open” to isolated-open

or isolated” or “flowing to isolated-open or isolated”)), with an added effect for how long since a

site transitioned from dry or isolated to flowing (which was represented by number of weeks the

site had been flowing when sampled and a 0 for dates that are isolated or dry). Similar to the

model for persistence, I estimated the effect of species type on species-specific values for

colonization in given alternative flow conditions. I also included two other covariates on

colonization, a bivariate term for weeks during the cool season (From November until March),

and the site distance to a perennial stream. Finally, I modeled detection as a function of three

covariates, also using a logit link whether a site was: completely isolated (e.g., a “small pool”

160

closed from any upstream or downstream movement); isolated-open (e.g., a “big pool” open to

upstream or downstream movement); and whether a stream was sampled with only an upstream

pass, or both an upstream and a downstream pass.

I modeled occupancy dynamics for adults and juveniles separately to evaluate evidence

that younger fish had higher persistence, or colonization, rates than adults. For each analysis, I

created a four-dimensional matrix of 21 species (model for adults including all taxa that occurred

in at least 5% of samples; 25 in the case of juveniles, Appendix C), by 12 sites, by two replicate

reaches sampled on each date, by 82 weekly samples (spanning June 2015 to January 2017).

Occasions when sites were not sampled in a particular week were coded as “NA”, as was the

second replicate on dates when I only sampled isolated pools at a site. I fit models with a

Bayesian framework implemented with the Markov chain Monte Carlo (MCMC) software JAGS

version 4.3.0 (Plummer 2003), run using the R package “jagsUI” (Kellner 2015) in R version

3.4.1 (R Core Team 2014). I used diffuse priors for parameter coefficients and ran three chains

for 30,000 iterations, thinned by four, after a burn-in of 3,000. I assessed convergence using the

Brooks-Gelman-Rubin statistic, R-hat (Brooks and Gelman 1998). I assessed model fit with a

Bayesian p-value for based on the discrepancy (Freeman-Tukey statistic) between the observed

and (model-based) expected number of species detected in each survey, and the same statistic

calculated for a replicate data set simulated using persistence, colonization, and detection

estimates at each MCMC iterations (Schaub and Kéry 2012). The Bayesian p-value was the

proportion of summed discrepancy values for the simulated data that exceeded the same for the

observed data. A value of less than about 0.05 or greater than about 0.95 would indicate

substantial model lack-of-fit (Schaub and Kéry 2012).

161

Appendix B.1: Directed acyclic graph illustrating the occupancy model structure. Green shading

represents the state model, blue shading represents the observation model, and the pink box

represents the observed survey data.

162

Refereces:

Brooks, S. P., and A. Gelman. 1998. General methods for monitoring convergence of iterative

simulations. Journal of Computational and Graphical Statistics 7:434-455.

Freeman, M. C., M. M. Hagler, P. M. Bumpers, K. Wheeler, S. Wenger, and B. Freeman. 2017.

Long-term monitoring data provide evidence of declining species richness in a river

valued for biodiversity conservation. Journal of Fish and Wildlife Management. 2017 in

press.

Kellner, K. 2015. jagsUI: a wrapper around rjags to streamline JAGS analyses. R package

version 1.

Plummer, M. 2003. JAGS: A program for analysis of Bayesian graphical models using Gibbs

sampling. Proceedings of the 3rd International workshop on distributed statistical

computing. Vienna, Austria.



Schaub, M., and M. Kéry. 2012. Combining information in hierarchical models improves

inferences in population ecology and demographic population analyses. Animal

Conservation 15:125-126.

163

APPENDIX C: Species and age class occurrences of taxa found at intermittent streams for

chapter 2. Total occurrences in a given stream state at intermittent sampling sites within the

Ichawaynochaway Creek Basin (June 2015-January 2017).

