U.S. Department of the Interior Southeast Climate Science Center

Insular Ecosystems of the Southeastern United States:

A Regional Synthesis to Support Biodiversity Conservation

in a Changing Climate

U.S. Department of the Interior

U.S. Geological Survey

Professional Paper 1828
Cover photographs, left column, top to bottom:

Photographs are by Alan M. Cressler, U.S. Geological Survey, unless noted otherwise. 

Ambystoma maculatum (spotted salamander) in a Carolina bay on the eastern shore of Maryland. Photograph

by Joel Snodgrass, Virginia Polytechnic Institute and State University.

Geum radiatum, Roan Mountain, Pisgah National Forest, Mitchell County, North Carolina.

Dalea gattingeri, Chickamauga and Chattanooga National Military Park, Catoosa County, Georgia.

Round Bald, Pisgah and Cherokee National Forests, Mitchell County, North Carolina, and Carter County, Tennessee. 

Cover photographs, right column, top to bottom:

Habitat monitoring at Leatherwood Ford cobble bar, Big South Fork Cumberland River, Big South Fork National 

River and Recreation Area, Tennessee. Photograph by Nora Murdock, National Park Service.

Soil island, Davidson-Arabia Mountain Nature Preserve, Dekalb County, Georgia. Photograph by Alan M. Cressler, 

U.S. Geological Survey.

Antioch Bay, Hoke County, North Carolina. Photograph by Lisa Kelly, University of North Carolina at Pembroke.
Insular Ecosystems of the Southeastern 

United States: A Regional Synthesis to 

Support Biodiversity Conservation in a 

Changing Climate

By Jennifer M. Cartwright and William J. Wolfe

U.S. Department of the Interior Southeast Climate Science Center

Professional Paper 1828

U.S. Department of the Interior

U.S. Geological Survey
U.S. Department of the Interior

SALLY JEWELL, Secretary

U.S. Geological Survey

Suzette M. Kimball, Director

U.S. Geological Survey, Reston, Virginia: 2016

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Suggested citation:

Cartwright, J.M., and Wolfe, W.J., 2016, Insular ecosystems of the southeastern United States—A regional synthesis 

to support biodiversity conservation in a changing climate: U.S. Geological Survey Professional Paper 1828, 162 p., 

http://dx.doi.org/10.3133/pp1828.

http://catalog.loc.gov/

ISSN 1044-9612 (print)

ISSN 2330-7102 (online)

ISBN 978-1-4113-4067-1
iii

Contents

Acknowledgments .....................................................................................................................................viii

Abstract...........................................................................................................................................................1

Chapter A. Introduction ...............................................................................................................................3

Background............................................................................................................................................4

Purpose and Scope ..............................................................................................................................4

Approach................................................................................................................................................4

References.............................................................................................................................................6

Chapter B. Granite Outcrops of the Piedmont .........................................................................................9

Introduction............................................................................................................................................9

Geographic Range................................................................................................................................9

Physical Geography............................................................................................................................11

Stress and Disturbance Regimes.....................................................................................................11

Community Types and Vegetation Dynamics .................................................................................12

Contributions to Regional Biodiversity............................................................................................15

Conservation Considerations............................................................................................................18

References...........................................................................................................................................20

Chapter C. Limestone Cedar Glades........................................................................................................23

Introduction..........................................................................................................................................23

Geographic Range..............................................................................................................................24

Physical Geography............................................................................................................................25

Stress and Disturbance Regimes.....................................................................................................26

Community Types and Vegetation Dynamics .................................................................................27

Differentiation from Related Ecosystems .......................................................................................28

Contributions to Regional Biodiversity............................................................................................29

Conservation Considerations............................................................................................................32

References...........................................................................................................................................35

Chapter D. Xeric Limestone Prairies .......................................................................................................39

Introduction..........................................................................................................................................39

Geographic Range and Differentiation from Related Ecosystems.............................................39

Physical Geography............................................................................................................................42

Stress and Disturbance Regimes.....................................................................................................42

Community Types and Vegetation Dynamics .................................................................................44

Contributions to Regional Biodiversity............................................................................................46

Conservation Considerations............................................................................................................47

References...........................................................................................................................................50

Chapter E. Mid-Appalachian Shale Barrens..........................................................................................55

Introduction..........................................................................................................................................55

Geographic Range..............................................................................................................................55

Physical Geography............................................................................................................................57

Stress and Disturbance Regimes.....................................................................................................58

Community Types and Vegetation Dynamics .................................................................................59

Contributions to Regional Biodiversity............................................................................................60

Conservation Considerations............................................................................................................62

References...........................................................................................................................................64
iv

Chapter F. High-Elevation Outcrops and Balds of the Southern Appalachians ...............................67

Introduction..........................................................................................................................................67

Geographic Range..............................................................................................................................68

Physical Geography............................................................................................................................69

Stress and Disturbance Regimes.....................................................................................................70

Community Types and Vegetation Dynamics .................................................................................72

Contributions to Regional Biodiversity............................................................................................76

Conservation Considerations............................................................................................................79

References...........................................................................................................................................81

Chapter G. Carolina Bays ..........................................................................................................................85

Introduction..........................................................................................................................................85

Geographic Range .............................................................................................................................85

Physical Geography............................................................................................................................87

Stress and Disturbance Regimes.....................................................................................................88

Community Types and Vegetation Dynamics .................................................................................90

Contributions to Regional Biodiversity............................................................................................94

Conservation Considerations............................................................................................................97

References.........................................................................................................................................100

Chapter H. Karst-Depression Wetlands................................................................................................105

Introduction........................................................................................................................................105

Geographic Range............................................................................................................................105

Physical Geography..........................................................................................................................107

Stress and Disturbance Regimes...................................................................................................111

Community Types and Vegetation Dynamics ...............................................................................113

Contributions to Regional Biodiversity..........................................................................................115

Conservation Considerations..........................................................................................................118

References.........................................................................................................................................121

Chapter I. Riverscour Ecosystems.........................................................................................................125

Introduction........................................................................................................................................125

Geographic Range............................................................................................................................126

Physical Geography..........................................................................................................................126

Stress and Disturbance Regimes...................................................................................................128

Community Types and Vegetation Dynamics ...............................................................................129

Contributions to Regional Biodiversity .........................................................................................133

Conservation Considerations..........................................................................................................135

References.........................................................................................................................................138

Chapter J. Conclusions and Implications.............................................................................................141

References.........................................................................................................................................143

Glossary.......................................................................................................................................................147

Appendix 1. Ecological System Names According to the International Terrestrial

Ecological Systems Classification ............................................................................................149

Appendix 2. Component Associations According to the International Terrestrial

Ecological Systems Classification ............................................................................................150

Appendix 3. Selected Plant Taxa of Conservation Concern in Insular Ecosystems

of the Southeastern United States ............................................................................................154

References Cited in Appendixes.............................................................................................................160
v

Figures [All are photographs unless noted otherwise]

B–1. Camp Meeting Rock Preserve, Heard County, Georgia .........................................................9

B–2. Map showing geographic distribution of Piedmont granite outcrops...............................10

B–3. Solution pits in a granite outcrop, Davidson-Arabia Mountain Nature Preserve,

Dekalb County, Georgia .............................................................................................................11

B–4. Moss-colonized rock surface, Davidson-Arabia Mountain Nature Preserve,

Dekalb County, Georgia .............................................................................................................13

B–5. Soil island, Davidson-Arabia Mountain Nature Preserve, Dekalb County, Georgia ......14

B–6. Characteristic plant taxa of Piedmont granite outcrops......................................................17

C–1. Cedar Glade Number 5, Chickamauga and Chattanooga National Military Park, 

Catoosa County, Georgia...........................................................................................................23

C–2. Map showing general geographic distribution of limestone cedar glades .....................24

C–3. Grass-filled cross joint, Chickamauga and Chattanooga National Military Park, 

Catoosa County, Georgia...........................................................................................................25

C–4. Standing water, Flat Rock Cedar Glades and Barrens State Natural Area, 

Rutherford County, Tennessee .................................................................................................26

C–5. Characteristic plant taxa of limestone cedar glades ...........................................................31

D–1. Grand Prairie, Coosa Valley Prairies, Floyd County, Georgia .............................................39

D–2. Map showing general geographic distribution of xeric limestone prairies ....................40

D–3. Ketona dolomite glade, Kathy Stiles Freeland Bibb County Glades Preserve, 

Bibb County, Alabama................................................................................................................41

D–4. Dry Prairie, Coosa Valley Prairies, Floyd County, Georgia ..................................................44

D–5. Rare plant taxa in Ketona dolomite glades at the Kathy Stiles Freeland 

Bibb County Glades Preserve, Bibb County, Alabama ........................................................47

E–1. Shale barren above a cowpasture, Bath County, Virginia...................................................55

E–2. Map showing general geographic distribution of mid-Appalachian shale barrens ......56

E–3. Shale barren, Edinburg Gap, George Washington and Jefferson National Forests, 

Shenandoah County, Virginia....................................................................................................57

E–4. Shale barren undercut by a stream, Bath County, Virginia .................................................58

E–5. Characteristic plant taxa of mid-Appalachian shale barrens.............................................61

F–1. Round Bald, Pisgah and Cherokee National Forests, Mitchell County, 

North Carolina, and Carter County, Tennessee ....................................................................67

F–2. Map showing geographic distribution of high-elevation outcrops and balds 

in the southern Appalachians ..................................................................................................68

F–3. Grass bald and heath bald mosaic at Round Bald, Pisgah and Cherokee National 

Forests, Mitchell County, North Carolina, and Carter County, Tennessee........................73

F–4. Grass bald at Round Bald, Pisgah and Cherokee National Forests, 

Mitchell County, North Carolina, and Carter County, Tennessee .......................................73

F–5. Rock outcrop at Sassafrass Mountain, Nantahala National Forest, 

Jackson County, North Carolina ..............................................................................................75

F–6. High-elevation rock outcrop community, Craggy Dome, Blue Ridge Parkway,

Buncombe County, North Carolina ..........................................................................................75

F–7. Characteristic taxa of high-elevation outcrops and balds ..................................................77
vi

G–1. Aerial view of Dry Bay, Savannah River National Environmental

Research Park, South Carolina ................................................................................................85

G–2. Map showing regions of concentrated occurrence of Carolina bays .............................86

G–3. A Carolina bay at Francis Marion National Forest, South Carolina ...................................89

G–4. Burned pocosin vegetation, Pocosin Lakes National Wildlife Refuge,

North Carolina .............................................................................................................................89

G–5. Seventeen Frog Pond, Scotland County, North Carolina .....................................................90

G–6. Antioch Bay, Hoke County, North Carolina.............................................................................91

G–7. Hamby’s Bay, Hoke County, North Carolina ...........................................................................91

G–8. Pocosin vegetation, Pocosin Lakes National Wildlife Refuge, North Carolina................93

G–9. Diagram showing relationship between Carolina bays and pocosins ..............................93

G–10. Oxypolis canbyi, Lee County, Georgia.....................................................................................95

G–11. Amphibian taxa supported by Carolina bay habitat..............................................................96

H–1. Harriston Ponds, Augusta County, Virginia ..........................................................................105

H–2. Map showing general geographic distribution of selected areas containing

concentrations of karst-depression wetlands in the southeastern United States........106

H–3. Nyssa aquatica stand in Westall Swamp, Arnold Air Force Base, 

Coffee County, Tennessee.......................................................................................................108

H–4. Deep Pond, Maple Flats Pond Complex, Augusta County, Virginia..................................109

H–5. Deep Run Ponds, Rockingham County, Virginia ..................................................................110

H–6. Two views of Sinking Pond, Arnold Air Force Base, Coffee County, Tennessee:

fully drained with leaves on in the summer, and fully drained with snow in

the winter...................................................................................................................................112

H–7. Neyami Pond, Lee County, Georgia ......................................................................................113

H–8. Davidson Run Pond, Rockbridge County, Virginia...............................................................115

H–9. Helenium virginicum at Twin Ponds, Maple Flats Pond Complex, 

Augusta County, Virginia .........................................................................................................116

I–1. Rough Shoals, Big South Fork Cumberland River, Big South Fork

National River and Recreation Area, Tennessee ................................................................125

I–2. Map showing selected locations of riverscour ecosystems in the

southeastern United States ...................................................................................................127

I–3. Appalachian Riverside Flatrock Community, Sandstone Falls, New River, 

New River Gorge National River, Raleigh County, West Virginia .....................................129

I–4. Bedrock outcrop exposures along the New River at Camp Brookside, 

New River Gorge National River, Summers County, West Virginia .................................130

I–5. Rare taxa associated with riverscour habitat in the southeastern United States ........134

I–6. Habitat monitoring at Leatherwood Ford cobble bar, Big South Fork Cumberland

River, Big South Fork National River and Recreation Area, Tennessee..........................137

Tables

C–1. Factors differentiating limestone cedar glades from xeric limestone prairies................29

I–1. Selected streams supporting riverscour ecosystems within the 

southeastern United States ....................................................................................................126

I–2. Selected occurrences of riverscour ecosystems on protected public land..................137

J–1. Examples of additional insular ecosystems in the southeastern United States............141

J–2. Examples of insular ecosystems outside the southeastern United States.....................142
vii

Conversion Factors

International System of Units to Inch/Pound

Multiply By To obtain

Length

centimeter (cm) 0.3937 inch (in.)

millimeter (mm) 0.03937 inch (in.)

meter (m) 3.281 foot (ft) 

kilometer (km) 0.6214 mile (mi)

kilometer (km) 0.5400 mile, nautical (nmi) 

meter (m) 1.094 yard (yd) 

Area

hectare (ha) 2.471 acre

square kilometer (km2

) 247.1 acre

hectare (ha) 0.003861 square mile (mi2

) 

square kilometer (km2

) 0.3861 square mile (mi2)

Flow rate

kilometer per hour (km/h) 0.6214 mile per hour (mi/h)

Inch/Pound to International System of Units

Multiply By To obtain

Length

foot (ft) 0.3048 meter (m)

mile (mi) 1.609 kilometer (km)

Area

acre 4,047 square meter (m2

)

acre 0.4047 hectare (ha)

acre 0.4047 square hectometer (hm2

) 

acre 0.004047 square kilometer (km2

)

square mile (mi2

) 259.0 hectare (ha)

square mile (mi2

) 2.590 square kilometer (km2) 

Flow rate

cubic foot per second (ft3

/s) 0.02832 cubic meter per second (m3/s)

Temperature in degrees Celsius (°C) may be converted to degrees Fahrenheit (°F) as follows:

°F =(1.8×°C)+ 32

Horizontal coordinate information is referenced to the World Geodetic System of 1984 (WGS 84).

Abbreviations

ITESC International Terrestrial Ecological Systems Classification 

USNVC U.S. National Vegetation Classification
viii

Acknowledgments

This literature synthesis was funded in part by the U.S. Department of the Interior Southeast 

Climate Science Center. The authors wish to acknowledge the guidance and collaboration 

of Jennifer Costanza and Jaime Collazo at North Carolina State University and Milo Pyne 

at NatureServe. 

The authors are grateful to the following colleagues for providing technical reviews of the draft 

report chapters and tables: Mason Brock,1

 Jennifer Costanza,2 Diane DeSteven,3 Timothy Diehl,4

Glenn Guntenspergen,4

 Benjamin Miller,4 Donald Phillips,5 Milo Pyne,6 Devin Rodgers,1

Joel Snodgrass,7

 Jeffrey Walck,8 Thomas Wentworth,2 Peter White,9 Thomas Wieboldt,7

Susan Wiser,10 Theo Witsell,11 and James Vanderhorst.12 The authors also wish to thank 

James Petersen,4

 Caitlin Elam,13 and Thomas Doyle,4 for providing colleague reviews of 

the compiled draft report. 

1

Austin Peay State University.

2

 North Carolina State University.

3

U.S. Forest Service.

4

U.S. Geological Survey.

5

U.S. Environmental Protection Agency.

6

 NatureServe.

7

 Virginia Polytechnic Institute and State University.

8

Middle Tennessee State University.

9

University of North Carolina.

10Landcare Research (New Zealand).

11Arkansas Natural Heritage Commission.

12West Virginia Division of Natural Resources.

13Tennessee Department of Environment and Conservation.
Insular Ecosystems of the Southeastern United States: 

A Regional Synthesis to Support Biodiversity Conservation 

in a Changing Climate

By Jennifer M. Cartwright and William J. Wolfe

Abstract

In the southeastern United States, insular ecosystems—such as rock outcrops, depression wetlands, high-elevation 

balds, flood-scoured riparian corridors, and insular prairies and barrens—occupy a small fraction of land area but 

constitute an important source of regional and global biodiversity, including concentrations of rare and endemic plant 

taxa. Maintenance of this biodiversity depends upon regimes of abiotic stress and disturbance, incorporating factors such 

as soil surface temperature, widely fluctuating hydrologic conditions, fires, flood scouring, and episodic droughts that may 

be subject to alteration by climate change. Over several decades, numerous localized, site-level investigations have yielded 

important information about the floristics, physical environments, and ecological dynamics of these insular ecosystems; 

however, the literature from these investigations has generally remained fragmented. This report consists of literature 

syntheses for eight categories of insular ecosystems of the southeastern United States, concerning (1) physical geography, 

(2) ecological determinants of community structures including vegetation dynamics and regimes of abiotic stress and 

disturbance, (3) contributions to regional and global biodiversity, (4) historical and current anthropogenic threats and 

conservation approaches, and (5) key knowledge gaps relevant to conservation, particularly in terms of climate-change 

effects on biodiversity. This regional synthesis was undertaken to discern patterns across ecosystems, identify knowledge 

gaps, and lay the groundwork for future analyses of climate-change vulnerability. Findings from this synthesis indicate 

that, despite their importance to regional and global biodiversity, insular ecosystems of the southeastern United States have 

been subjected to a variety of direct and indirect human alterations. In many cases, important questions remain concerning 

key determinants of ecosystem function. In particular, few empirical investigations in these ecosystems have focused on 

possible climate-change effects, despite the well-documented ecological effects of climate change at a global level. Long-term 

management of these ecosystems could benefit from increased scientific effort to characterize and quantify the linkages 

between changing environmental conditions and the ecological processes that sustain biodiversity.