Species Age Class Flowing

n=77

Isolated

n=57

Isolated and

Open

n=9

Ameiurus natalis Adult 1 5 0

Ameiurus natalis Juvenile 1 14 1

Amia calva Adult 1 1 0

Amia calva Juvenile 3 3 0

Aphredoderus sayanus Adult 5 6 0

Aphredoderus sayanus Juvenile 27 39 5

Centrarchus macropterus Adult 4 5 0

Centrarchus macropterus Juvenile 13 14 1

Cyprinella venusta Adult 6 3 0

Cyprinella venusta Juvenile 0 0 0

Elassoma zonatum Adult 18 28 1

Elassoma zonatum Juvenile 16 28 2

Erimyzon sucetta Adult 3 0 2

Erimyzon sucetta Juvenile 6 2 0

Esox americanus Adult 6 2 0

Esox americanus Juvenile 24 9 3

Esox niger Adult 1 2 0

Esox niger Juvenile 10 4 0

Etheostoma edwini Adult 24 9 1

Etheostoma edwini Juvenile 28 12 1

Etheostoma fusiforme Adult 7 8 0

Etheostoma fusiforme Juvenile 12 5 1

Etheostoma swaini Adult 11 1 0

Etheostoma swaini Juvenile 7 2 0

Fundulus dispar Adult 1 1 0

Fundulus dispar Juvenile 1 0 0

Gambusia sp. Adult 53 48 8

Gambusia sp. Juvenile 38 49 6

Heterandria formosa Adult 0 0 0

Heterandria formosa Juvenile 0 1 0

Hybopsis sp. cf. H.

winchelli

Adult 0 0 0

164

Hybopsis sp. cf. H.

winchelli

Juvenile 0 2 0

Labidesthes sicculus Adult 8 2 0

Labidesthes sicculus Juvenile 8 9 2

Lepisosteus oculatus Adult 0 1 0

Lepisosteus oculatus Juvenile 0 0 0

Lepomis auritus Adult 15 8 1

Lepomis auritus Juvenile 15 11 2

Lepomis cyanellus Adult 2 3 0

Lepomis cyanellus Juvenile 4 2 0

Lepomis gulosus Adult 4 7 0

Lepomis gulosus Juvenile 2 7 0

Lepomis macrochirus Adult 17 22 2

Lepomis macrochirus Juvenile 35 38 4

Lepomis marginatus Adult 1 0 0

Lepomis marginatus Juvenile 1 1 0

Lepomis microlophus Adult 2 12 0

Lepomis microlophus Juvenile 8 14 2

Lepomis punctatus x

miniatus

Adult 11 9 1

Lepomis punctatus x

miniatus

Juvenile 13 13 3

Micropterus salmoides Adult 2 2 0

Micropterus salmoides Juvenile 10 12 3

Minytrema melanops Adult 2 4 0

Minytrema melanops Juvenile 5 5 1

Notemigonus crysoleucas Adult 6 8 0

Notemigonus crysoleucas Juvenile 10 25 3

Notropis harperi Adult 40 13 0

Notropis harperi Juvenile 50 33 5

Notropis petersoni Adult 11 8 0

Notropis petersoni Juvenile 9 10 1

Notropis texanus Adult 19 6 0

Notropis texanus Juvenile 9 4 0

Noturus leptacanthus Adult 1 1 0

Noturus leptacanthus Juvenile 0 0 0

Opsopoeodus emiliae Adult 1 2 0

Opsopoeodus emiliae Juvenile 1 0 0

Percina nigrofasciata Adult 20 4 0

Percina nigrofasciata Juvenile 14 8 0

Pteronotropis grandipinnis Adult 0 19 7

Pteronotropis grandipinnis Juvenile 0 9 5

165

APPENDIX D: R Code Used for Dynamic Occupancy Model. Model code for dynamic

occupancy used to explore the effects of intermittency on fishes in the Ichawaynocaway Creek

Basin from June 2015- January 2017. Model was fit with a Bayesian framework implemented

using Markov chain Monte Carlo (MCMC).