Chapter A. Introduction

Although these insular ecosystems are critically 

important to the natural heritage of the southeastern 

United States (Stein and others, 2000; Noss, 2013), they 

have also been vulnerable to damage and destruction from 

many types of human activities (Noss and others, 1995). 

In some cases, insular ecosystems have been the subjects 

of extensive botanical inventories and localized ecological 

investigations; however, regional and conceptual 

syntheses are less common. Effective conservation and 

management of these ecosystems—especially in the 

face of climate-change—will likely require improved 

understanding of the key drivers of ecosystem dynamics, 

their biotic and abiotic interactions, and their sensitivities 

to human alteration.

The southeastern United States is rich in biodiversity 

(Noss, 2013) and contains multiple “hotspots” of rarityweighted richness (bold terms can be found in glossary) for 

globally rare taxa (Stein and others, 2000; NatureServe, 2013). 

In part, this is because of concentrations of rare and endemic

plant taxa (Marcinko, 2007), which are clustered into several 

“centers of endemism” including the southern Appalachian 

Mountains, the Nashville (Central) Basin of Tennessee and 

Alabama, and the Mid-Atlantic Coastal Plain (see maps in 

Estill and Cruzan, 2001; Sorrie and Weakley, 2001). 

Across the southeastern United States, rare-species 

biodiversity is commonly associated with clusters of “habitat 

islands” (Collins and others, 2001; Edwards and Weakley, 

2001) that form regional “archipelagos” (Loehle, 2006). 
4 Insular Ecosystems of the Southeastern United States

Background

Recognition of the ecological importance of islands dates 

back to the work of Darwin in the Galapagos Archipelago and 

that of Wallace in the Malay Archipelago (Simberloff, 1974). 

MacArthur and Wilson (1963, 1967) pioneered the study 

of island biogeography, and soon afterward the conceptual 

models and analytic approaches derived from studies of 

oceanic islands were adapted to metaphorical islands such as 

mountain tops (Brown, 1971; White and others, 1984) and 

lakes (Keddy, 1976; Browne, 1981). The ecological definition 

of “island” was expanded to include any “patch of suitable 

habitat surrounded by unfavorable environment that limits 

the dispersal of individuals” (Brown, 1978). 

More recently, general patterns have been discerned 

linking biodiversity-rich insular ecosystems to localized 

aberrations in geologic, topographic, edaphic, or geomorphic 

conditions (Collins and others, 2001), termed “landscape 

anomalies” by Kelso and others (2001). Spatial heterogeneity 

in these conditions creates geodiversity (Gray, 2013), a landscape quality that may be important to conserving biodiversity 

as climatic conditions change (Anderson and Ferree, 2010; 

Beier and Brost, 2010; Schloss and others, 2011; Groves and 

others, 2012). Indeed, geodiversity conservation has been 

increasingly recognized for its importance in planning for 

climate-change mitigation of biodiversity losses (Anderson 

and others, 2015; Comer and others, 2015). Thus, scientifically based conservation of insular ecosystems may prove 

important not only to protecting current biodiversity but also 

to improving regional resiliency to climate change.

Purpose and Scope

This report presents a state-of-the-science review of 

the literature on selected ecosystems of the southeastern 

United States that meet five key criteria for ecological 

insularity:

1. Geographic discreteness, meaning spatial isolation 

from other occurrences of the same ecosystem type, 

2. Distinctive geographic distributions reflecting highly 

specific geologic, edaphic, geomorphic, and (or) 

topographic controls, 

3. Biogeographic endemism1 and (or) disjunction for 

vascular plant taxa,

4. Abiotic conditions and (or) disturbance regimes that 

produce stressful conditions for plant growth, and 

1

Anderson (1994) differentiated “habitat endemism” (confined to a 

particular habitat that may be geographically widespread) from “geographic 

endemism” (geographically limited to an area or region). The insular ecosystems reviewed here typically support plant taxa that are narrowly endemic 

based on habitat and display varying degrees of geographic endemism, 

depending on the species and geographic scale considered.

5. Sharply defined boundaries with steep environmental 

and ecological gradients to the surrounding landscape.

Rather than attempt an exhaustive review of all 

ecosystems meeting this definition within the southeastern 

United States2, this report focuses on eight categories of insular 

ecosystems for which sufficient scientific literature exists to 

enable a synthesis across site-level investigations:

1. Granite outcrops of the Piedmont region,

2. Limestone cedar glades,

3. Xeric limestone prairies,

4. Mid-Appalachian shale barrens,

5. High-elevation outcrops and balds of the 

southern Appalachians,

6. Carolina bays,

7. Karst-depression wetlands, and

8. Riverscour ecosystems.

The geographic focus of this selection was the area east 

of the Mississippi River and south of the Ohio and Potomac 

Rivers, although some insular ecosystems have occurrences 

beyond these boundaries. Peninsular Florida was excluded 

from this review because this region contains many insular 

ecosystems and arguably deserves its own regional synthesis. 

Approach

The approach to assessing the state-of-the-science for 

the selected insular ecosystems took the form of literature 

syntheses addressing the physical geography, abiotic conditions, biodiversity contributions, plant community types, 

vegetation dynamics, and documented anthropogenic threats 

to each ecosystem. The body of this report is organized in eight 

chapters (B through I), each focused on one of the selected 

insular ecosystems. These eight chapters:

1. Summarize basic information concerning geographic 

distribution3, geologic and topographic context, and 

physical environmental characteristics,

2

Additional ecosystems in the southeastern United States meeting this 

definition are listed in table J–1.

3

A map is presented for each insular ecosystem, showing the general 

geographic distribution of ecosystem occurrence or, where appropriate, 

selected sites where previous investigations have been performed. Exhaustive 

depiction of all ecosystem occurrences was outside the scope of this study 

and would likely be impracticable for most ecosystems due to incompleteness of regional surveys. Similarly, showing map locations of all place names 

mentioned in the text would have been prohibitively complex at the map 

scales used in this report; therefore, the maps included here are for general 

illustrative purposes only.
Chapter A 5

2. Provide a synthesis of known ecological drivers and 

determinants of community structures, including 

factors regulating spatial distributions of taxa and 

successional dynamics,

3. Discuss prominent contributions to regional and 

global biodiversity, especially in terms of endemic 

plant taxa, disjunct plant populations, and provision 

of faunal habitat,

4. Describe historical and current threats to ecosystems 

as well as established conservation approaches to 

mitigate these threats, and 

5. Identify key areas where further research is needed to 

help inform management and conservation of these 

ecosystems, particularly in terms of climate-change 

effects on biodiversity.

Particular focus was given to the distinctive regimes of 

abiotic stress factors (for example, thin soils, widely fluctuating hydrology, low soil pH, or seasonally high soil surface 

temperatures) and disturbance events (for example, fires, 

droughts, and episodic scouring from high-energy floods) 

that characterize each ecosystem. Illustrative examples are 

included of how these abiotic stresses and disturbance regimes 

are reflected in specialized adaptations promoting stress 

tolerance in the plant taxa that are endemic to, or characteristic 

of, each ecosystem. Where applicable, the report discusses 

the mechanisms by which stress and disturbance regimes help 

shape successional dynamics and spatial vegetation patterns, 

for example by regulating rates of woody encroachment into 

herbaceous communities or by producing species turnover 

along abiotic gradients. For each ecosystem, existing conservation practices are described along with human activities and 

resulting processes that have been documented in the literature 

as prominent threats to ecosystem integrity. 

Additionally, this report identifies key knowledge 

gaps concerning fundamental ecological questions for each 

ecosystem. Documentation of basic ecological characteristics 

is uneven across ecosystems in the published literature. Some 

of these systems, such as limestone cedar glades (Quarterman, 

1950; Quarterman and others, 1993; Baskin and others, 

2007), Appalachian shale barrens (Platt, 1951; Keener, 1983; 

Norris and Sullivan, 2002), and Carolina bays (Ross, 1987; 

Sharitz, 2003) have received attention for decades from 

botanists and geographers and are the subjects of relatively 

recent regional reviews. Others, such as karst-depression 

wetlands (Wolfe, 1996; Buhlmann and others, 1999) and 

riverscour ecosystems (Murdock and others, 2007; Wolfe and 

others, 2007) are known primarily from individual site-level 

studies across a number of discrete occurrences in several 

physiographic provinces, the findings of which have not 

previously been synthesized at a regional level. To expand the 

knowledge base from which to derive general patterns and 

pose new questions, this approach included reviews not only 

of easily accessible publications such as peer-reviewed journal 

articles but also of government reports, academic theses, and 

Internet-inaccessible publications. In cases where inconsistencies in terminology and nomenclature have created confusion 

or ambiguity in the ecological literature, this report attempts 

to articulate these inconsistencies and clarify the definition 

of terms. In cases where disagreements have existed among 

researchers concerning important ecological questions, the 

conflicting viewpoints are presented, with particular emphasis 

on those supported by findings from rigorous empirical 

research. Scientific names of taxa are in accordance with those 

used by NatureServe (2015) unless otherwise noted.4

 Names 

of ecological systems and plant associations generally follow 

the International Terrestrial Ecological Systems Classification 

(ITESC) and the U.S. National Vegetation Classification 

(USNVC), unless otherwise attributed (Jennings and others, 

2009; Franklin and others, 2012). Conservation status rankings 

for associations and individual taxa are in accordance with 

NatureServe (2015) and are explained in the glossary.

This regional synthesis for insular ecosystems of 

the southeastern United States addresses one of the six 

U.S. Geological Survey, 2007 science strategy goals, “Understanding ecosystems and predicting ecosystem change” 

(U.S. Geological Survey, 2007). Particular attention was given 

to the ecological factors and forces that may prove sensitive 

to climate change. Effective management of these ecosystems 

in the face of increasing regional temperatures and changing 

precipitation patterns requires improved understanding 

in a number of areas, including the abiotic regulation of 

community structure, drivers of and constraints on succession, 

species-level physiological tolerance thresholds, and interspecific relationships. In general, these interactions and processes 

are inadequately understood for purposes of climate-change 

vulnerability assessment in the reviewed ecosystems. As such, 

an attempt was made to frame relevant questions for future 

research and, where warranted based on sufficient scientific 

documentation of climate-regulated ecosystem dynamics, to 

propose plausible and testable hypotheses concerning possible 

ecological effects of climate change. Because climate-change 

projections vary between different regionally downscaled 

climate models for the southeastern United States (Chen and 

others, 2003; Mearns and others, 2003; Mitchell and others, 

2014), these hypotheses involved consideration of general 

and qualitative climatic changes rather than quantitative 

projections. Variability between models notwithstanding, 

climate change in the southeastern United States is likely 

to involve spatially and temporally variable temperature 

increases combined with changes in the timing, frequency, 

and magnitude of precipitation, possibly including increased 

storm intensity and increased drought severity (Karl and 

Knight, 1998; Chen and others, 2003; Mearns and others, 

2003; Anchukaitis and others, 2006; Ibáñez and others, 2006; 

Mitchell and others, 2014). Thus, these types of changes were 

4

Recently renamed or reclassified taxa names may not appear in the updated 

form in NatureServe (2015). Naming revisions, reclassifications, synonyms, 

and basionyms are available in the International Plant Names Index (2015), 

the Integrated Taxonomic Information System (2015), and Natural Resources 

Conservation Service (2015).
6 Insular Ecosystems of the Southeastern United States

given primary consideration in this report for the formulation 

of climate-change hypotheses for insular ecosystems of the 

southeastern United States. Given the paucity of previous 

scientific investigations addressing climate-change effects in 

these ecosystems, the hypotheses presented here are necessarily speculative and should be viewed not as predictions but 

rather as starting points for future empirical work. Although 

answers are rare and questions are abundant concerning 

climate-change ecology in insular ecosystems, regional and 

conceptual syntheses constitute an important first step toward 

the formulation of scientifically grounded approaches to the 

conservation and management of these biodiverse ecosystems 

in the decades to come.

References

Anchukaitis, K.J., Evans, M.N., Kaplan, A., Vaganov, E., 

Hughes, M.K., Grissino-Mayer, H.D., and Cane, M., 2006, 

Forward modeling of regional scale tree-ring patterns in 

the southeastern United States and the recent influence 

of summer drought: Geophysical Research Letters, v. 33, 

no. 4, p. L04705.

Anderson, S., 1994, Area and endemism: The Quarterly 

Review of Biology, v. 69, no. 4, p. 451– 471.

Anderson, M., Comer, P., Beier, P., Lawler, J., Schloss, C., 

Buttrick, S., Albano, C., and Faith, D., 2015, Case studies 

of conservation plans that incorporate geodiversity: 

Conservation Biology, v. 29, no. 3, p. 680–691.

Anderson, M., and Ferree, C., 2010, Conserving the stage—

Climate change and the geophysical underpinnings of 

species diversity: PLOS ONE, v. 5, no. 7, p. e11554.

Baskin, J., Baskin, C., and Quarterman, E., 2007, Flow 

diagrams for plant succession in the Middle Tennessee 

cedar glades: Journal of the Botanical Research Institute 

of Texas, v. 1, no. 2, p. 1131–1140.

Beier, P., and Brost, B., 2010, Use of land facets to plan for 

climate change—Conserving the arenas, not the actors: 

Conservation Biology, v. 24, no. 3, p. 701–710.

Brown, J., 1971, Mammals on mountaintops—Nonequilibrium 

insular biogeography: The American Naturalist, v. 105, 

no. 945, p. 467–478.

Brown, J., 1978, The theory of insular biogeography and the 

distribution of boreal birds and mammals: Great Basin 

Naturalist Memoirs, p. 209–227.

Browne, R., 1981, Lakes as islands—Biogeographic 

distribution, turnover rates, and species composition in the 

lakes of central New York: Journal of Biogeography, v. 8, 

no. 1, p. 75–83.

Buhlmann, K., Mitchell, J., and Smith, L., 1999, Descriptive 

ecology of the Shenandoah Valley sinkhole pond system in 

Virginia: Banisteria, v. 13, p. 23–51.

Chen, M., Pollard, D., and Barron, E., 2003, Comparison of 

future climate change over North America simulated by two 

regional models: Journal of Geophysical Research, v. 108, 

no. D12, p. ACL 3–1—ACL 3–19.

Collins, B., White, P., and Imm, D., 2001, Introduction to 

ecology and management of rare plants of the southeast: 

Natural Areas Journal, v. 21, no. 1, p. 4 –11.

Comer, P., Pressey, R., Hunter, M., Schloss, C., Buttrick, S., 

Heller, N., Tirpak, J., Faith, D., Cross, M., and Shaffer, M., 

2015, Incorporating geodiversity into conservation decisions: Conservation Biology, v. 29, no. 3, p. 692–701.

Edwards, A., and Weakley, A., 2001, Population biology and 

management of rare plants in depression wetlands of the 

southeastern Coastal Plain, USA: Natural Areas Journal, 

v. 21, no. 1, p. 12–35.

Estill, J., and Cruzan, M., 2001, Phytogeography of rare 

plant species endemic to the southeastern United States: 

Castanea, v. 66, no. 1, p. 3–23.

Franklin, S., Faber-Langendoen, D., Jennings, M., 

Keeler-Wolf, T., Loucks, O., Peet, R., Roberts, D., and 

McKerrow, A., 2012, Building the United States National 

Vegetation Classification: Annali di Botanica, v. 2, p. 1–9.

Gray, M., 2013, Geodiversity—Valuing and conserving abiotic 

nature: Chichester, Wiley-Blackwell.

Groves, C., Game, E., Anderson, M., Cross, M., Enquist, C., 

Ferdaña, Z., Girvetz, E., Gondor, A., Hall, K., Higgins, J., 

Marshall, R., Popper, K., Schill, S., and Shafer, S., 2012, 

Incorporating climate change into systematic conservation planning: Biodiversity and Conservation, v. 21, no. 7, 

p. 1651–1671.

Ibáñez, I., Clark, J., Dietze, M., Feeley, K., Hersh, M., 

LaDeau, S., McBride, A., Welch, N., and Wolosin, M., 

2006, Predicting biodiversity change—Outside the climate 

envelope, beyond the species-area curve: Ecology, v. 87, 

no. 8, p. 1896–1906.

International Plant Names Index, 2015, The international 

plant names index, accessed October 5, 2015, at 

http://www.ipni.org. 

Integrated Taxonomic Information System, 2015, The 

Integrated Taxonomic Information System, accessed 

October 5, 2015, at http://www.itis.gov.

Jennings, M., Faber-Langendoen, D., Loucks, O., Peet, R., and 

Roberts, D., 2009, Standards for associations and alliances 

of the U.S. National Vegetation Classification: Ecological 

Monographs, v. 79, no. 2, p. 173–199.
Chapter A 7

Karl, T.R., and Knight, R.W., 1998, Secular trends of precipitation amount, frequency, and intensity in the United States: 

Bulletin of the American Meteorological Society, v. 79, 

no. 2, p. 231–241.

Keddy, P., 1976, Lakes as islands—The distributional ecology 

of two aquatic plants, Lemna minor L. and L. trisulca L.: 

Ecology, v. 57, no. 2, p. 353–359.

Keener, C., 1983, Distribution and biohistory of the endemic 

flora of the mid-Appalachian shale barrens: The Botanical 

Review, v. 49, no. 1, p. 65–115.

Kelso, S., Hall, C., and Maentz, G., 2001, The role of landscape anomalies in regional plant conservation, in Southwestern Rare and Endangered Plants: Proceedings of the 

Third Conference: Flagstaff, Ariz., p. 13–19.

Loehle, C., 2006, Endemic plant distributions in Eastern 

North America—Implications for conservation: Journal 

of Forestry, v. 104, no. 8, p. 415–418.

MacArthur, R., and Wilson, E., 1963, An equilibrium theory of 

insular zoogeography: Evolution, v. 17, no. 4, p. 373–387.

MacArthur, R., and Wilson, E., 1967, The theory of island 

biogeography: Princeton, N.J., Princeton University Press. 

Marcinko, S., 2007, Pattern and process in rare plant conservation—An assessment of the southeastern U.S.: Chapel 

Hill, N.C., University of North Carolina, Masters Thesis.

Mearns, L.O., Giorgi, F., Mcdaniel, L., and Shields, C., 2003, 

Climate scenarios for the southeastern U.S. based on GCM 

and regional model simulations: Climate, v. 60, p. 7–35.

Mitchell, R.J., Liu, Y., O’Brien, J.J., Elliott, K.J., Starr, G., 

Miniat, C.F., and Hiers, J.K., 2014, Future climate and fire 

interactions in the southeastern region of the United States: 

Forest Ecology and Management, v. 327, p. 316–326.