Model {

for (m in 1:ntaxa){

psi1[m] ~ dunif(0, 1) #Occupancy probability for each species, 1st sample date

for (i in 1:nsite){

for (k in 1:(ndate-1)){

logit(phi[m,i,k]) <-mean.phi[m] +

beta.phi.weeks.slack[m]*weeks.slack[i,k] + ## Number of weeks slack

beta.phi.close.open*isolated.open[i,k] + # 1 if interval between states is “flowing to

isolated-open” or “isolated-open to isolated-open”

epsilon.phi.site.date[i,k]

logit(gamma[m,i,k]) <- mean.gamma[m] +

beta.gamma.cool*cold[i,k+1] + ## 1 if cold on second date

beta.gamma.distance.s*distance.s[i] + ## Scaled distance to perennial stream

beta.gamma.weeks.flowing[m]*weeks.flowing[i,k]* (1-col.isolated[i,k])*(1-dry[i,k]) +

## The number of weeks flowing, centered on mean of 14.98 weeks.

beta.gamma.pools*col.isolated[i,k] + # Allows colonization to occur during isolated

events

epsilon.gamma.site.date[i,k]

}}}

for (m in 1:ntaxa){

for (i in 1:nsite){

for (k in 1:(ndate)){ ### model detection on each date as species specific probabilities +

random variation among sample events; assuming equal probability of detection among

replicates on a given date

logit(p[m,i,k]) <-mean.p[m] +

beta.p.small.pool*small.pool[i,k] + ## 1 if “isolated”

beta.p.big.pool*big.pool[i,k] + ## 1 if “isolated-open”

beta.p.updown*updown[i,k] + ## 1 if sampled with up and downstream passes

epsilon.p[i,k] }}

####specify flat or uninformative priors on mean persistence, colonization, detection logit scale

for (m in 1:ntaxa){

beta.phi.weeks.slack[m]<- mean.beta.phi.weeks.slack +

beta.phi.weeks.slack.other.effect*other.spps[m] +

beta.phi.weeks.slack.perennial.effect*perennial.spps[m] +

epsilon.phi.weeks.slack[m]

166

epsilon.phi.weeks.slack[m]~dnorm(0, tau.phi.weeks.slack)

beta.gamma.weeks.flowing[m]<-mean.beta.gamma.weeks.flowing +

beta.gamma.weeks.flowing.other.effect*other.spps[m] +

beta.gamma.weeks.flowing.perennial.effect*perennial.spps[m] +

epsilon.gamma.weeks.flowing[m]

epsilon.gamma.weeks.flowing[m]~dnorm(0, tau.gamma.weeks.flowing)

mean.p[m]~dnorm(0, 0.37) #flat prior on logit scale

mean.gamma[m]~dnorm(0, 0.37)

mean.phi[m]~dnorm(0, 0.37)

p.sp[m]<-1/(1+exp(-mean.p[m])) ### Back-transform to get the estimated detection rate for

each species

phi.sp[m]<-1/(1+exp(-mean.phi[m])) # This is the species-specific persistence when flowing

gamma.sp[m]<-1/(1+exp(-mean.gamma[m])) ## This is the species-specific gamma when

flowing for the average amount of time in the data set

}

beta.phi.weeks.slack.other.effect ~dnorm(0, .37)

beta.gamma.weeks.flowing.other.effect ~dnorm(0, .37)

beta.phi.weeks.slack.perennial.effect ~dnorm(0, .37)

beta.gamma.weeks.flowing.perennial.effect ~dnorm(0, .37)

beta.phi.close.open ~dnorm(0, 0.37)

beta.gamma.pools ~ dnorm(0, 0.37)

beta.gamma.distance.s~dnorm(0,0.37)

mean.beta.phi.weeks.slack~dnorm(0, 0.37)

beta.gamma.cool~dnorm(0, 0.37)

mean.beta.gamma.weeks.flowing~dnorm(0, 0.37)

beta.p.small.pool~dnorm(0, 0.37)

beta.p.big.pool~dnorm(0, 0.37)

beta.p.updown~dnorm(0, 0.37)

#### Specify flat or uninformative priors on random effects on persistence, colonization,

detection, where tau is precision, = 1/variance = 1/sd^2, where sigma is sd (standard deviation).