Murdock, N., Hughes, J., and Emmott, R., 2007, Understanding hydrologic links between “river prairies” and other 

threatened riparian resources of the Cumberland Plateau, 

in Selleck, J. ed., Natural resource year in review—2006, 

publication D-1859: Denver, Colo., National Park Service, 

Natural Resource Program Center, p. 45–46.

Natural Resources Conservation Service, 2015, The USDA 

plants database: Greensboro, N.C., U.S. Department of 

Agriculture, Natural Resources Conservation Service, 

National Plant Data Team, accessed September 25, 2015, 

at http://plants.usda.gov. 

NatureServe, 2013, NatureServe rarity-weighted richness 

model of critically imperiled and imperiled species

in the United States (updated in 2013), accessed 

November 15, 2015, at http://www.natureserve.org/

conservation-tools/natureserve-hotspots-map.

NatureServe, 2015, NatureServe Explorer—An online 

encyclopedia of life [web application], ver. 7.1, accessed 

October 7, 2015, at http://www.natureserve.org/explorer.

Norris, S., and Sullivan, R., 2002, Conservation assessment 

for the mid-Appalachian shale barrens: Elkins, W.Va., 

West Virginia Division of Natural Resources. 

Noss, R., 2013, Forgotten grasslands of the south: Washington, 

D.C., Island Press. 

Noss, R., LaRoe, E., and Scott, J., 1995, Endangered ecosystems of the United States—A preliminary assessment of

loss and degradation: Washington, D.C., U.S. Department

of the Interior. 

Platt, R., 1951, An ecological study of the mid-Appalachian 

shale barrens and of the plants endemic to them: Ecological 

Monographs, v. 21, no. 4, p. 269–300.

Quarterman, E., 1950, Major plant communities of Tennessee 

cedar glades: Ecology, v. 31, no. 2, p. 234–254.

Quarterman, E., Burbanck, M., and Shure, D., 1993, Rock 

outcrop communities—Limestone, sandstone, and granite, 

in Martin, W., Boyce, S., and Echternacht, A. eds., 

Biodiversity of the southeastern United States—Upland 

terrestrial communities: New York, John Wiley and Sons, 

Inc., p. 35–86.

Ross, T., 1987, A comprehensive bibliography of the Carolina 

bays literature: The Journal of the Elisha Mitchell Scientific 

Society, v. 103, no. 1, p. 28–42.

Schloss, C., Lawler, J., Larson, E., Papendick, H., Case, M., 

Evans, D., DeLap, J., Langdon, J., Hall, S., and McRae, B.,

2011, Systematic conservation planning in the face of 

climate change—Bet-hedging on the Columbia Plateau: 

PLOS ONE, v. 6, no. 12, p. e28788.

Sharitz, R., 2003, Carolina bay wetlands—Unique habitats 

of the southeastern United States: Wetlands, v. 23, no. 3, 

p. 550–562.

Simberloff, D., 1974, Equilibrium theory of island biogeography: Annual Review of Ecology and Systematics, 

v. 5, p. 161–182.

Sorrie, B., and Weakley, A., 2001, Coastal Plain vascular plant 

endemics—Phytogeographic patterns: Castanea, v. 66, 

no. 1, p. 50–82.

Stein, B., Kutner, L., and Adams, J., eds., 2000, Precious 

heritage—The status of biodiversity in the United States: 

New York, N.Y., Oxford University Press.

U.S. Geological Survey, 2007, Facing tomorrow’s challenges—

U.S. Geological Survey Science in the decade 2007–2017: 

U.S. Geological Survey Circular 1309, 67 p.
8 Insular Ecosystems of the Southeastern United States

White, P., Miller, R., and Ramseur, G., 1984, The speciesarea relationship of the southern Appalachian high 

peaks—Vascular plant richness and rare plant distributions: 

Castanea, v. 49, no. 2, p. 47–61.

Wolfe, W.J., 1996, Hydrology and tree-distribution patterns 

of karst wetlands at Arnold Engineering Development 

Center, Tennessee: U.S. Geological Survey Water-Resouces 

Investigations Report 96–4277, 46 p.

Wolfe, W.J., Fitch, K.C., and Ladd, D.E., 2007, Alluvial 

bars of the Obed Wild and Scenic River, Tennessee: 

U.S. Geological Survey Scientific Investigations 

Map 2007–2972, 1 sheet.
Chapter B. Granite Outcrops of the Piedmont

Geographic Range

Granite outcrops are widely distributed throughout the 

Piedmont Physiographic Province, from eastern Alabama to 

southern Virginia (fig. B–2). The Piedmont Physiographic 

Province slopes from the Blue Ridge Mountains in the east 

toward the Fall Line, which is the boundary with the Atlantic 

Coastal Plain. Granite outcrops are concentrated in the zone 

of the Piedmont nearest the Fall Line, a zone approximately 

120 miles (about 195 kilometers) wide that extends about 

700 miles (about 1,130 kilometers) from Alabama to Virginia 

(Oosting and Anderson, 1939; McVaugh, 1943). Within this 

zone, granite outcrops range in size from less than 100 square 

feet to hundreds of acres (a few square meters to hundreds 

of hectares) (McVaugh, 1943; Tate, 2005). The total area of 

Piedmont granite outcrops is estimated at close to 8,000 acres 

(about 3,200 hectares), with 6,000 acres (about 2,500 hectares) 

Introduction

Within the Piedmont Physiographic Province (Fenneman, 

1938), differential weathering has produced outcroppings of 

relatively resistant granite schists and gneisses surrounded 

by deeper soils from more easily weathered parent material 

(Shure, 1999). Piedmont granite outcrops are thus “terrestrial 

islands of rock” surrounded by forest (Caspary and Affolter, 

2012) that contain a diverse array of microhabitats and 

associated plant communities subject to a range of hydrologic 

conditions, from generally xeric to seasonally variable to 

semipermanently inundated (Quarterman and others, 1993). 

These outcrops are particularly noteworthy for their high 

numbers of rare and endemic vascular plant species (Shure, 

1999; Loehle, 2006; Marcinko, 2007) and for the presence of 

soil patches within topographic depressions of the rock surface 

that function as “islands within islands” (fig. B–1).

Figure B–1. Camp Meeting Rock Preserve, Heard County, Georgia. 

Photograph by Alan M. Cressler, U.S. Geological Survey.
10 Insular Ecosystems of the Southeastern United States

in Georgia where the largest outcrops are present (McVaugh, 

1943; Tate, 2005). Although more attention has been paid to 

granite outcrops along the Fall Line, Wyatt and Fowler (1977) 

pointed out that, at least in North Carolina, two other bands of 

granite outcrops run roughly parallel to the Fall Line zone of 

outcrops but considerably to the west: one along the center of 

the Piedmont, the other just east of the Blue Ridge Mountains. 

These three bands of outcrops generally correspond to the 

three main granite belts within the Piedmont, all oriented in a 

northeast-to-southwest direction. The western outcrops tend 

to be more steeply sloping and have more fissures and crevices 

than the outcrops along the Fall Line (Wyatt and Fowler, 1977). 

Granite outcrops also are present in a number of other regions 

of North America, including the Wichita Mountains of Oklahoma, the Sierra Nevada of California, and the Central Mineral 

Region of Texas, with the Texas outcrops showing similarities 

in life-form spectra to those in the Piedmont but with substantially lower rates of endemism (Walters and Wyatt, 1982). 

Figure B–2. Geographic distribution of Piedmont granite outcrops.

Figure B-2

85° 80° 75°

40°

35°

30°

25°

Base modifed from U.S. Department of Commerce, 

Census Bureau, U.S. States, 2015

Mercator Auxiliary Sphere Projection

Horizontal datum is World Geodedic System of 1984

EXPLANATION

Piedmont Physiographic Province—

 From Fenneman and Johnson (1946)

Southern Piedmont Granite Flatrock

 and Outcrop Ecological System—

 Modified from U.S. Geological 

 Survey (2011)

Fall Line—From Fenneman and

 Johnson (1946)

Selected occurrence of Piedmont

 granite outcrop—Modified from

 Shure (1999)

0 100 200 300 KILOMETERS

0 100 200 300 MILES

TENNESSEE

ALABAMA GEORGIA

SOUTH

CAROLINA

NORTH

CAROLINA

WEST

VIRGINIA

OHIO

PENNSYLVANIA

DELAWARE

NEW

JERSEY

INDIANA

VIRGINIA

FLORIDA

KENTUCKY

MARYLAND
Chapter B 11

Physical Geography

At least two distinct types of granite outcrops are 

present in the Piedmont: relatively flat outcrops at the same 

general elevation as the surrounding landscape and domed 

monadnocks (Stone Mountain in DeKalb County, Georgia, 

being the most well-known example) which rise as high as 

700 feet (about 215 meters) above the surrounding rolling 

terrain (Keever and others, 1951; Burbanck and Platt, 1964; 

Shure, 1999; Belden and others, 2003). The exposed rock may 

be true granite or may contain porphyritic granite, gneiss, or 

granodiorite, and is generally high in silica and aluminum 

and low in iron and magnesium (McVaugh, 1943; Belden and 

others, 2003; Tate, 2005). In comparison to other rock outcrop 

types, granite outcrops are relatively smooth with few cracks, 

crevices, or fractures, and they lack the surface layer of rock 

fragments that is typical of limestone cedar glades (Oosting 

and Anderson, 1937; McVaugh, 1943). A characteristic feature 

of flat and domed granite outcrops is the presence of shallow 

depressions (sometimes called “solution pits,” fig. B–3) that 

may be seasonally ponded if the rock rim is continuous but 

otherwise are xeric (Oosting and Anderson, 1939; Burbanck 

and Platt, 1964; Shure, 1999). The formation of solution pits is 

primarily driven by physical weathering—including scouring 

by wind and water—and by the accumulation of organic and 

inorganic deposits rather than by biochemical weathering from 

lichens or mosses (McVaugh, 1943; Burbanck and Platt, 1964). 

Stress and Disturbance Regimes

The primary components of the stress regime in granite 

outcrops stem from the paucity—and in many places the 

absence—of soil. Even within granite outcrop depressions 

in the more advanced stages of succession, soil thickness is 

markedly less than in the surrounding woodland and forest. 

Canopy cover is virtually nonexistent because so few trees 

reach maturity (Keever and others, 1951). As a result, the 

stress regime of many microhabitats within granite outcrops 

is characterized by high solar irradiance, widely fluctuating 

temperatures with maximums as high as 55 degrees Celsius 

(°C), limited water availability due to rapid evaporation and 

limited water storage by thin soil: a combination of factors 

producing a “microenvironmental desert” environment within 

a region with ample precipitation (Phillips, 1982; Shure, 

1999; Austin and others, 2009). Wet depressions experience 

additional abiotic stress conditions due to fluctuation between 

hydrologic extremes because the depressions become inundated 

by rain events but also become desiccated during dry periods 

(Oosting and Anderson, 1937; Keever and others, 1951).

Figure B–3. Solution pits in a granite outcrop, Davidson-Arabia Mountain Nature Preserve, Dekalb County, Georgia. 

Photograph by Alan M. Cressler, U.S. Geological Survey.
12 Insular Ecosystems of the Southeastern United States

Disturbance from fire may play some role in ecosystem 

maintenance and the inhibition of succession. Oosting and 

Anderson (1939) observed that “every outcrop has evidence of a 

fairly recent fire,” an idea supported by McVaugh (1943), Keever 

and others (1951), and Wyatt and Fowler (1977). However, 

the low biomass levels and patchy distribution of vegetation 

in granite outcrops might be expected to limit the spread and 

intensity of fire. The relative importance of fire as a disturbance 

factor maintaining ecosystem integrity, including its possible 

role in suppressing invasion by weedy species within outcrop 

ecotones, is not fully understood (Caspary and Affolter, 2012). 

On the sloping outcrops of the eastern Piedmont, vegetation is susceptible to several types of disturbance. For example, 

disturbance events can result when runoff from heavy rains 

releases and washes away vegetation mats that had formed 

over hundreds or thousands of years (Oosting and Anderson, 

1937; McVaugh, 1943; Wyatt and Fowler, 1977; Tate, 2005). 

On Rocky Face Mountain (Alexander County, North Carolina), 

Keever and others (1951) speculated that vegetation mats might 

have been dislodged and removed when logging operations 

in the early 1900s rolled logs down the mountain across areas 

of granite outcrop. Another form of disturbance is exfoliation, an erosion process in which shells of granite ranging in 

thickness from less than an inch to more than ten feet (a few 

centimeters to several meters) are loosened from the bedrock 

by freeze-thaw cycles and gradual weathering (Burbanck and 

Platt, 1964). Highly localized disturbance events can also 

result when a dead or windblown tree is uprooted because this 

uprooting typically destroys the vegetation mats in which the 

tree was growing. McVaugh (1943) and Keever and others 

(1951) noted that moisture conditions typically limit the size 

and age of trees, which are prone to being uprooted due to 

insufficient anchorage in shallow soils. Wyatt and Fowler 

(1977), Phillips (1981), and Burbanck and Phillips (1983) 

suggested that succession on granite outcrops is thus fundamentally limited by moisture and nutrient availability and that 

disturbance effects from drought, erosion, and wind generally 

preclude the achievement of a climax or steady-state seral 

stage such as that of the surrounding forest.

Community Types and Vegetation 

Dynamics

Granite outcrops harbor several distinct microhabitats, 

each with its own characteristic species assemblages and 

patterns of succession. These microhabitats include exposed 

rock surfaces, rock crevices, dry depressions, periodically 

ponded depressions, persistent pools, soil islands (accumulations of deeper soil that are sometimes called vegetation 

islands or vegetation mats), seepage areas, and ecotones to the 

surrounding forest (Oosting and Anderson, 1937; McVaugh, 

1943; Wyatt and Fowler, 1977; Shure, 1999). 

Exposed Rock Surfaces 

Much of the area of granite outcrops is exposed bedrock 

with virtually no soil development. The first pioneers of bare 

granite surfaces are invariably lichens and mosses (fig. B – 4) 

such as Staurothele diffractella, Peccania kansana, Grimmia 

laevigata, and Rhacomitrium heterostichum) (Oosting and 

Anderson, 1937; McVaugh, 1943; Keever and others, 1951; 

Wyatt and Stoneburner, 1982). Ability to colonize this harsh 

environment requires tolerance of xeric conditions, temperature fluctuations, and high levels of light intensity (Keever, 

1957). Pioneer establishment facilitates the initial stages of soil 

development by way of several interrelated processes. Mosses 

and lichens trap and accumulate mineral particles that are 

blown across rock surfaces by wind or washed across by storm 

runoff and contribute organic matter (Oosting and Anderson, 

1937; Keever and others, 1951). Through their associations 

with bacteria, lichens and bryophytes also play roles in 

soil nitrogen fixation (Snyder and Wullstein, 1973). Spatial 

distribution of species in these early successional communities 

is likely governed by complex competitive dynamics. Grimmia

mats typically continue their pioneer expansion at their 

periphery and are simultaneously invaded in their centers 

(the oldest, deepest part of the mats) by lichens (Cladonia 

leporina), other mosses (Selaginella rupestris, Polytrichum 

ohioense), and sometimes vascular plants. The presence of 

lichens often retards the growth of Grimmia (Keever, 1957; 

Shure, 1999). Keever and others (1951) noted that the surfaces 

of these mats are extremely compact and speculated that this 

surface inhibits seedling establishment, resulting in relatively 

slow succession on moss/lichen mats that form on level, bare 

rock. For example, the granite outcrop communities studied 

by Leslie and Burbanck (1979) were dominated by mats of 

Polytrichum and Selaginella with relatively few other species 

present, indicating limited or slow succession. However, 

Oosting and Anderson (1939) maintained that, given sufficient 

time, these mats will develop soil layers sufficient to support 

drought-resistant woody shrub species, although these shrubs 

tend to be stunted and slow growing. Shure (1999) suggested 

that different rates of mat formation and successional development may be geographically determined, with slower rates in 

Georgia where lichens are the primary invaders and faster rates 

in North Carolina and Virginia where mats are more successfully invaded by vascular plants. 

Rock Crevices 

Crevices range in size from hairline cracks sustaining 

only one or two individual herbaceous plants to large fissures 

supporting woody shrubs and small trees. Explorers of granite 

outcrops in the eastern Piedmont (along the Fall Line) have 

noted the scarcity of crevices and fissures and generally have 

excluded them from their discussions of community types or 

succession pathways (Oosting and Anderson, 1939; Keever 
Chapter B 13

and others, 1951; Shure, 1999). Wyatt and Fowler (1977) 

observed that crevice formation is more common on more 

steeply sloped granite outcrops of the western Piedmont near 

the Blue Ridge Mountains, but did not extensively describe 

crevices as a community or successional type.

Dry Depressions 

For depressions in which the downslope rock rim has been 

broken or partially eroded, such that rainfall is not retained 

and moisture evaporates quickly, hydrologic conditions are 

seasonally to semipermanently xeric. These depressions do 

accumulate and retain soil and minerals washed across them; 

however, this soil is typically extremely thin, dry, and hard. 

Oosting and Anderson (1937) suggested that these dry depressions “are perhaps less favorable habitats than the bare rock 

itself” because the soil layer is insufficient for most plants but 

its presence excludes the bare rock pioneers. In summer, dry 

depressions may appear completely unvegetated but in spring 

can support populations of Phemeranthus species, Portulaca 

smallii, Agrostis elliottiana, and—given sufficiently deep soil—

Diamorpha smallii (Oosting and Anderson, 1937). 

Periodically Ponded Depressions 

Where the rock rim of a depression is continuous—

allowing rainfall to be retained—and where the depression 

is shallow enough that water completely evaporates during 

dry periods, hydrologic conditions span both extremes, 

from xeric to inundated (Oosting and Anderson 1937; 

Keever and others, 1951). The few pioneer species that 

can survive these widely fluctuating hydrologic conditions 

include members of Grimmia, Cladonia, and Polytrichum

genera. Once these species become established, they 

moderate the hydrologic regime and build soil within the 

depression. This process facilitates succession similar 

to that on bare rock, but with typically thicker soil and 

more moisture availability allowing for a more diverse 

assemblage of species (Oosting and Anderson, 1939). 

The semiaquatic endemics Isoetes melanospora, Isoetes 

tegetiformans, and Amphianthus pusillus apparently can 

survive some degree of desiccation and so may be able to 

colonize some periodically ponded depressions depending 

on the duration of dryness (Shure, 1999).

Figure B–4. Moss-colonized rock surface, Davidson-Arabia Mountain Nature Preserve, Dekalb County, Georgia. Photograph 

by Alan M. Cressler, U.S. Geological Survey.
14 Insular Ecosystems of the Southeastern United States

Persistent Pools

Generally located atop granitic domes as opposed to 

flat outcrops, persistent pools are the least common of the 

depression types because few depressions are sufficiently large 

or deep to be persistently inundated (Oosting and Anderson, 

1939; Burbanck and Platt, 1964; Shure, 1999). These pools 

support the Piedmont Granite Flatrock Glade Seasonal Pool 

association, which is ranked G1. Persistent pools have been 

documented primarily in Georgia and are conspicuously 

absent in North Carolina (Wyatt and Fowler, 1977) and 

Virginia (Shure, 1999). These pools are the primary habitat for 

the three semiaquatic endemics of granite outcrops: I. melanospora, I. tegetiformans, and A. pusillus. Around the pools’ 

circumferences at the water’s edge are communities often 

dominated by Polytrichum ohioense that resemble those found 

in periodically ponded depressions (Oosting and Anderson, 

1939). These communities around the water’s edge, along with 

Eleocharis obtusa and Typha latifolia, which form clumps 

in the inundated zone, contribute to soil formation such that 

persistent pools follow succession pathways to become moist 

or marshy depressions (Oosting and Anderson, 1939).

Soil Islands 

In certain depressions, substantial accumulation of 

soil has occurred (fig. B–5). In many cases, the surrounding 

exposed rock is sloped and the lower rim of the depression 

has partially eroded away, allowing rainfall to enter and exit 

(Burbanck and Platt, 1964). This topography, combined 

with the water-retaining capacity of accumulated soil, helps 

moderate the hydrology of soil islands such that they tend not 

to experience the xeric or saturated conditions characteristic 

of periodically ponded depressions. As succession proceeds, 

soil organic matter, plant biomass, and vertical stratification 

increase (Shure and Ragsdale, 1977). 

Burbanck and Platt (1964) and Shure (1999) assigned 

soil island communities to one of four general types in order 

of increasing soil thickness and island area: Diamorpha, 

lichen-annual herb, annual-perennial herb, and herb-shrubtree. Burbanck and Phillips (1983) resampled soil islands in 

Georgia over 9- and 20-year intervals to confirm that these 

four community types do indeed represent seral stages, 

with increased soil depth corresponding to floristic changes 

for about half of the soil islands studied. Phillips (1981) 

Figure B–5. Soil island, Davidson-Arabia Mountain Nature Preserve, Dekalb County, Georgia. Photograph by Alan M. Cressler, 

U.S. Geological Survey.
Chapter B 15

demonstrated that, although community composition tends to 

progress through the seral stages in the order described here, 

tree mortality from droughts may cause reversion from the 

herb-shrub-tree seral stage to the herb-shrub stage. 

Houle (1990) demonstrated at Arabia Mountain—a 

domed monadnock east of Atlanta, Georgia—that plantspecies richness was positively associated with island 

area (in support of the species-area hypothesis of island 

biogeography) and even more strongly with soil depth (in 

support of the habitat-diversity hypothesis if the soil depth 

gradient is considered a series of distinct habitats). The 

findings of Houle (1990) and Houle and Phillips (1988) also 

indicate that plant diversity and patterns of seed abundance, 

respectively, change with advancing seral stages, which are 

in turn associated with declining abiotic stress (deeper soil, 

increased water retention and nutrient availability, and reduction in temperature fluctuation). Shure and Ragsdale (1977) 

demonstrated a peak in plant diversity at intermediate seral 

stages and stressed the importance of interspecific competition 

for nutrients and moisture as a regulator of plant community 

composition as succession proceeds. Houle and Phillips (1988, 

1989) also demonstrated that interspecific interactions play 

important roles in determining community structure of soil 

islands. Specifically, in communities at the lichen-annual seral 

stage, community structure is influenced by the presence of 

shallow-soil species (which facilitate the establishment of 

deeper-soil species) and by the presence of Diamorpha smallii

(which helps maintain lichen and moss cover). 

Although studied less extensively than plant community 

dynamics, animal community composition has also been 

shown to change over the course of succession in soil islands. 

Austin and others (2009) demonstrated that island soils at 

different seral stages harbor different community structures 

of soil nematodes. Shure and Ragsdale (1977) found that 

macroarthropod densities, biomass, and diversity increased 

as succession proceeded and that microarthropod diversity 

also increased with decreasing seasonal variability in soil 

physical conditions.

Marginal Zones and Seepage Areas 

Ecotones from bare rock to soil islands or to the 

surrounding forest contain marginal zones and associated 

areas of seepage. Such marginal zones support associations 

such as the Pine/Granitic Flatrock Border Woodland, a 

G3-ranked association. Soil inputs and losses are believed to 

be fairly balanced in these ecotones, a factor promoting relative 

stability in community structure (Oosting and Anderson, 1939; 

Shure, 1999). These ecotones are vulnerable to invasions of 

exotic species, however, such that natural disturbance from fire 

(or, in its absence, controlled burn programs) are important 

regulators of community composition (Caspary and Affolter, 

2012). On outcrops with gradual slopes, precipitation that 

percolates down through the vegetation and soil of isolated 

vegetation mats, soil islands, or the surrounding woodland and 

forest is intercepted by the granite bedrock and diverted laterally. Thus, in transitional zones to exposed bedrock downslope 

of these vegetated areas, seepage is an important contributor to 

water availability. In some cases, seepage is sufficient to support 

characteristic bog species, including Sphagnum species, 

Bryum pseudotriquetrum, and Philonotis fontana (Oosting 

and Anderson, 1939; Shure, 1999). Marginal zones adjacent 

to forest are often partially shaded, adding to their ability to 

retain moisture, and protected from the scouring action of 

heavy rain such that a layer of organic debris accumulates 

(Oosting and Anderson, 1939). A number of herbaceous and 

“weedy” species (not endemic to the outcrops) are capable of 

colonizing this habitat.

Contributions to Regional Biodiversity

Like other rock outcrops in the southeastern United 

States, Piedmont granite outcrops make important contributions to regional biodiversity (Quarterman and others, 1993). 

In particular, Piedmont granite outcrops provide habitat to 

plant and animal taxa of conservation concern, including 

endemic and biogeographically disjunct taxa (Shure, 1999; 

Marcinko, 2007). 

Endemic and Biogeographically Disjunct Taxa

Piedmont granite outcrops support a number of plant 

taxa of conservation concern (table 3–1 in appendix 3). 

Although many of these taxa are endemic (or nearly so) to 

Piedmont granite outcrops, the enumeration and identification 

of endemics has involved several types of ambiguity. Over 

decades, a number of lists have been compiled representing 

species referred to as “endemic” or else as characteristic of 

Piedmont granite outcrops (McVaugh, 1943; Burbanck and 

Platt, 1964; Shure, 1999). These lists were based on study 

areas ranging from hundreds of outcrops across several States 

to single visits to a relatively small number of outcrops. 

Although some overlap exists between lists, application of 

the terms “endemic” and “near endemic” are highly variable 

among investigators. For example, McVaugh (1943) included 

in his list of “supposedly endemic” taxa Liatris microcephala, 

which also is found on the Cumberland Plateau in sandstone 

outcrop ecosystems (Schmalzer and others, 1985) and on 

flood-scoured cobble bars along rivers (Bailey and Coe, 2001). 

Indeed, a number of taxa are generally restricted to rock 

outcrop ecosystems but also are found on sandstone or other 

substrates and have distributions that include other physiographic provinces. These taxa include Diamorpha smallii, 

Minuartia uniflora, Talinum teretifolium, and Cuscuta harperi

(Baskin and Baskin, 1988; Reinhard and Ware, 1989; 

Spaulding, 2013). Some taxa common to granite outcrops 

(including Diodia teres, T. teretifolium, Commelina erecta, 

Oenothera fruticosa, Opuntia humifusa var. humifusa, and 

Croton willdenowii) have also been documented on diabase 
16 Insular Ecosystems of the Southeastern United States

cedar glades of the Piedmont, which are habitats with considerably higher pH (LeGrand, 1988). Additionally, disagreements and ambiguity regarding taxonomic classification have 

complicated the enumeration of endemics, as has been the case 

with Isoetes taxa (Matthews and Murdy, 1969; Boom, 1982). 

These uncertainties notwithstanding, the narrow endemism of several taxa is widely accepted. These taxa include 

the semiaquatic Amphianthus pusillus and Isoetes species 

associated with persistent pools and species occupying various 

niches along gradients between bare rock and surrounding 

ecosystems, such as Sedum pusillum, Portulaca smallii, and 

Phacelia maculata (Shure, 1999). Three of these taxa are 

shown in figure B–6. Some endemic taxa belong to geographically widespread genera containing many other species (such 

as Juncus georgianus and Quercus georgiana), whereas others 

belong to monotypic genera (such as Amphianthus pusillus) 

(Braun, 1955; Hilton and Boyd, 1996). Certain genera, such as 

Isoetes, have several endemic sub-taxa with different degrees 

of geographic restriction and varying habitat characteristics 

(Matthews and Murdy, 1969). 

The biogeography of granite outcrop endemics has been 

studied in some detail. The central Piedmont of Georgia 

has been referred to as the “center of endemism” for granite 

outcrop species because it has the greatest concentration 

of these endemics within the Piedmont region, geographically coinciding with the greatest concentration of exposed 

rock surfaces (Quarterman and others, 1993; Shure, 1999; 

Tate, 2005). Wyatt and Fowler (1977) noted characteristic 

geographic differences in community composition between 

western and eastern granite outcrops in North Carolina, 

highlighted by the phenomenon of competitive replacement 

whereby a species adapted to a particular habitat in eastern 

outcrops is replaced by another species of the same genus in 

the corresponding habitat in western outcrops. For example, 

Murdy and others (1970) described the competitive replacement of Talinum mengesii by T. teretifolium, which appear to 

occupy the same niche within granite outcrop communities. 

In line with the general principles of island biogeography, 

the number of endemics present on a given granite outcrop 

is determined by factors including the size of the outcrop, 

proximity to the nearest outcrop, outcrop age, and degree 

of recent disturbance (Wyatt and Fowler, 1977). Dispersal 

capabilities of individual taxa also play an important role in 

distribution patterns of granite outcrop plants. In a genetic 

analysis of Minuartia uniflora, Wyatt and others (1992) 

noted unusually low genetic similarity between populations, 

suggesting that “pollen and seed dispersal between outcrop 

populations [of M. uniflora] is negligible” resulting in very 

low gene flow between populations that have an “island-like 

distribution” on granite outcrops across the species range. 

Houle (1990) presented concurrent evidence about overall 

species diversity for soil islands within a single outcrop—

showing a strong negative correlation between species 

richness of soil islands and the distance to the closest uphill 

island community—and suggested that the stress regime in 

early seral stages may be inhospitable to seedling germination, 

thus limiting immigration and gene flow and acting as a barrier 

to dispersal from one soil island to another. The geographic 

ranges of the aquatic endemics are particularly restricted 

because their aquatic habitat, available on some outcrops in 

Alabama, Georgia, and South Carolina, is generally not found 

farther north in the Piedmont (Wyatt and Fowler, 1977). Thus, 

distributions of taxa such as A. pusillus may be limited by 

habitat suitability and also by dispersal limitations (Hilton and 

Boyd, 1996).

Piedmont granite outcrops also provide habitat to a 

number of species with disjunct distributions and those with 

ranges that extend into the outcrops from nearby provinces. 

These outcrops allow range extensions for some Coastal Plain 

species (such as Packera tomentosa and Utricularia juncea) 

by providing moist, sandy habitats in outcrop seepage areas 

comparable to those available in the Coastal Plain, and for 

some mountain species (such as Saxifraga michauxii) for 

which populations along the eastern limit of the range are 

clustered in granite outcrops (Wyatt and Fowler, 1977). 

Habitat in Piedmont granite outcrops has also extended the 

ranges for Pleurochaete squarrosa, a moss previously thought 

to be restricted to calcareous substrates in the Interior Low 

Plateaus and Valley and Ridge Physiographic Provinces, 

and for Pleurozium schreberi, a boreal moss native to 

higher elevations in the Appalachian Mountains (Wyatt and 

Stoneburner, 1982). A few occurrences of Draba aprica in the 

Piedmont of Georgia and South Carolina are present in granite 

outcrops, disjunct from the Interior Highlands of Missouri, 

Arkansas, and Oklahoma (Quarterman and others, 1993). 

Other taxa exhibit more profound disjunctions. These include 

Cladina evansii, a type of lichen disjunct by approximately 

200 miles (about 320 kilometers) from its population center in 

the sandhills region of the Atlantic Coastal Plain (Wyatt and 

Stoneburner, 1982). A population of Pellaea wrightiana—a 

rare fern native to the southwestern United States—is found 

on Piedmont granite outcrops, representing a disjunction 

of more than 1,000 miles (more than 1,500 kilometers). 

Additionally, Piedmont granite outcrops support taxa, such 

as Lotus unifoliolatus var. helleri, that are the only eastern 

members of taxonomic groups otherwise belonging to western 

North America (McVaugh, 1943; Walters and Wyatt, 1982).

The presence of endemic and disjunct plant taxa in 

granite outcrops is related to the particular abiotic stress 

regime of outcrop environments (Quarterman and others, 

1993). Endemics generally are obligate heliophytes and 

are restricted to granite outcrops primarily due to lack of 

competitive ability in mesic environments rather than because 

of adaptation to rock substrate chemistry or genetic factors 

(Baskin and Baskin, 1988; Shure, 1999). Granite outcrop 

endemics have evolved an array of adaptations to cope with 

thin soils, xeric or highly variable hydrologic conditions, 

extreme temperature fluctuation, and intense solar irradiation. 

For example, Schoenolirion croceum has a deeply buried corm; 

Packera tomentosa has a deep root system and woody underground stem; Talinum teretifolium is a succulent that is capable 

of highly efficient water storage; and Tradescantia hirsuticaulis
Chapter B 17

Figure B–6. Characteristic plant taxa of Piedmont granite outcrops, A, Amphianthus pusillus, Davidson-Arabia 

Mountain Nature Preserve, Dekalb County, Georgia; B, Sedum pusillum, Heggie’s Rock Preserve; C, Isoetes 

tegetiformans, Heggie’s Rock Preserve, Columbia County, Georgia. Photographs by Alan M. Cressler, 

U.S. Geological Survey.

A

C

B
18 Insular Ecosystems of the Southeastern United States

is covered with light-colored hairs that reflect light and limit 

water loss (Burbanck and Platt, 1964; Tate, 2005). The life 

form of granite outcrop taxa (the degree of exposure of their 

perennating tissues) also reflects adaptation to harsh environmental conditions. Granite outcrop species are much more 

likely to be therophytes (having their only perennating tissues 

in the seed, an adaptation that promotes tolerance to drought 

and heat) than the surrounding Piedmont flora, and in fact, 

their life-form spectra more closely approximate desert areas 

such as Death Valley, California (Phillips, 1982). Based on 

physiology, habitat, and evolved tolerance to a particular stress 

regime, granite outcrop endemics have been compared to and 

associated with species found in deserts of the southwestern 

United States and northern Mexico (McVaugh, 1943; Braun, 

1955; Tate, 2005). The three species endemic to persistent 

pools have unique adaptations to their particular stress 

regime. Isoetes melanospora and Isoetes tegetiformans use an 

unusual photosynthetic pathway similar to crassulacean acid 

metabolism photosynthesis (CAM photosynthesis) (Shure, 

1999). Amphianthus pusillus has submerged and floating 

leaves, and bears its flowers on elongated floating stems and 

also submerged at the plant base where they are cleistogamous 

(self-pollinating) (Hilton and Boyd, 1996; Shure, 1999). 

A. pusillus is a winter annual, germinating when water levels 

in pools are typically highest and completing its life cycle very 

quickly (Hilton and Boyd, 1996). Although these adaptations 

to abiotic stress are particularly noteworthy for endemic taxa, 

abundant and geographically widespread taxa also have been 

shown to display heightened drought tolerance in their granite 

outcrop populations (Chapman and Jones, 1975).

Animal Taxa of Conservation Concern

Piedmont granite outcrops provide habitat to a 

number of arthropods of regional and global conservation 

concern. Trimerotropis saxatilis (lichen grasshopper), a 

G3-ranked species apparently restricted to rock outcrop 

and glade ecosystems, is an important and characteristic 

herbivore in Piedmont granite outcrops (Duke and Crossley, 

1975). Other regionally rare arthropods present in granite 

outcrops include Pardosa lapidicina (the lichen spider) and 

Anisomorpha ferruginea5 (a walking stick) (Quarterman and 

others, 1993). Dendrocephalus lithaca (the Stone Mountain 

fairy shrimp) was described in the 1950s from depressions 

on the summit of Stone Mountain in Georgia; however, this 

branchipod has not been reported in several decades and is 

suspected to be extinct (Rogers and others, 2004). Animals 

inhabiting granite outcrops also display certain adaptations to 

heightened abiotic stress. For example, soil nematodes exhibit 

anhydrobiosis—an ametabolic, inactive state coupled with a 

coiled morphology—in granite outcrop soils where desiccation 

is a major stress factor (Austin and others, 2009).

5

Species name according to the Integrated Taxonomic Information 

System (2015).

Conservation Considerations

Conservation of Piedmont granite outcrops and the 

biodiversity they support involves several key considerations. 

These include an assessment of current and historical threats, 

an evaluation of existing conservation strategies, and an 

identification of gaps in scientific knowledge such as those 

concerning possible ecological effects of climate change.