for (i in 1:nsite){

for (k in 1:ndate){

epsilon.phi.site.date[i,k]~dnorm(0, tau.phi) ### Random effect of survey on persistence

epsilon.gamma.site.date[i,k]~dnorm(0, tau.gamma) ### Random effect of survey on

colonization

epsilon.p[i,k]~dnorm(0, tau.p) ## Random effect of survey on overall detection

}}

sigma.phi~dunif(0,1)

tau.phi<-pow(sigma.phi,-2)

var.phi<-pow(sigma.phi,2)

sigma.gamma~dunif(0,1)

tau.gamma<-pow(sigma.gamma,-2)

var.gamma<-pow(sigma.gamma,2)

sigma.p~dunif(0,1)

tau.p<-pow(sigma.p,-2)

167

var.p<-pow(sigma.p,2)

sigma.phi.weeks.slack~dunif(0,1)

tau.phi.weeks.slack<-pow(sigma.phi.weeks.slack,-2)

var.phi.weeks.slack<-pow(sigma.phi.weeks.slack,2)

sigma.gamma.weeks.flowing~dunif(0,1)

tau.gamma.weeks.flowing<-pow(sigma.gamma.weeks.flowing,-2)

var.gamma.weeks.flowing<-pow(sigma.gamma.weeks.flowing,2)

## Extracting effects from model for species effects.

avg.mean.phi.perennial<-(sum(mean.phi*perennial.spps))/(sum(perennial.spps)) # Average

'mean.phi' for perennial species

avg.mean.phi.other<-(sum(mean.phi*other.spps))/(sum(other.spps)) # Average 'mean.phi' for

other species

avg.mean.phi.intermittent <- (sum(mean.phi*(1-(other.spps+perennial.spps))))/(sum(1-

(other.spps+perennial.spps))) # Average 'mean.phi' for intermittent species

avg.mean.gamma.perennial <- (sum(mean.gamma*perennial.spps))/(sum(perennial.spps)) #

Average ‘mean.gamma’ for perennial species

avg.mean.gamma.other <- (sum(mean.gamma*other.spps))/(sum(other.spps)) # Average

‘mean.gamma’ for other species

avg.mean.gamma.intermittent <-(sum(mean.gamma*(1-(other.spps+perennial.spps))))/(sum(1-

(other.spps+perennial.spps))) # Average ‘mean.gamma’ for intermittent species

##Extracting species specific effects of the number of weeks slack and the number of weeks

flowing for all species where 29 and 101 are the number of weeks isolated or flowing,

respectively, multiplied by two.

for (i in 1:29){

phi.perennial[i]<-1/(1+exp(-avg.mean.phi.perennial - ((mean.beta.phi.weeks.slack +

beta.phi.weeks.slack.perennial.effect)*weeks.isolated[i]))) # When weeks.slack=0, this is just

phi in an isolated pool.

phi.other[i]<-1/(1+exp(-avg.mean.phi.other - ((mean.beta.phi.weeks.slack +

beta.phi.weeks.slack.other.effect)*weeks.isolated[i])))

phi.intermittent[i]<-1/(1+exp(-avg.mean.phi.intermittent -

(mean.beta.phi.weeks.slack*weeks.isolated[i]))) }

for(i in 1:101){

gamma.perennial[i]<-1/(1+exp(-avg.mean.gamma.perennial -

((mean.beta.gamma.weeks.flowing +

beta.gamma.weeks.flowing.perennial.effect)*weeks.continuous.flow[i])))

gamma.other[i]<-1/(1+exp(-avg.mean.gamma.other - ((mean.beta.gamma.weeks.flowing +

beta.gamma.weeks.flowing.other.effect)*weeks.continuous.flow[i])))

gamma.intermittent[i]<-1/(1+exp(-avg.mean.gamma.intermittent -

(mean.beta.gamma.weeks.flowing*weeks.continuous.flow[i])))

}

for (m in 1:ntaxa){

for (i in 1:29){

phi.species.slack[m,i]<- 1/(1+exp(-mean.phi[m] -

(beta.phi.weeks.slack[m]*weeks.isolated[i])))

} }

168

for (m in 1:ntaxa){

for (i in 1:101){

gamma.species.flowing[m,i] <-1/(1+exp(-mean.gamma[m] -

(beta.gamma.weeks.flowing[m]*weeks.continuous.flow[i])))