Threats to Ecosystem Integrity

The Piedmont Physiographic Province is one of the most 

densely populated regions in the United States, resulting in 

threats to granite outcrop ecosystems—especially those on 

private land—from suburban development, outdoor recreation, and the dumping of refuse (Tate, 2005). Some granite 

outcrops have been destroyed by the creation of reservoirs 

(Wyatt and Stoneburner, 1982; Quarterman and others, 

1993). Many surviving granite outcrops have been adversely 

affected by logging operations, quarrying for granite, and 

livestock grazing (McVaugh, 1943; Keever and others, 1951; 

Burbanck and Platt, 1964; Quarterman and others, 1993; 

Shure, 1999; Tate, 2005). Granite outcrops make attractive 

recreational destinations, and thus are especially vulnerable to 

physical destruction from trampling (Wyatt and Fowler, 1977; 

Burbanck and Phillips, 1983; Shure, 1999; Belden and others, 

2003). Air pollution and acid rain have also been mentioned 

as threats to granite outcrop species, especially bryophytes 

(Quarterman and others, 1993; Tate, 2005). Woody encroachment within ecotones of granite outcrops—possibly facilitated 

by fire suppression and the establishment of exotic invasive 

species—has also been suggested as a threat to ecosystem 

integrity (Caspary and Affolter, 2012).

Conservation Strategies

Throughout the Piedmont, a number of state, Federal, 

and private natural areas and preserves have been established 

to protect granite outcrop communities (see Quarterman and 

others [1993] for a list of sites). The fact that several endemic 

species of granite outcrops are federally listed as endangered or 

threatened may provide some additional habitat protection on 

private land (Shure, 1999). Conservation approaches generally 

focus on protecting granite outcrops from physical destruction 

(such as from quarrying or recreation) and educating the 

public about their ecological importance. In a few preserves, 

prescribed fire has been used as a management tool to combat 

woody encroachment by invasive species at outcrop-forest 

ecotones (Caspary and Affolter, 2012).
Chapter B 19

Knowledge Gaps

Some aspects of granite outcrop community structure 

have been well studied, particularly the patterns of succession 

for soil island communities (Oosting and Anderson, 1939; 

McVaugh, 1943; Keever and others, 1951; Burbanck and Platt, 

1964; Wyatt and Fowler, 1977; Burbanck and Phillips, 1983; 

Houle and Phillips, 1988, 1989; Houle, 1990). Little is known 

about other important dynamics of granite outcrop community structure, however, including plant relationships with 

arthropod herbivores and pollinators (Quarterman and others, 

1993). In some cases, the taxonomic relationships among 

granite outcrop species and between them and the surrounding 

vegetation need clarification (Tate, 2005). Additionally, the 

mechanisms and resilience capacities of plant species in 

response to climatic stressors have not been fully explored, 

such that the effects of changing abiotic stress on particular 

taxa of conservation concern cannot currently be quantified 

(Shure, 1999). Research on ecological effects of abiotic stress 

factors is needed. For example, droughts could be used to 

examine differential effects on endemic versus nonendemic 

taxa or among plants occupying different microenvironments 

within granite outcrops (Amyx, 2015).

Possible Ecological Effects from Climate Change

Research is generally lacking concerning the effects of 

climate change on Piedmont granite outcrop species. Such 

effects could plausibly include changes to hydrologic regimes, 

temperature changes, and interactive effects with known 

biotic factors such as invasive species. Ibáñez and others 

(2006) predicted that the Piedmont will experience increased 

aridity from climate change, possibly resulting in a dramatic 

restructuring of forest communities in the region. If this were 

to occur, the different community types sustained by various 

microhabitats within granite outcrops would likely be affected 

in different ways. Microhabitats that have historically been 

dry to xeric (including bare rock surfaces and dry depressions) 

would presumably experience stress regime intensification, 

with unknown consequences for their associated rare taxa 

and succession dynamics. Zones with widely fluctuating 

hydrologic conditions, such as periodically wet depressions, 

would likely experience reductions in hydroperiod that 

could lead to restructuring of plant communities. Persistent 

pools might shift from being semipermanently inundated 

to having increased frequency of complete drying, which 

would be expected to degrade habitat for the rare semiaquatic 

endemics Isoetes melanospora, I. tegetiformans, and 

Amphianthus pusillus. Davenport (2007) highlighted the 

importance of precipitation seasonality and suggested that 

A. pusillus “would survive as long as [its] life cycles coincide 

with available spring rainfall.” Microhabitats such as soil 

islands and seepage areas with hydrologic conditions that 

have historically been relatively mesic might shift toward 

drier conditions with associated restructuring of community 

composition. Hypotheses such as these are necessarily 

speculative in the absence of large-scale empirical examinations of climate-change-driven effects on Piedmont granite 

outcrop ecology. The exact nature of community restructuring 

in these outcrop communities and the implications for rare 

plant conservation are impossible to predict with any certainty. 

An additional unknown is how climate change might affect 

disturbance regimes on granite outcrops, such as fire, tree falls, 

exfoliation enabled by freeze-thaw cycles, and the dislodging 

of vegetation mats during storm events. 

As abiotic stress regimes change, the microhabitat 

requirements and dispersal limitations of many rare taxa 

associated with Piedmont granite outcrops may restrict their 

ability to migrate to new geographic areas, such that simple 

climate-envelope models based on poleward shifts in suitable 

climate will be inadequate to describe the influences of climate 

change on the geographic distributions of these taxa (Noss and 

others, 1995; Ibáñez and others, 2006). Given such barriers 

to range-shifting, adaptations such as phenophase change 

might be expected to occur as granite outcrop taxa cope 

with seasonal shifts in temperature and precipitation. One 

possible approach for detecting such effects could include the 

resampling of well-studied communities and comparison to 

earlier published data. For example, Burbanck and Platt (1964) 

produced a chart of seasonal timing for various granite outcrop 

species indicating months of vegetative growth, flowering, and 

fruiting. A resampling of those outcrops similarly focused on 

these phenological indicators might demonstrate whether any 

changes have occurred between the 1960s and the present. 

Some recent discussions of biodiversity conservation in 

the face of rapid climate change have focused on preserving 

regional diversity of geologic, edaphic, and topographic 

conditions, an approach sometimes referred to as “conserving 

the geophysical stage” (Game and others, 2010; Anderson and 

Ferree, 2010). Piedmont granite outcrops exist in stark contrast 

to the surrounding landscape and provide regionally unique 

environments for rare taxa, such that their conservation would 

certainly be an important component to any regional climatechange-mitigation strategies seeking to maintain such diversity 

of the physical environment.
20 Insular Ecosystems of the Southeastern United States

References

Amyx, H., 2015, Helianthus porteri (Porter’s sunflower) 

survives drought at Rocky Face Mountain, accessed 

November 17, 2015, at https://bwwellsassociation.wordpress.com/2015/09/30/helianthus-porteri-porters-sunflowersurvives-drought-at-rocky-face-mountain/.

Anderson, M., and Ferree, C., 2010, Conserving the stage—

Climate change and the geophysical underpinnings of 

species diversity: PLOS ONE, v. 5, no. 7, p. e11554.

Austin, E., Semmens, K., Parsons, C., and Treonis, A., 2009, 

Granite rock outcrops—An extreme environment for soil 

nematodes?: Journal of Nematology, v. 41, no. 1, p. 84–91.

Bailey, C., and Coe, F., 2001, The vascular flora of the 

riparian zones of the Clear Fork River and the New River 

in the Big South Fork National River and Recreation Area 

(BSFNRRA): Castanea, v. 66, no. 3, p. 252–274.

Baskin, J., and Baskin, C., 1988, Endemism in rock outcrop 

plant communities of unglaciated eastern United States—

An evaluation of the roles of the edaphic, genetic and light 

factors: Journal of Biogeography, v. 15, no. 5, p. 829–840.

Belden, A., Chazal, A., and Hobson, C., 2003, A natural 

heritage inventory of fourteen headwater sites in the 

Dragon Run Watershed: Richmond, Va., Virginia 

Department of Conservation and Recreation.

Boom, B., 1982, Synopsis of Isoetes in the southeastern 

United States: Castanea, v. 47, no. 1, p. 38–59.

Braun, E., 1955, The phytogeography of unglaciated eastern 

United States and its interpretation: The Botanical Review, 

v. 21, no. 6, p. 297–375.

Burbanck, M., and Phillips, D., 1983, Evidence of plant 

succession on granite outcrops of the Georgia Piedmont: 

American Midland Naturalist, v. 109, no. 1, p. 94–104.

Burbanck, M., and Platt, R., 1964, Granite outcrop communities of the Piedmont Plateau in Georgia: Ecology, v. 45, 

no. 2, p. 292–306.

Caspary, M., and Affolter, J., 2012, Using prescribed burning 

to restore granite rock outcrop ecotones in the Piedmont 

of the southeastern United States: Ecological Restoration, 

v. 30, no. 3, p. 228–236.

Chapman, R., and Jones, S., 1975, Ecotypic differentiation in 

Andropogon virginicus (Gramineae): Bulletin of the Torrey 

Botanical Club, v. 102, no. 4, p. 166–171.

Davenport, L., 2007, Climate change and its potential effects 

on Alabama’s plant life: Birmingham, Ala., Samford 

University, Vulcan Materials Center for Environmental 

Stewardship and Education.

Duke, K., and Crossley, D., 1975, Population energetics and 

ecology of the rock grasshopper, Trimerotropis saxatilis: 

Ecology, v. 56, no. 5, p. 1106–1117.

Fenneman, N., 1938, Physiography of the Eastern United States:

New York, McGraw-Hill.

Fenneman, N., and Johnson, D., 1946, Physical divisions of 

the United States: U.S. Geological Survey map prepared 

in cooperation with the Physiographic Commission, scale 

1:7,000,000.

Game, E., Groves, C., Andersen, M., Cross, M., Enquist, C., 

Ferdaña, Z., Girvetz, E., Gondor, A., Hall, K., Higgins, J., 

Marshall, R., Popper, K., Schill, S., and Shafer, S., 2010, 

Incorporating climate change adaptation into regional 

conservation assessments: Arlington, Va., The Nature 

Conservancy.

Hilton, J., and Boyd, R., 1996, Microhabitat requirements 

and seed/microsite limitation of the rare granite outcrop 

endemic Amphianthus pusillus (Scrophulariaceae): Bulletin 

of the Torrey Botanical Club, v. 123, no. 3, p. 189–196.

Houle, G., 1990, Species-area relationship during primary 

succession in granite outcrop plant communities: American 

Journal of Botany, v. 77, no. 11, p. 1433–1439.

Houle, G., and Phillips, D., 1988, The soil seed bank of granite 

outcrop plant communities: Oikos, v. 52, no. 1, p. 87–93.

Houle, G., and Phillips, D., 1989, Seed availability and biotic 

interactions in granite outcrop plant communities: Ecology, 

v. 70, no. 5, p. 1307–1316.

Ibáñez, I., Clark, J., Dietze, M., Feeley, K., Hersh, M., 

LaDeau, S., McBride, A., Welch, N., and Wolosin, M., 

2006, Predicting biodiversity change—Outside the climate 

envelope, beyond the species-area curve: Ecology, v. 87, 

no. 8, p. 1896–1906.

Integrated Taxonomic Information System, 2015, The 

Integrated Taxonomic Information System, accessed 

October 5, 2015, at http://www.itis.gov.

Keever, C., 1957, Establishment of Grimmia laevigata on 

bare granite: Ecology, v. 38, no. 3, p. 422–429.

Keever, C., Oosting, H., and Anderson, L., 1951, Plant 

succession on exposed granite of Rocky Face Mountain, 

Alexander County, North Carolina: Bulletin of the Torrey 

Botanical Club, v. 78, no. 5, p. 401– 421.

LeGrand, H., 1988, Cedar glades on diabase outcrops—A 

newly described community type: Castanea, v. 53, no. 2, 

p. 168–172.

Leslie, K., and Burbanck, M., 1979, Vegetation of granitic 

outcroppings at Kennesaw Mountain, Cobb County, 

Georgia: Castanea, v. 44, no. 2, p. 80–87.
Chapter B 21

Loehle, C., 2006, Endemic plant distributions in Eastern 

North America—Implications for conservation: Journal 

of Forestry, v. 104, no. 8, p. 415–418.

Marcinko, S., 2007, Pattern and process in rare plant conservation—An assessment of the southeastern U.S.: Chapel 

Hill, N.C., University of North Carolina, Masters Thesis.

Matthews, J., and Murdy, W., 1969, A study of Isoetes

common to the granite outcrops of the southeastern Piedmont, United States: Botanical Gazette, v. 130, no. 1, 

p. 53–61.

McVaugh, R., 1943, The vegetation of the granitic flat-rocks 

of the southeastern United States: Ecological Monographs, 

v. 13, no. 2, p. 119–166.

Murdy, W., Johnson, T., and Wright, V., 1970, Competitive 

replacement of Talinum mengesii by T. teretifolium in 

granite outcrop communities of Georgia: Botanical Gazette, 

v. 131, no. 3, p. 186–192.

Noss, R., LaRoe, E., and Scott, J., 1995, Endangered ecosystems of the United States—A preliminary assessment of

loss and degradation: Washington, D.C., U.S. Department

of the Interior.

Oosting, H., and Anderson, L., 1937, The vegetation of a 

barefaced cliff in western North Carolina: Ecology, v. 18, 

no. 2, p. 280–292.

Oosting, H., and Anderson, L., 1939, Plant succession on 

granite rock in eastern North Carolina: Botanical Gazette, 

v. 100, no. 4, p. 750–768.

Phillips, D., 1981, Succession in granite outcrop shrubtree communities: American Midland Naturalist, v. 106, 

p. 313–317.

Phillips, D., 1982, Life-forms of granite outcrop plants: 

American Midland Naturalist, v. 107, no. 1, p. 206–208.

Quarterman, E., Burbanck, M., and Shure, D., 1993, Rock 

outcrop communities—Limestone, sandstone, and granite, 

in Biodiversity of the southeastern United States—Upland 

Terrestrial Communities: New York, John Wiley and Sons, 

Inc., p. 35–86.

Reinhard, R., and Ware, S., 1989, Adaptation to substrate in 

rock outcrop plants—Interior Highlands Talinum (Portulacaceae): Botanical Gazette, v. 150, no. 4, p. 449–453.

Rogers, D., Jensen, J., and Floyd, T., 2004, Eubranchipus 

stegosus n. sp.(Crustacea: Anostraca), a bizarre new species 

of fairy shrimp from the southeastern United States: Hydrobiologia, v. 528, p. 201–207.

Schmalzer, P., Patrick, T., and DeSelm, H., 1985, Vascular 

flora of the Obed Wild and Scenic River, Tennessee: 

Castanea, v. 50, no. 2, p. 71–88.

Shure, D., 1999, Granite outcrops of the southeastern 

United States, in Anderson, R., Fralish, J., and Baskin, J., 

eds., Savannas, barrens, and rock outcrop plant communities of North America: New York, Cambridge University 

Press, p. 99–118.

Shure, D., and Ragsdale, H., 1977, Patterns of primary 

succession on granite outcrop surfaces: Ecology, v. 58, 

p. 993–1006.

Snyder, J., and Wullstein, L., 1973, Nitrogen fixation on 

granite outcrop pioneer ecosystems: Bryologist, v. 76, 

no. 1, p. 196–199.

Spaulding, D., 2013, Key to the dodders (Cuscuta, Convolvulaceae) of Alabama and adjacent states: Phytoneuron, 

v. 74, p. 1–15.

Tate, S., 2005, Effects of stratification on the germination of 

six Piedmont rock outcrop species and development of a 

model Piedmont rock outcrop habitat garden: Athens, Ga., 

University of Georgia, Masters Thesis.

U.S. Geological Survey, National Gap Analysis Program, 

2011, Land cover data portal, accessed December 2015 at 

http://gapanalysis.usgs.gov/gaplandcover/data/download/.

Walters, T., and Wyatt, R., 1982, The vascular flora of granite 

outcrops in the Central Mineral Region of Texas: Bulletin of 

the Torrey Botanical Club, v. 109, no. 3, p. 344–364.

Wyatt, R., Evans, E., and Sorenson, J., 1992, The evolution 

of self-pollination in granite outcrop species of Arenaria

(Caryophyllaceae). VI. Electrophoretically detectable 

genetic variation: Systematic Botany, v. 17, no. 2, 

p. 201–209.

Wyatt, R., and Fowler, N., 1977, The vascular flora and 

vegetation of the North Carolina granite outcrops: Bulletin 

of the Torrey Botanical Club, v. 104, no. 3, p. 245–253.

Wyatt, R., and Stoneburner, A., 1982, Range extensions 

for some cryptogams from granite outcrops in Alabama: 

Bryologist, v. 85, no. 4, p. 405–409.

Chapter C. Limestone Cedar Glades

Introduction

Limestone cedar glades (fig. C–1) are edaphic climax 

ecosystems, characterized by herbaceous angiosperms and 

cryptogams growing in thin soil, with areas of exposed 

limestone bedrock, rock pavement, gravel, cobbles, and 

(or) flagstone (Quarterman and others, 1993; Baskin and 

Baskin, 2003; Baskin and others, 2007b). Because the 

perimeters of these rock outcrops are typically inhabited by 

Juniperus virginiana (eastern redcedar), they are commonly 

called cedar glades (Baskin and others, 1968). Cedar glade 

vegetation is dominated by C4

 summer annual grasses (for 

example, Sporobolus vaginiflorus) and C3 forbs, which may 

be winter or summer annuals or perennials, with sparse shrub 

and tree cover (Baskin and Baskin, 1999). In microhabitats 

of cedar glades where soil is insufficiently thick to support 

vascular plants, important community constituents are mosses, 

cyanobacteria such as Nostoc commune6

, and fruticose, 

foliose, and crustose lichens (Quarterman, 1950b; Baskin 

and others, 2007b). Some limestone cedar glades contain 

distinct zonal communities that are seasonally saturated, 

support specialized associations of hydrophytic vegetation, 

and arguably meet the standards for delineation as wetlands 

(U.S. Army Corps of Engineers, 2010; Taylor and Estes, 

2012). Based on their high degree of endemism, limestone 

cedar glades are considered to be among the most botanically 

unique ecosystems in the southeastern United States (Somers 

and others, 1986; Estill and Cruzan, 2001). 

6

Species name according to the Integrated Taxonomic Information 

System (2015).

Figure C–1. Cedar Glade Number 5, Chickamauga and 

Chattanooga National Military Park, Catoosa County, Georgia. 

Photograph by Alan M. Cressler, U.S. Geological Survey.
24 Insular Ecosystems of the Southeastern United States

Geographic Range

Limestone cedar glades are present in three physiographic 

provinces of the southeastern United States (the Interior 

Low Plateaus, Appalachian Plateaus, and Valley and Ridge; 

fig. C–2) in regions of five States (the Outer Bluegrass and 

Pennyroyal Plain regions of central Kentucky; “The Cedars” 

area in Lee County in extreme southwestern Virginia; the 

Western Valley, Nashville Basin, and Valley and Ridge regions 

of Tennessee; the Tennessee Valley, Little Mountain, Moulton 

Valley, and Sequatchie Valley regions of northern Alabama; 

and the Valley and Ridge region of extreme northwestern 

Georgia; Baskin and Baskin, 1999, 2003). The greatest 

concentration and best developed examples of limestone 

cedar glades are within the Cumberland River and Duck River 

drainage basins of the Inner Nashville Basin of Tennessee, 

Figure C–2. General geographic distribution of limestone cedar glades. 

Area of 

map below

Base modifed from U.S. Department of Commerce, 

Census Bureau, U.S. States, 2015

Mercator Auxiliary Sphere Projection

Horizontal datum is World Geodedic System of 1984

0 200 300 KILOMETERS 100

0 100 200 300 MILES

90° 85° 80° 75°

40°

35°

30°

EXPLANATION

Nashville Basin Limestone Glade and 

 Woodland Ecological System—Modified

 from U.S. Geological Survey (2011)

Inner Nashville Basin subsection

Outer Nashville Basin subsection

Interior Low Plateaus Physiographic Province

Cumberland Plateau section of Appalachian 

 Plateaus Physiographic Province

Tennessee section of Valley and Ridge

 Physiographic Province

Physiographic division—From Fenneman and Johnson

 (1946) and U.S. Department of Agriculture (2015)

Selected occurrences of limestone cedar glades—

 Modified from Quarterman and others (1993)

Figure C-2

TENNESSEE

TENNESSEE

ALABAMA

ALABAMA

GEORGIA

SOUTH

CAROLINA

NORTH CAROLINA

WEST

VIRGINIA

OHIO

PENNSYLVANIA

NEW

JERSEY

INDIANA

VIRGINIA

FLORIDA

KENTUCKY

KENTUCKY

MISSOURI DELAWARE

ARKANSAS

LOUISIANA

MISSISSIPPI

ILLINOIS

MARYLAND
Chapter C 25

corresponding to the Nashville Basin Limestone Glade and 

Woodlands ecological system according to the International 

Terrestrial Ecological Systems Classification (ITESC). This 

region also supports the greatest number of cedar glade 

endemic plant taxa, with the second-highest concentration of 

endemics found in northern Alabama (Harper, 1926; Baskin 

and others, 1995; Baskin and Baskin, 1999; Baskin and 

Baskin, 2003; Sutter and others, 2011). 

Physical Geography

Limestone cedar glades develop on rock outcrops of level 

or moderate slope (5–15 percent) at an average elevation of 

600 feet (about 180 meters) on thinly bedded limestone or 

dolomite, typically of Ordovician, Silurian, or Mississippian 

age (Harper, 1926; Quarterman, 1950a; Baskin and Baskin, 

1999). In the Inner Nashville Basin, limestone cedar glades are 

present primarily on the Talbott-Rock Outcrop and the TalbottGladeville-Barfield-Rock Outcrop soil associations on limestones of Ordovician age of the Stones River Group (Baskin 

and Baskin, 2004). In the Moulton Valley of northern Alabama, 

cedar glades are present primarily on Colbert and Talbott soil 

series on the Bangor Limestone (Baskin and others, 1995). At 

some sites, the underlying limestone bedrock is extensively 

cracked, creating a patchwork habitat of thin soil above intact 

areas of bedrock surrounded by linear zones of relatively deep 

soil in crevices that are in some places wide enough to support 

J. virginiana trees (Quarterman, 1950a, b; Baskin and others, 

2007a). Figure C–3 shows an example of vegetation patterns 

reflecting bedrock fissures in a limestone cedar glade.

Cedar glades are typically associated with lithic mollisols 

(rendolls and udolls) and alfisols, and the limestone parent 

material contributes to greater alkalinity and calcium content 

relative to surrounding forest soils (Mann and others, 1999; 

Baskin and others, 2007a). Cedar glade soils characteristically 

lack a B horizon and have been classified as stony loams 

(Quarterman, 1950b) and as flaggy silty clay loams (Baskin 

and Baskin, 1999, 2004). Various researchers have characterized the soil thickness of cedar glades using slightly different 

ranges; however, all researchers agree that, in general, cedar 

glade soils are less than 12 inches (less than 30 centimeters) 

in thickness (Quarterman, 1950a; Baskin and Baskin, 1977; 

Anderson and others, 1999; Baskin and Baskin, 2003; Baskin 

and others, 2007a). Spatial gradients in soil depth are associated with gradients in a number of ecologically relevant soil 

properties and processes. For example, increasing soil depth 

is associated with decreasing pH, increasing organic matter 

content, increasing sand-to-silt ratio (Cartwright, 2014), and 

increasing soil respiration rates (Cartwright and Hui, 2014).

Figure C–3. Grass-filled cross joint, Chickamauga and Chattanooga National Military Park, Catoosa County, Georgia. 

Photograph by Alan M. Cressler, U.S. Geological Survey.
26 Insular Ecosystems of the Southeastern United States

Stress and Disturbance Regimes

The stress regime of limestone cedar glades is characterized by thin soil, high insolation, high soil temperatures 

in summer, and widely fluctuating soil moisture availability 

(Quarterman, 1989; Baskin and others, 1995; Baskin and 

Baskin, 2003). Seasonally high temperatures at the soil surface 

are especially common in unshaded zones with very thin soil 

in glade interiors. Temperatures near the soil surface have been 

recorded as high as 43 degrees Celsius (°C) (Dubois, 1993), 

50 °C (Baskin and Baskin, 1999), and 50.5 °C (Freeman, 

1933). The combination of high temperatures and thin soils in 

summer can result in seasonally dry to xeric soil conditions. 

For example, working primarily during the growing season, 

Freeman (1933) reported soil water content levels below the 

wilting coefficient 11 times between mid-June and late October. 

Spatial heterogeneity in canopy coverage and soil thickness 

commonly produce gradients of moisture availability. Based 

on five sites described as “gravel, grass, shrub, shrub-cedar and 

cedar-hardwood,” Martin and Sharp (1983) reported a decrease 

in available soil moisture along that gradient from forest to 

open glade. Seasonal water stress in cedar glades is commonly 

reflected in vegetation composition. Some plant taxa use the C4

photosynthetic pathway (for example, Andropogon virginicus, 

Sporobolus vaginiflorus, Cyperus squarrosus, some Euphorbia

species, and some Panicum species) or the crassulacean acid 

metabolism (CAM) pathway (for example, Manfreda virginica

and Opuntia humifusa), indicating adaptation to xeric conditions (Baskin and Baskin, 1981; Eickmeier, 1986). 

Several investigations have also noted components 

of cedar glade vegetation suggestive of adaptation to wet 

or saturated soil conditions (for example, Carex species, 

Isoetes butleri, Eleocharis bifida, and Schoenolirion croceum) 

(Freeman, 1933; Rollins, 1997; Baskin and others, 2007b). 

Drainage of cedar glades occurs primarily as sheet flow, which 

often drains into subterranean cave systems (Quarterman 

and others, 1993). Drainage ability may be highly variable 

spatially, depending on location relative to drainage pathways 

such as joints in limestone strata. Zones of lateral seepage 

and topographic depressions above relatively impermeable 

bedrock are commonly associated with seasonally saturated or 

inundated conditions from winter to early spring (fig. C– 4), 

when precipitation is relatively high and evapotranspiration 

is relatively low (Taylor and Estes, 2012; Nelson and others, 

2013). Winter-annual taxa, including a number of hydrophytic 

taxa, complete their life cycles during this season. Quarterman 

(1950a) emphasized the importance of widely fluctuating soil 

water content to the stress regime of cedar glades, asserting 

that “the occurrence of both wet and dry extremes must limit 

the number of species capable of living in glade areas.” 

Figure C–4. Standing water, Flat Rock Cedar Glades and Barrens State Natural Area, Rutherford County, Tennessee. 

Photograph by Alan M. Cressler, U.S. Geological Survey.
Chapter C 27

Community Types and Vegetation 

Dynamics

Relatively high abiotic stress levels, combined with low 

biomass and low levels of competition for light, generally 

mean that plant community structure within glade interiors 

is influenced more by stress tolerance than by competitive 

displacement (Quarterman, 1950b; Walck and others, 2001; 

Norton, 2010). In particular, vegetation patterns in limestone 

cedar glades commonly reflect gradients in soil depth 

(Freeman, 1933; Quarterman, 1950b; Norton, 2010). As a 

result, vegetation density can be highly spatially variable, 

ranging from zones nearly devoid of vascular plants in the 

vicinity of exposed bedrock to thickly vegetated zones in 

areas of deeper soil (Quarterman, 1950b; Nordman, 2004). 

Shrub communities characterized by Forestiera ligustrina, 

Rhus aromatica, and small Juniperus virginiana trees 

commonly develop at glade margins and along fissures 

(Quarterman, 1989). The sub-climax or climax forest 

communities that are found on glade-forming soils of the Inner 

Nashville Basin and that surround cedar glades are typically 

dominated by J. virginiana, Fraxinus americana, Carya ovata,

and Quercus muehlenbergii (Adams and others, 2012). In 

addition to vascular plant species, cedar glades provide habitat 

to a number of fruticose and foliose lichens and to communal 

groups of cyanobacteria including Oscillatoria, Stigonema, 

Scytonema, and Nostoc species (Quarterman, 1950b; Mahr and 

Mathis, 1981; Somers and others, 1986; Quarterman, 1989).

Various researchers have used differing classification 

schemes to describe the communities present in cedar glades 

(Baskin and others, 2007b). Furthermore, the scale used for 

classification affects community classification outcomes, such 

that small-scale classification primarily reflects microhabitat 

differences, whereas larger-scale classification reflects vegetation differences between glades (Rollins, 1997). Working 

in the Inner Nashville Basin of Middle Tennessee, Somers 

and others (1986) delineated seven plant communities in 

cedar glades by using cluster analysis: Panicum capillare, 

foliose lichen, Nostoc commune7

-Sporobolus vaginiflorus, 

and Dalea gattingeri (xeric communities); and Sporobolus

vaginiflorus, Pleurochaete squarrosa, and Panicum flexilePleurochaete squarrosa-Sporobolus vaginiflorus (sub-xeric 

communities). The four xeric communities generally were 

found on soils less than 2 inches (about 5 centimeters) deep, 

whereas the three sub-xeric communities occupied deeper 

soils. This soil depth distinction between xeric and sub-xeric 

communities corresponds roughly to Quarterman’s (1973) 

Zone I (gravel glades) having soil depth less than about 

2 inches (5 centimeters) and Zone II (grass glades) having soil 

depth greater than this threshold (Baskin and others, 2007b).

7

Species name according to the Integrated Taxonomic Information 

System (2015).

Several distinct community types that exist as zonal 

components of limestone cedar glades are characterized 

by seasonally saturated or inundated conditions. Such 

communities meet many of the hydrologic and vegetation 

criteria required for delineation as wetlands (U.S. Army 

Corps of Engineers, 2010; Taylor and Estes, 2012), although 

wetland delineation based on hydric soils is made difficult 

by shallowness and seasonal drying of soils (Taylor and 

Estes, 2012). These communities include the Limestone Seep 

Glade/Kentucky Glade Seep, the Limestone Glade Streamside 

Meadow community, and the Interior Low Plateau Limestone 

Glade Ephemeral Pool (Nordman, 2004; Norton, 2010). 

Characteristic species of the seasonally wet zones of limestone 

cedar glades include Isoetes butleri, Carex species (especially 

C. crawei), Eleocharis bifida, Nothoscordum bivalve, and 

Schoenolirion croceum as well as Nostoc cyanobacteria 

(Nordman, 2004; Baskin and others, 2007b).

Some debate and confusion has existed as to whether—

or to what extent—limestone cedar glades are subject to 

succession. For example, Baskin and others (2007b) suggested 

conceptual models for plant succession while simultaneously 

asserting that cedar glades are characterized by “natural, 

long-persisting (edaphic climax) plant communities.” 

Quarterman (1950b) delineated two primary types of succession in limestone cedar glades and later revised these through 

development of flow diagrams (Baskin and others, 2007b): 

succession beginning on bare rock (usually in very shallow 

depressions that retain some moisture after rains) and succession beginning in cracks and crevices in the bedrock. In the 

case of bare rock, early pioneers that establish the first thin soil 

layer are generally lichens, mosses, and cyanobacteria, and soil 

development is a process whereby isolated mats in depressions 

grow deeper and wider (usually showing concentric bands of 

pioneer species at the periphery and invader species in the 

center) until mats meet and merge (Baskin and others, 2007b). 

In the case of bedrock cracks, vegetation becomes established 

in fissures and deposits organic matter onto the surface of 

adjacent bedrock blocks, contributing to gradual accumulation 

of soil. Deep-rooted perennials established in fissures are able 

to tap into a source of soil moisture—in the form of decayed 

shale below the surface layers of limestone—that is unavailable 

to plants in shallow soil on intact bedrock (Quarterman, 1989). 

Quarterman (1950b) asserted that both succession types 

progress through seral stages (defined as gravel glade, grass 

glade, glade-shrub, shrub-redcedar, redcedar stand) toward 

redcedar-hardwood forest as the climax stage. Quarterman 

(1950b) acknowledged, however, that in both cases, succession 

is probably extremely slow, and that the “open glade stage” 

and “cedar stand in crevices with expanses of flat-topped 

blocks of limestone between trees” community types can be 

long-persisting subclimax stages for the two types, respectively. 

Thus, the competing hypotheses of how best to conceptualize 

the different observed communities within cedar glades (seral 

stages of succession versus climax communities along stress 

gradients) may be in part a question of time scales. At relatively 
28 Insular Ecosystems of the Southeastern United States

short time scales, these community types probably represent 

sequences along a gradient, whereas at longer timescales that 

allow for the weathering of bedrock and the very slow accumulation of soil, more xeric community types might follow a 

succession pathway toward more mesic conditions. 

Additionally, periodic disturbance—in the form of soil 

erosion from sheet flow during storm events, tree and shrub 

death from droughts, and fire—may result in changes in 

vegetation structure that are cyclical rather than directional 

(Quarterman, 1989; Quarterman and others, 1993). However, 

disturbance from fire is generally considered to be less 

important in maintaining southeastern limestone cedar glades 

than it is for “glades” (actually xeric limestone prairies, see 

chapter D) of the Ozark Plateaus and Midwestern States 

(Harper, 1926; Baskin and others, 2007a). Community 

composition in cedar glades may also be determined in part 

by geographical attributes. For a group of 40 cedar glades 

in Middle Tennessee, Cofer and others (2008) found that 

richness—of total species, native and exotic species, and 

endemic and nonendemic species—all increased significantly 

(p<0.05) with increasing glade area and perimeter and 

decreased significantly with increasing isolation (distance 

between glade edge and the edge of the nearest glade).

Differentiation from Related 

Ecosystems

Limestone cedar glades differ in several ways from a 

closely related ecosystem type, xeric limestone prairies (see 

chapter D). Although both ecosystems develop on shallow, 

rocky soils over calcareous substrates, they differ in terms of 

topography, community composition, and factors maintaining 

ecosystem integrity (table C–1). Limestone cedar glades 

generally are found on areas with level to moderate slopes, are 

dominated by summer annual grasses, and are considered to be 

an edaphic climax ecosystem, whereas xeric limestone prairies 

may be steeply sloped, are dominated by perennial grasses, 

and are maintained in part by disturbance (Baskin and Baskin, 

2000; Lawless and others, 2006a, b; Baskin and others, 

2007a). Additionally, mosses, lichens, cyanobacteria, and 

forbs are important in limestone cedar glades but much less so 

in xeric limestone prairies, and may be entirely absent in the 

latter (Baskin and others, 2007a). Despite these differences, 

even experts have had some difficulty differentiating between 

the two ecosystems (Baskin and Baskin, 2004). For example, 

Baskin and Baskin (1977) first described a community in 

Wilson County, Tennessee, as “an undescribed cedar glade” 

due to its dominance by Schizachyrium scoparium; however, 

they later stated that this site was actually a xeric limestone 

prairie (Baskin and Baskin, 2004).

Calcareous forest openings associated with stands of 

J. virginiana have been referred to as “cedar glades” in States 

including Oklahoma, Kansas, Missouri, Illinois, Indiana, 

Wisconsin, and Ohio (Baskin and Baskin, 1999; McClain and 

Ebinger, 2002; Mills, 2008). Although midwestern “cedar 

glades” are somewhat similar to southeastern cedar glades 

in terms of physiognomy and vegetation, Baskin and Baskin 

(2000, 2003) and Baskin and others (2007a) argued that they 

are fundamentally different ecosystems—more accurately 

categorized as barrens or as xeric limestone prairies—because 

they are maintained primarily by disturbance (fire and grazing) 

rather than by fundamental edaphic constraints and in the 

absence of disturbance will quickly succeed to redcedar and 

(or) hardwood forests. Forest openings referred to as “cedar 

glades” in the Ozark Plateaus are topographically different 

from true limestone cedar glades. Ozark “glades” commonly 

have slopes up to 50 percent, allowing them to be better 

drained than cedar glades of the Nashville Basin, which tend 

to be on level terrain (Baskin and others, 1968; Gates and 

others, 1982). “Cedar glades” in the Ozarks, like similar 

ecosystems in Midwest States, are primarily disturbance 

maintained and are not edaphic climax communities (Baskin 

and others, 1995). Similarly, “cedar glades” in West Virginia 

(Bartgis, 1993) have some physical and floristic similarity to 

limestone cedar glades of the Nashville Basin; however, the 

“glades” in West Virginia are more steeply sloping, tend to be 

well drained during the winter, support few winter annuals, 

and have a different suite of community dominants, leading 

Lawless and others (2006a) to categorize them as xeric 

limestone prairies. 

“Cedar glades” have also been described as outcrop 

communities on diabase bedrock in the Piedmont region. 