} }

# Ecological submodel: Define state conditional on parameters

for (m in 1:ntaxa){

for (i in 1:nsite){

z.wet[m,i,1] ~ dbern(psi1[m]) #psi1 = prob of occurrence on first date is constant across sites,

for each species

z.possible[m,i,1]<-max(0.00001, min((1-dry[i,1]), z.wet[m,i,1]))

z[m,i,1]~dbern(z.possible[m,i,1]) # if dry, set z on 1st date to 0

for (k in 2:ndate){

muZ[m,i,k]<- max((dry[i,k]*0.00001), min((1-dry[i,k]), (z[m,i,k-1]*(phi[m,i,k-1]) + (1-

z[m,i,k-1])*gamma[m,i,k-1])))

z[m,i,k] ~ dbern(muZ[m,i,k])

}}}

# Observation model

for (m in 1:ntaxa){

for (i in 1:nsite){

for (j in 1:nrep){

for (k in 1:ndate){

muy[m,i,j,k] <- z[m,i,k]*p[m,i,k]

y[m,i,j,k] ~ dbern(muy[m,i,j,k])

ynew[m,i,j,k]~dbern(muy[m,i,j,k]) #simulated observations so we can check model fit,

posterior predictive check below

}}}

########## Posterior predictive check ########

for (i in 1:nsite){

for (j in 1:nrep){

for (k in 1:ndate){

obsrich[i,j,k]<-sum(y[,i,j,k]) ## observed richness by survey

simrich[i,j,k]<-sum(ynew[,i,j,k]) ## simulated richness by survey

exprich[i,j,k]<-sum(muy[,i,j,k]) ## expected richness

depobs[i,j,k]<-pow((pow(obsrich[i,j,k], 0.5)-pow(exprich[i,j,k], 0.5)), 2) # freeman-tukey

measure of departure from expected, observed data

depsim[i,j,k]<-pow((pow(simrich[i,j,k], 0.5)-pow(exprich[i,j,k], 0.5)), 2) # freeman-tukey

measure of departure from expected, simulated data

} }}

fit<-sum(depobs[,,]) #discrepancy, observed data

fit.sim<-sum(depsim[,,]) #discrepancy, simulated data

}

169

APPENDIX E: Indicator analysis and classification for species strategist endpoints, showing species results from indicator species

analysis for reach type, and strategy weight and classification for species strategist endpoints. Indicator species analysis assessed how

strongly species were correlated with a given stream type (intermittent or perennial), where A is the probability that the surveyed

stream site belongs to the stream type given the fact that the taxon has been found, with a value of 1.0 if a species has only been found

in this group and B is the probability of finding the taxon in a site belonging to this stream type with a value of 1.0 if a species appears

in all sites belonging to this group. The p-value was used to assign stream type associations, where nonindicative species were not

significantly correlated with a stream type. Species strategy weight and assignment for Soft Classification for opportunistic (Opp),

periodic (Per), and equilibrium (Dudgeon et al. 2006) strategist end points were calculated following Mims et al. (2010) for species

identified in the Ichawaynochaway Creek Basin (June 2015-January 2017). Species strategy weight was not assessed for any species

missing life history traits.

Species A B p. value Stream Type Opp Per Equ

Soft

Classification

Elassoma zonatum 0.84 0.67 0.005** Intermittent 0.55 0.19 0.30 Opportunistic

Gambusia sp. 0.59 0.86 0.005** Intermittent 0.43 0.14 0.34 Opportunistic

Notropis harperi 0.71 0.73 0.005** Intermittent 0.80 0.24 0.17 Opportunistic

Centrarchus macropterus 0.79 0.27 0.02* Intermittent 0.35 0.50 0.41 Periodic

Notemigonus crysoleucas 0.70 0.37 0.05* Intermittent 0.32 0.57 0.19 Periodic

Notropis petersoni 0.75 0.22 0.07 Nonindicative 0.64 0.42 0.26 Opportunistic

Erimyzon sp. 0.73 0.10 0.26 Nonindicative 0.30 0.71 0.36 Periodic

Esox niger 0.63 0.12 0.44 Nonindicative 0.21 0.75 0.48 Periodic

Lepomis cyanellus 0.66 0.07 0.59 Nonindicative 0.38 0.43 0.47 Equilibrium

Ameiurus natalis 0.54 0.17 0.85 Nonindicative 0.18 0.39 0.75 Equilibrium

Cyprinella venusta 0.87 0.32 0.005** Perennial 0.64 0.30 0.32 Opportunistic

Esox americanus 0.68 0.68 0.005** Perennial 0.43 0.60 0.43 Periodic

Etheostoma edwini 0.67 0.89 0.005** Perennial - - - -

170

Hybopsis sp. cf. H. winchelli 0.96 0.32 0.005** Perennial - - - -

Ichthyomyzon gagei 1.00 0.13 0.005** Perennial 0.43 0.57 0.47 Periodic

Labidesthes sicculus 0.74 0.54 0.005** Perennial 0.60 0.47 0.28 Opportunistic

Lepisosteus oculatus 0.96 0.14 0.005** Perennial 0.19 0.69 0.56 Periodic

Lepomis auritus 0.72 0.84 0.005** Perennial 0.37 0.41 0.50 Equilibrium

Lepomis gulosus 0.77 0.54 0.005** Perennial 0.27 0.48 0.46 Periodic

Lepomis marginatus 0.95 0.21 0.005** Perennial - - - -

Lepomis punctatus x miniatus 0.75 0.93 0.005** Perennial 0.31 0.37 0.45 Equilibrium

Micropterus salmoides 0.74 0.54 0.005** Perennial 0.14 0.54 0.58 Equilibrium

Notropis chalybaeus 1.00 0.09 0.005** Perennial 0.74 0.32 0.17 Opportunistic

Notropis longirostris 1.00 0.39 0.005** Perennial 0.77 0.28 0.17 Opportunistic

Notropis texanus 0.82 0.98 0.005** Perennial 0.67 0.37 0.20 Opportunistic

Noturus leptacanthus 0.96 0.75 0.005** Perennial 0.43 0.04 0.40 Opportunistic

Opsopoeodus emiliae 0.92 0.29 0.005** Perennial 0.58 0.31 0.40 Opportunistic

Percina nigrofasciata 0.79 0.91 0.005** Perennial 0.64 0.28 0.35 Opportunistic

Pteronotropis grandipinnis 0.83 0.96 0.005** Perennial 0.63 0.33 0.29 Opportunistic

Etheostoma swaini 0.72 0.36 0.01** Perennial 0.67 0.21 0.32 Opportunistic

Etheostoma parvipinne 1.00 0.05 0.02* Perennial 0.69 0.28 0.32 Opportunistic

Notropis maculatus 0.92 0.07 0.02* Perennial 0.67 0.30 0.29 Opportunistic

Heterandria formosa 0.83 0.09 0.05* Perennial 0.63 0.00 0.20 Opportunistic

Lepomis macrochirus 0.54 0.73 0.07 Nonindicative 0.23 0.50 0.50 Periodic

Minytrema melanops 0.59 0.21 0.18 Nonindicative 0.20 0.76 0.44 Periodic

Aphredoderus sayanus 0.53 0.70 0.24 Nonindicative 0.54 0.19 0.39 Opportunistic

Lepomis microlophus 0.55 0.27 0.47 Nonindicative 0.21 0.55 0.51 Periodic

Amia calva 0.57 0.09 0.60 Nonindicative 0.00 0.46 0.71 Equilibrium

Etheostoma fusiforme 0.53 0.21 0.71 Nonindicative 0.70 0.24 0.29 Opportunistic

171

References:

Dudgeon, D., A. H. Arthington, M. O. Gessner, Z. Kawabata, D. J. Knowler, C. Leveque,

R. J. Naiman, A. H. Prieur-Richard, D. Soto, M. L. Stiassny, and C. A. Sullivan.

2006. Freshwater biodiversity: importance, threats, status and conservation

challenges. Biological Reviews of the Cambridge Philosophical Society 81:163-

182.

Mims, M. C., J. D. Olden, Z. R. Shattuck, and N. L. Poff. 2010. Life history trait

diversity of native freshwater fishes in North America. Ecology of Freshwater

Fish 19:390-400.

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