Although these “cedar glades” have some floristic similarity 

to Nashville Basin limestone cedar glades, they also share 

some characteristic species with Piedmont granite outcrops 

(LeGrand, 1988). The Ketona dolomite glades of Bibb County, 

Alabama, also share some floristic elements with limestone 

cedar glades of the Nashville Basin; however, they are more 

steeply sloping and sustain a suite of endemic taxa not found 

in Nashville Basin glades (Allison and Stevens, 2001; Duncan 

and others, 2008). The Ketona dolomite glades were classified 

as xeric limestone prairies by Baskin and Baskin (2003) and are 

discussed in chapter D. 

As a general rule, Quarterman (1989) and Baskin and 

Baskin (2004) reserve use of the term “glades” only for 

communities with less than 50 percent perennial grass cover, 

and they refer to communities with greater than 50 percent 

perennial grass cover as “barrens” or xeric limestone prairies. 

Quarterman (1989) similarly asserted that “barren” is a better 

term than “glade” for communities that are dominated by 

Schizachyrium scoparium, noting that this community type is 

generally found only on soil at least 8 inches (about 20 centimeters) deep. Early studies often used the term “cedar glades” 

to describe the outcrop areas and the accompanying redcedar 

forest; however, authors more recently have restricted use of 

the term to the outcrops only (Baskin and Baskin, 2004).
Chapter C 29

Table C–1. Factors differentiating limestone cedar glades from xeric limestone prairies, adapted from Baskin and others (2007a).

[<, less than; cm, centimeter; m, meter]

Factor Limestone cedar glades Xeric limestone prairies References

Ecosystem 

maintenance

Edaphic climax communities, 

disturbance is less important 

to ecosystem maintenance

Greater importance of disturbance (fire and grazing)Baskin and Baskin (2000), Lawless

and others (2006a), Baskin and

others (2007a)

Topography Level to moderate slopes Moderate to steep slopes, 

typically with southern

to western aspect

Erickson and others (1942), Bartgis 

(1993), Quarterman and others 

(1993), Baskin and Baskin (2000), 

Laughlin and Uhl (2003), Baskin and 

others (2007a)

Soil depth (depth

to bedrock)

< 30 cm <1 m Quarterman (1950a), Baskin and Baskin 

(1977), Anderson and others (1999), 

Baskin and Baskin (2003), Lawless

and others (2006a), Baskin and

others (2007a)

Geographic range 

(physiographic 

provinces)

Interior Low Plateaus, 

Appalachian Plateaus, 

and Valley and Ridge

Ozark Plateaus, Central Lowlands, Interior Low Plateaus, 

Appalachian Plateaus, Valley 

and Ridge, and Coastal Plain

Baskin and Baskin (1999), Baskin

and Baskin (2003), Lawless and

others (2006a)

Geographic range 

(States)

Tennessee, Kentucky, Virginia, 

Alabama, and Georgia

Arkansas, Missouri, Illinois, 

Indiana, Ohio, Pennsylvania, 

West Virginia, Virginia, 

Kentucky, Tennessee, 

Alabama, and Georgia

Baskin and Baskin (1999), Baskin

and Baskin (2003), Lawless and

others (2006a)

Center(s) of endemism Inner Nashville Basin of 

Middle Tennessee

Valley and Ridge of Alabama 

(Ketona dolomite glades) and 

Ozark Plateaus of Arkansas 

and Missouri

Baskin and Baskin (1999), Baskin

and Baskin (2003), Lawless and

others (2006a)

Dominant vegetation C4 summer annual grasses 

(for example, Sporobolus 

vaginiflorus) and forbs

C4 perennial grasses (for 

example, Schizachyrium 

scoparium)

Quarterman (1989), Baskin and others 

(1995), Baskin and Baskin (1999, 

2000), Lawless and others (2006a, b)

Winter annual taxa Relatively common Relatively few Lawless and others (2004)

Importance of 

bryophytes, 

cryptogams, and 

cyanobacteria

High Low Somers and others (1986), Lawless

and others (2006b), Baskin and

others (2007a)

Contributions to Regional Biodiversity

Limestone cedar glades are ecosystems of conservation 

concern primarily because of their contributions to plant 

biodiversity (Noss, 2013). Some regionally rare plants 

in limestone cedar glades are endemics, whereas others 

represent disjunct populations of geographically widespread 

taxa from outside the southeastern United States. Limestone 

cedar glades also provide habitat to a number of animals of 

conservation concern.

Endemic and Biogeographically Disjunct Taxa

Limestone cedar glades support concentrations of plant 

taxa that are of conservation concern. Of all the plant taxa 

listed as rare in Tennessee by the Tennessee Division of 

Natural Heritage, 10 percent are found in limestone cedar 

glades (Vandevender, 2006). In Kentucky, 51 taxa that are 

considered rare or threatened at the State or Federal level 

are found within “forest openings” that include limestone 

cedar glades and related ecosystems such as xeric limestone 
30 Insular Ecosystems of the Southeastern United States

prairies (Rhoades and others, 2004). The rare species associated with limestone cedar glade habitat include a number of 

endemic species, including so-called “strict endemics” that 

have only been found in limestone cedar glades as well as 

“near endemics” that are primarily restricted to cedar glades 

but have occasionally been found in related ecosystems 

such as xeric limestone prairies or outcrop communities on 

noncalcareous substrates. Some plant taxa are endemic only 

to cedar glades within certain geographic areas. Astragalus 

bibullatus, Echinacea tennesseensis, Leavenworthia stylosa, 

and Phacelia dubia var. interior are restricted to the Nashville 

Basin; Delphinium alabamicum, Leavenworthia alabamica, 

Leavenworthia crassa, and Lesquerella lyrata are restricted to 

northern Alabama; and Leavenworthia exigua var. laciniata is 

restricted to the Outer Bluegrass region of northern Kentucky 

(Baskin and others, 1995; Baskin and Baskin, 2003). Many of 

the cedar glade endemics and near-endemics are regionally or 

globally rare and are of conservation concern. For example, 

E. tennesseensis, A. bibullatus, and Dalea foliosa have been 

federally listed as endangered or threatened species (Norton, 

2010; Sutter and others, 2011), although E. tennesseensis was 

delisted in 2011 (Bowen, 2011). Several characteristic glade 

plants are shown in figure C–5. Of these, A. bibullatus is 

considered a strict endemic to limestone cedar glades, while 

Dalea gattingeri, Leavenworthia exigua, and Pediomelum 

subacaule are considered near-endemics because they also 

are present in xeric limestone prairies.

Limestone cedar glade endemics and near-endemics are 

primarily herbaceous dicots, therophytes, and hemicryptophytes, and all have a C3

 photosynthetic pathway (Baskin 

and Baskin, 1985, 1999). Generally, cedar glade interiors are 

areas of high abiotic stress and low competition relative to 

the surrounding forest, and cedar glade endemics tend to be 

specially adapted to stress conditions and poor competitors 

with mesic species (Quarterman, 1950b; Norton, 2010). Cedar 

glade endemics tend to be obligate heliophytes, unable to 

compete in even moderately shaded environments (Baskin and 

others, 1968; Baskin and Baskin, 1988; Quarterman, 1989). 

Most glade endemics have no special morphological adaptations for seed dispersal, and their mode of dispersal is gravity; 

however, some seeds of endemics are dispersed by wind and 

(or) water (Baskin and Baskin, 1989; Cofer and others, 2008).

Although considerable research has been done on the 

autecology and growth requirements of cedar glade endemics, 

the full suite of factors controlling their distribution and 

local abundance is still not fully understood. Some evidence 

suggests that certain glade species grow preferentially in 

specific microhabitats. Somers and others (1986) found that 

Talinum calcaricum, Leavenworthia stylosa, and L. exigua

grew in soils that were substantially shallower than average 

cedar glade soils and that Viola egglestonii, Pediomelum 

subacaule, and L. torulosa grew preferentially toward the 

edges of glades near the margin of the surrounding redcedar 

forest. These results are indicative of niche partitioning among 

cedar glade endemics, based in part on soil depth and shading 

gradients. The spatial distribution of some glade endemics is 

also apparently regulated in part by allelopathy, as several 

glade species contain chemicals that inhibit seed germination 

of other taxa (Quarterman, 1973; Quarterman, 1989). For 

example, germination of Minuartia patula, a nonendemic, 

is inhibited by extracts from Dalea gattingeri (Baskin and 

Baskin, 1989; Quarterman and others, 1993).

Attempts to identify the proximal cause(s) of endemism 

for cedar glade species have been only partially successful. 

Although research indicates that all of the endemics are 

shade-intolerant and compete poorly with mesic species, these 

factors alone do not appear sufficient to fully explain their 

endemism, and no other unique biological or physiological 

traits have been identified that fully explain why endemics 

are restricted to cedar glade habitat (Baskin and Baskin, 1988, 

1989). Edaphic factors (soil chemical and physical properties) 

and genetic factors were explored by Baskin and Baskin 

(1988), but neither adequately accounted for the observed 

endemism of limestone cedar glades. Baskin and others (1997) 

compared Echinacea tennesseensis (a cedar glade endemic) 

and Solidago shortii (which grows in glade-like communities) 

to their geographically widespread congeners—E. angustifolia

and S. altissima, respectively—and ruled out a number of 

physiological, ecological, cytological, genetic, and life-history 

factors in explaining the narrow geographic restriction of the 

endemics, leading them to conclude that historical factors, 

namely the presence or absence of past opportunities for range 

expansion, likely played some role in fostering endemism for 

these glade species; see also Walck and others (2001).

One set of traits shared by many cedar glade endemics 

involves special adaptations to the particular stress regime of 

their ecosystem (high levels of sunlight, thin soil, and widely 

fluctuating soil temperatures and hydrologic conditions). 

Types of adaptations common to cedar glade endemics 

and to other characteristic glade taxa include tolerance of 

high irradiance and drought (Baskin and Baskin, 1974; 

Quarterman and others, 1993) as well as tolerance of seasonal 

inundation and seasonally saturated soil conditions (Taylor 

and Estes, 2012). Quarterman (1950a) pointed to the timing 

of life cycles, particularly the timing of dormancy, as being 

a critically important adaptive strategy to survive seasonally 

extreme hydrologic conditions. A number of taxa complete 

their life cycles during the wet season and survive the dry 

season as seeds, corms, or bulbs. Adaptive advantages of 

the winter-annual life cycle include not only the ability to 

survive summer drought conditions through dormancy but 

also the ability to persist as seedlings during the winter when 

frost heaving of the thin soil is excessive (Quarterman and 

others, 1993). A special adaptation common to the winterannual glade endemics is self-compatibility (the ability to 

self-pollinate and set seed), which is necessary for life-cycle 

completion because few pollinators are active during the 

winter and early spring flowering period (Quarterman and 

others, 1993; Noss, 2013). Some other glade taxa are dormant 

during the period of saturation and complete their life cycle 

in the summer and fall. Some glade taxa possess specialized 

adaptations to cope with seasonal saturation and anaerobic 
Chapter C 31

Figure C–5. Characteristic plant taxa 

of limestone cedar glades, A, Astragalus 

bibullatus, Flat Rock Cedar Glades and 

Barrens State Natural Area, Rutherford 

County, Tennessee; B, Dalea gattingeri, 

Chickamauga and Chattanooga National 

Military Park, Catoosa County, Georgia; 

C, Leavenworthia exigua, Chickamauga 

and Chattanooga National Military Park; 

D, Pediomelum subacaule, Flat Rock Cedar 

Glades and Barrens State Natural Area. 

Photographs by Alan M. Cressler, 

U.S. Geological Survey.

A

C

D

B
32 Insular Ecosystems of the Southeastern United States

conditions. Leavenworthia uniflora, for example, does not 

accumulate ethanol in roots the way most plants do when 

in saturated soil conditions, suggesting that it has metabolic 

adaptations to protect it from cytotoxic effects of root zone 

hypoxia (Baskin and Baskin, 1989; Norton, 2010).

Another characteristic shared by several cedar glade 

taxa is biogeographic disjunction, meaning that occurrences 

within limestone cedar glades are geographically remote 

from primary ranges in other regions or represent substantial 

range expansions; see Braun (1955) and Brown and others 

(1996). Harper (1926) observed that a number of plant taxa 

in limestone cedar glades were disjunct to home ranges in the 

Great Plains. Since then, several generations of researchers 

have noted glade species disjunctions to the Ozark Plateaus 

and to the Great Plains (Bridges and Orzell, 1986; Quarterman 

and others, 1993; Norton, 2010). Glade-associated taxa with 

population centers in physiographic regions where limestone 

cedar glades are not present include Dalea foliosa, which is in 

glaciated northern regions of the Midwest, and Oxalis priceae

ssp. priceae, whose range includes the Coastal Plain of 

Alabama (Baskin and Baskin, 2003). Additionally, cedar 

glades support a number of eastern species of western genera, 

for example the narrow endemic Astragalus bibullatus, which 

belongs to a genus mostly found in arid western regions 

(Noss, 2013). Other primarily western genera with species 

found in cedar glades of the southeastern United States include 

Echinacea, Onosmodium, Pediomelum, Penstemon, and 

Phemeranthus (Noss, 2013).

Animal Taxa of Conservation Concern

In comparison to studies of vascular plants, relatively 

few investigations in limestone cedar glades have focused on 

animals or microbes. Martin and Sharp (1983) compiled species 

lists of soil protozoa and found that species richness was not 

well correlated with available soil moisture or pH. Martin and 

Sharp (1983) also noted a conspicuous absence of the ciliate 

genus Colpoda, which is usually abundant in soil. Meyer (1937) 

published detailed records of invertebrate species (annelids, 

mollusks, and arthropods) collected in a Middle Tennessee 

cedar glade, but found no substantial difference in species 

composition among different habitat types when open glades 

were compared to redcedar forest and to shrub-dominated 

areas. Hill (2010; 2012), however, focusing exclusively on 

grasshoppers, found discernible differences in species composition between open glade habitat, redcedar forest, and adjacent 

xeric limestone prairie. This survey of grasshoppers in cedar 

glades and associated ecosystems also resulted in documentation of four species for the first time in Tennessee, as well as 

the description of a new species, Melanoplus ingrami, which 

is suspected to be a cedar glade endemic. Although aquatic 

invertebrate fauna were observed by Taylor and Estes (2012) in 

seasonally inundated zones of cedar glades, no formal documentation or study of aquatic fauna has yet been published.

Working in another Middle Tennessee cedar glade, Jordan 

and others (1968) reported a community composition of 

amphibians and reptiles different from that normally found in 

the surrounding region—most notably the absence of species 

requiring moist or aquatic conditions—and suggested that the 

dry conditions of cedar glades provided a favorable habitat for 

lizards and lizard-consuming snakes. Some amphibian taxa—

such as Anaxyrus fowleri (Fowler’s toad)—were more abundant in the open, rocky glade areas relative to the surrounding 

forest, whereas other species—such as Anaxyrus americanus

(the American toad)—showed the opposite trend. Based on 

20 years of herpetofaunal survey data from seven sites in 

Middle Tennessee containing cedar glade habitat, Niemiller 

and others (2011) concluded that although no amphibian or 

reptile species are endemic to cedar glades, several species 

are more frequently encountered in cedar glade habitat than in 

surrounding mesic forest and that cedar glades and associated 

woodlands provide habitat to several species of conservation 

concern. Schultz (1930) found that many of the bird species 

present in a Middle Tennessee cedar glade were characteristic 

of early successional fields and other open or edge habitat, with 

a corresponding paucity of bird species requiring deciduous 

forest conditions. Although no bird species have been identified 

as relying solely on limestone cedar glades for habitat, cedar 

glades do support a number of avian species of conservation 

concern, including Thryomanes bewickii (Bewick’s wren), 

Setophaga discolor (prairie warbler), and Vermivora cyanoptera (blue-winged warbler) (Ford and others, 2000).

Conservation Considerations

A number of researchers familiar with the biodiversity 

contributions of limestone cedar glades have stressed the 

importance of their conservation (Quarterman and others, 

1993; Noss and others, 1995; Baskin and others, 2007b). These 

arguments commonly have involved descriptions of current 

and historical threats to limestone cedar glades, assessment 

of existing conservation approaches, and identification of 

scientific knowledge gaps.

Threats to Ecosystem Integrity

According to estimates by the Tennessee Department 

of Environment and Conservation, about 50 percent of total 

limestone cedar glade area has been lost and 90 percent of 

ecologically intact limestone cedar glades have been lost from 

the Nashville Basin of Middle Tennessee, rendering limestone 

cedar glades an “endangered ecosystem” (Noss and others, 

1995). Population growth and suburban sprawl are commonly 

cited as major threats to limestone cedar glades forming on 

land that is not already publicly owned or otherwise protected 

from development (Nordman, 2004; Baskin and others, 
Chapter C 33

2007b). Comparison of Freeman’s (1933) hand-drawn map of 

his study site to current aerial photography indicates that the 

cedar glades upon which he based his well-cited publication 

have likely been destroyed by human development. In Middle 

Tennessee, the construction of the Percy Priest Reservoir 

flooded more than 202,000 acres (about 82,000 hectares) in 

the “heart of glade country” and almost certainly resulted in 

the destruction of many cedar glades and associated redcedar 

forests, as well as encouraged private development along the 

lakeshore (Quarterman and others, 1993).

Harper (1926) observed limestone cedar glades in Middle 

Tennessee that had been severely damaged by lumbering and 

grazing. Quarterman (1950b) mentioned grazing, the cutting 

of Juniperus virginiana trees for fenceposts, and the scraping 

of glades to harvest gravel as primary threats. Niemiller 

and others (2011) described degradation of cedar glades 

and related ecosystems resulting from off-road recreational 

vehicles, rock gathering, and suburban sprawl. Dumping of 

garbage in cedar glades has also been a problem, arising from 

a popular misperception of cedar glades as “wastelands” 

(Quarterman and others, 1993). Field reconnaissance in 

Tennessee and Kentucky by Norton (2010) indicated that 

most cedar glades on private land showed evidence of severe 

anthropogenic destruction—grazing, recreational vehicle use, 

garbage dumping, and scraping for gravel—and these glades 

on private land supported relatively few rare taxa. 

One of the indirect effects of anthropogenic disturbance 

is often the introduction of exotic species. Quarterman and 

others (1993) suggested that hikers and off-road vehicles can 

introduce weeds such as Ambrosia artemisiifolia and Cirsium 

discolor, which can out-compete glade flora and degrade 

the structure of glade communities. Sutter and others (2011) 

reported a marked increase of woody species, particularly 

the exotic Ligustrum sinense, in calcareous glades between 

1993 and 2008 and attributed these changes primarily to fire 

suppression and the introduction of invasive species; however, 

they did not differentiate between limestone cedar glades 

(edaphic climax communities) and xeric limestone prairies 

(already known to be maintained by disturbance), and thus 

the relevance of their findings to only cedar glades is unclear. 

Comparing L. sinense, which is listed as a “severe threat” to 

native plant communities by the Tennessee Exotic Pest Plant 

Council (2009), to Forestiera ligustrina, which is a native 

shrub with a similar growth habit, Morris and others (2002) 

found that L. sinense had spatial and temporal advantages in 

capturing light as well as the production of a large quantity 

of fruits, both attributes that might account for its ability to 

rapidly colonize and displace native species.

In recent studies, the percentage of total glade flora 

found to be exotic species has generally been between 

10 and 20 percent: 18.8 percent in a cluster of cedar glades in 

Middle Tennessee (Cofer and others, 2008); 13.4 percent for 

a group of glades in northern Alabama (Baskin and others, 

1995); and 17.6 percent for glade flora across the southeastern 

United States as a whole (Baskin and Baskin, 2003). By 

contrast, forest canopies growing on glade-forming soils in 

the Inner Nashville Basin contain few if any exotic species 

(Adams and others, 2012). 

Because most glade endemics are poor competitors 

and fill a niche based on high abiotic stress, the greatest 

threat from exotic invasive species would presumably be the 

arrival of exotics that are competitive and have adaptations 

for drought or anaerobic tolerance (Norton, 2010). In the 

southeastern United States, the exotic species considered 

to be of greatest potential threat to limestone cedar glades 

are Lonicera japonica, L. maackii, and Ligustrum sinense

(Nordman, 2004; Cofer and others, 2008; Norton, 2010). 

Herbaceous exotic species may present a greater threat than 

woody exotics because of their ability to invade areas beyond 

crevices and glade perimeters (Cofer and others, 2008; Norton, 

2010). Examples of herbaceous exotics with potential to invade 

limestone cedar glades include Carduus nutans, Daucus 

carota, Lespedeza cuneata, Leucanthemum vulgare, and 

Melilotus officinalis (Cofer and others, 2008; Norton, 2010), 

all of which are listed as current or potential threats to native 

plant communities in Tennessee by the Tennessee Exotic Pest 

Plant Council (2009). 

Conservation Strategies

A number of State and Federal conservation programs are 

underway in an attempt to preserve limestone cedar glades. 

A few examples of State, Federal, and private preserves 

containing protected limestone cedar glades are Chickamauga 

and Chattahoochee National Military Park, Cedar Glades 

State Forest and Natural Area, Cedars of Lebanon State 

Park, Long Hunter State Park, Mount View Cedar Glade, 

the Percy Priest Lake area, Flat Rock Cedar Glade Preserve, 

and Stones River National Battlefield (Quarterman and 

others, 1993). In some cedar glades on protected and publicly 

owned land, native plant restoration programs are underway 

(Vandevender, 2006), including attempts to establish new 

and self-sustaining populations of rare cedar glade endemics 

in suitable glade habitat where those endemics do not 

currently grow (Nordman, 2004; Albrecht and McCue, 2010). 

Quarterman and others (1993) pointed out that the sparse or 

patchy distribution of some glade endemic species complicates 

the process of selecting individual glades or glade clusters for 

conservation. Species inventories at the level of individual 

glades may be necessary to ensure that conservation strategies 

encompass the full range of gene pools for endemic species. 

Also, conservation of small glades may be equally important 

to that of larger glades because small glades may serve 

as corridors between larger glades and thus help facilitate 

dispersal and establishment of glade endemics, especially 

those whose seeds are dispersed by gravity (Mann and others, 

1999; Cofer and others, 2008).
34 Insular Ecosystems of the Southeastern United States

Knowledge Gaps

Although several generations of researchers have 

intensively studied limestone cedar glades and their endemic 

species, many important questions remain unanswered. 

For example, few researchers have conducted systematic 

hydrologic investigations of limestone cedar glades, and 

currently the nonprecipitation sources of water inputs to 

these ecosystems as well as their connectivity to subterranean 

karst features is largely unknown (Norton, 2010). Although 

the presence of aquatic fauna has been noted (Taylor and 

Estes, 2012), systematic documentation of aquatic fauna or 

their trophic relationships has yet to be conducted. Some 

possible anthropogenic effects on cedar glades have not been 

scientifically examined, including nitrogen deposition due to 

air pollution (Jeffrey Walck, Middle Tennessee State University, 

written commun., 2014) and localized changes to precipitation 

regimes due to heat-island effects from nearby urban areas 

(Shepherd and others, 2002). The degree to which limestone 

cedar glades are vulnerable to woody encroachment has not 

been conclusively established. Also, compared to studies of 

the vegetation of cedar glade interiors and associated redcedar 

stands, less attention has been paid to the surrounding mixed 

hardwood forests that occupy the same glade-forming soils 

(Adams and others, 2012).

Possible Ecological Effects from 

Climate Change

Although cases of relatively obvious anthropogenic 

destruction have been noted repeatedly by cedar glade 

researchers, very little attention has been paid to the potential 

effects of climate change. As discussed in chapter D, MolanoFlores and Bell (2012) developed a population variability 

analysis for Dalea foliosa based on climate-change scenarios. 

This analysis, however, was for a single plant species (present 

in limestone cedar glades but also in related prairie contexts 

in the northern Midwest) and was based on populations in 

noncedar-glade habitat in Illinois, rendering its applicability 

to limestone cedar glade ecosystems unclear. For example, 

Molano-Flores and Bell (2012) identified changes in seasonal 

timing of snowmelt as important to Illinois populations of 

D. foliosa. Such effects on populations in limestone cedar 

glades farther south would likely be different because of 

regional climatic differences.

Regional changes in the timing, amount, and (or) 

variability of precipitation have not yet been systematically 

investigated as a possible threat to limestone cedar glades. 

Such changes might be expected to affect glade-associated taxa 

with life histories that are closely tied to historical patterns of 

seasonal soil desiccation and saturation. Furthermore, regional 

temperature changes might affect such taxa by altering evapotranspiration dynamics. For example, Norton (2010) suggested 

that increased temperatures during the spring might lead to 

seasonally saturated zones within cedar glades drying out 

earlier in the year, making it more difficult for hydrophytic taxa 

within these zones to complete their life cycles. This hypothesis 

was not empirically tested and does not account for interactive 

effects of temperature changes with possible seasonal increases 

in regional precipitation. Increased summer temperatures or 

changes in precipitation timing and magnitude might also alter 

seasonal patterns of xeric soil conditions in cedar glades, with 

ecological effects that could be difficult to predict. For example, 

increased duration or intensity of summer drought conditions might be expected to help exclude mesic competitors 

from cedar glade habitat; however, such changes might also 

overwhelm the adaptive capacities of glade endemics to survive 

xeric conditions. 

Predicting ecological effects of climate change in limestone cedar glades is also complicated by spatial variability 

in soil moisture dynamics and by the coexistence of plant 

taxa with differing adaptive strategies for surviving extreme 

hydrologic conditions. In addition to—or interacting with—

climate-change-induced alteration of abiotic stress regimes, 

alterations to biological interactions could represent a possible 

threat to cedar glade ecological integrity. For example, in other 

ecosystems, changes to the seasonal timing of the availability 

of pollinators has resulted in phenological asynchrony, a 

phenomenon whereby differential shifts in seasonal timing of 

life cycles can disrupt interspecific interactions (Walther and 

others, 2002; Walther, 2010). An assessment of the potential for 

such disruptions to occur in limestone cedar glade communities would require more detailed understanding of the interspecific interactions governing key constituents of cedar glade 

ecosystems and of their tolerance to seasonal shifts.
Chapter C 35

References

Adams, D., Walck, J., Howard, R., and Milberg, P., 2012, 

Forest composition and structure on glade-forming 

limestones in Middle Tennessee: Castanea, v. 77, no. 4, 

p. 335–347.

Albrecht, M., and McCue, K., 2010, Changes in demographic 

processes over long time scales reveal the challenge of 

restoring an endangered plant: Restoration Ecology, v. 18, 

no. S2, p. 235–243.

Allison, J., and Stevens, T., 2001, Vascular flora of Ketona 

dolomite outcrops in Bibb County, Alabama: Castanea, 

v. 66, no. 1, p. 154–205.

Anderson, R., Fralish, J., and Baskin, J., 1999, Savannas, 

barrens, and rock outcrop plant communities of North 

America: Cambridge University Press, New York.

Bartgis, R., 1993, The limestone glades and barrens of 

West Virginia: Castanea, v. 58, no. 2, p. 69–89.

Baskin, C., and Baskin, J., 1974, Responses of Astragalus 

tennesseensis to drought: Oecologia, v. 17, no. 1, p. 11–16.

Baskin, J., and Baskin, C., 1977, An undescribed cedar glade 

community in Middle Tennessee: Castanea, v. 42, no. 2, 

p. 140–145.

Baskin, J., and Baskin, C., 1981, Photosynthetic pathways 

indicated by leaf anatomy in 14 summer annuals of cedar 

glades: Photosynthetica, v. 15, p. 205–209.

Baskin, J., and Baskin, C., 1985, Photosynthetic pathway 

in 14 southeastern cedar glade endemics, as revealed by 

leaf anatomy: American Midland Naturalist, v. 114, no. 1, 

p. 205–208.

Baskin, J., and Baskin, C., 1988, Endemism in rock outcrop 

plant communities of unglaciated eastern United States—

An evaluation of the roles of the edaphic, genetic and light 

factors: Journal of Biogeography, v. 15, no. 5, p. 829–840.

Baskin, J., and Baskin, C., 1989, Cedar glade endemics in 

Tennessee, and a review of their autecology: Journal of the 

Tennessee Academy of Science, v. 64, no. 3, p. 63–74.

Baskin, J., and Baskin, C., 1999, Cedar glades of the southeastern United States, in Anderson, R., Fralish, J., and 

Baskin, J., eds., Savannas, barrens, and rock outcrop plant 

communities of North America: New York, Cambridge 

University Press, p. 206–219.

Baskin, J., and Baskin, C., 2000, Vegetation of limestone and 

dolomite glades in the Ozarks and Midwest regions of the 

United States: Annals of the Missouri Botanical Garden, 

v. 87, no. 2, p. 286–294.

Baskin, J., and Baskin, C., 2003, The vascular flora of cedar 

glades of the southeastern United States and its phytogeographical relationships: Journal of the Torrey Botanical 

Society, v. 130, no. 2, p. 101–118.

Baskin, J., and Baskin, C., 2004, History of the use of 

“cedar glades” and other descriptive terms for vegetation

on rocky limestone soils in the Central Basin of Tennessee: 

The Botanical Review, v. 70, no. 4, p. 403–424.

Baskin, J., Baskin, C., and Lawless, P., 2007a, Calcareous rock 

outcrop vegetation of eastern North America (exclusive of 

the Nashville Basin), with particular reference to use of the 

term “cedar glades”: The Botanical Review, v. 73, no. 4, 

p. 303–325.

Baskin, J., Baskin, C., and Quarterman, E., 2007b, Flow 

diagrams for plant succession in the Middle Tennessee 

cedar glades: Journal of the Botanical Research Institute 

of Texas, v. 1, no. 2, p. 1131–1140.

Baskin, J., Quarterman, E., and Caudle, C., 1968, Preliminary 

check-list of the herbaceous vascular plants of cedar glades: 

Journal of the Tennessee Academy of Science, v. 43, no. 3, 

p. 65–71.

Baskin, J., Snyder, K., Walck, J., and Baskin, C., 1997, The 

comparative autecology of endemic, globally-rare, and 

geographically-widespread, common plant species—Three 

case studies: The Southwestern Naturalist, v. 42, no. 4, 

p. 384–399.

Baskin, J., Webb, D., and Baskin, C., 1995, A floristic plant 

ecology study of the limestone glades of northern Alabama: 

Bulletin of the Torrey Botanical Club, v. 122, no. 3, 

p. 226–242.

Bowen, B., 2011, Natural areas protection at its best; 

protecting the Tennessee purple coneflower (Echinacea 

tennesseensis): Natural Areas Journal, v. 31, no. 4, 

p. 326–330.

Braun, E., 1955, The phytogeography of unglaciated eastern 

United States and its interpretation: The Botanical Review, 

v. 21, no. 6, p. 297–375.

Bridges, E., and Orzell, S., 1986, Distribution patterns of the 

non-endemic flora of Middle Tennessee limestone glades: 

ASB Bulletin, v. 33, p. 155–166.
36 Insular Ecosystems of the Southeastern United States

Brown, J., Stevens, G., and Kaufman, D., 1996, The 

geographic range—Size, shape, boundaries, and internal 

structure: Annual Review of Ecology and Systematics, 

v. 27, no. 1996, p. 597–623.

Cartwright, J., 2014, Soil ecology of a rock outcrop 

ecosystem—Abiotic stresses, soil respiration, and microbial 

community profiles in limestone cedar glades: Nashville, 

Tenn., Tennessee State University, Ph.D. Dissertation.

Cartwright, J., and Hui, D., 2014, Soil respiration patterns and 

controls in limestone cedar glades: Plant and Soil, v. 389, 

no. 1–2, p. 157–169.

Cofer, M., Walck, J., and Hidayati, S., 2008, Species richness and exotic species invasion in Middle Tennessee cedar 

glades in relation to abiotic and biotic factors: The Journal 

of the Torrey Botanical Society, v. 135, no. 4, p. 540–553.

Dubois, J., 1993, Biological dinitrogen fixation in two cedar 

glade communities of Middle Tennessee: Journal of the 

Tennessee Academy of Science, v. 68, no. 4, p. 101–105.

Duncan, R., Anderson, C., Sellers, H., and Robbins, E., 2008, 

The effect of fire reintroduction on endemic and rare plants 

of a southeastern glade ecosystem: Restoration Ecology, 

v. 16, no. 1, p. 39–49.

Eickmeier, W., 1986, The distribution of photosynthetic 

pathways among cedar glade plants: ASB Bulletin, v. 33, 

p. 200–205.

Erickson, R., Brenner, L., and Wraight, J., 1942, Dolomitic 

glades of east-central Missouri: Annals of the Missouri 

Botanical Garden, v. 29, no. 2, p. 89–101.

Estill, J., and Cruzan, M., 2001, Phytogeography of rare 

plant species endemic to the southeastern United States: 

Castanea, v. 66, no. 1, p. 3–23.

Fenneman, N., and Johnson, D., 1946, Physical divisions of 

the United States: U.S. Geological Survey map prepared 

in cooperation with the Physiographic Commission, scale 

1:7,000,000.

Ford, R., Carr, S., Hunter, C., York, J., and Reodel, M., 2000, 

Partners in flight bird conservation plan for the Interior 

Low Plateaus (Physiographic Area 14): Washington, D.C., 

American Bird Conservancy.

Freeman, C., 1933, Ecology of the cedar glade vegetation near 

Nashville Tennessee: Journal of the Tennessee Academy of 

Science, v. 8, p. 143–228.

Gates, J., Settergren, C., Henderson, G., and Krstansky, J., 

1982, Hydrology of a south-central Missouri cedar glade: 

Water and Soil Management, v. 1, p. 36–41.

Harper, R., 1926, The cedar glades of Middle Tennessee: 

Ecology, v. 7, no. 1, p. 48–54.

Hill, J., 2010, A new species of Melanoplus (Orthoptera: 

Acrididae: Melanoplinae) from the cedar glades of 

Tennessee, USA: Journal of Orthoptera Research, v. 19, 

no. 2, p. 341–345.

Hill, J., 2012, Habitat associations of grasshoppers (Orthoptera: Acrididae) in the heterogeneous cedar glade landscape 

of the Central Basin of Tennessee: Journal of Orthoptera 

Research, v. 21, no. 2, p. 227–233.

Integrated Taxonomic Information System, 2015, The 

Integrated Taxonomic Information System, accessed 

October 5, 2015, at http://www.itis.gov.

Jordan, O., Garton, J., and Ellis, R., 1968, The amphibians and 

reptiles of a Middle Tennessee cedar glade: Journal of the 

Tennessee Academy of Science, v. 43, no. 3, p. 72–78.

Laughlin, D., and Uhl, C., 2003, The xeric limestone prairies 

of Pennsylvania: Castanea, v. 68, no. 4, p. 300–316.

Lawless, P., Baskin, J., and Baskin, C., 2004, The floristic 

ecology of xeric limestone prairies in Kentucky, and a 

comparison to limestone cedar glades and deep-soil barrens: 

Sida, v. 21, no. 70, p. 1055–1079.

Lawless, P., Baskin, J., and Baskin, C., 2006a, Xeric limestone 

prairies of eastern United States—Review and synthesis: 

The Botanical Review, v. 72, no. 3, p. 235–272.

Lawless, P., Baskin, J., and Baskin, C., 2006b, Scaledependent classification of xeric limestone prairies—

Annual or perennial grasslands?: Annals of the 

Missouri Botanical Garden, v. 93, no. 3, p. 455–464.

LeGrand, H., 1988, Cedar glades on diabase outcrops—

A newly described community type: Castanea, v. 53,

no. 2, p. 168–172.

Mahr, W., and Mathis, P., 1981, Foliose and fruticose lichens 

of the cedar glades in Stones River National Battlefield Park 

(Rutherford County, Tennessee): Journal of the Tennessee 

Academy of Science, v. 56, no. 2, p. 66–67.

Mann, L., King, A., Dale, V., Hargrove, W., WashingtonAllen, R., Pounds, L., and Ashwood, T., 1999, The role 

of soil classification in geographic information system 

modeling of habitat pattern—Threatened calcareous 

ecosystems: Ecosystems, v. 2, p. 524–538.

Martin, E., and Sharp, R., 1983, Soil protozoa from a cedar 

glade in Rutherford County, Tennessee: Journal of the 

Tennessee Academy of Science, v. 58, no. 3 and 4, p. 31–36.
Chapter C 37

McClain, W., and Ebinger, J., 2002, A comparison of the 

vegetation of three limestone glades in Calhoun County, 

Illinois: Southeastern Naturalist, v. 1, no. 2, p. 179–188.

Meyer, A., 1937, An ecological study of cedar glade invertebrates near Nashville, Tennessee: Ecological Monographs, 

v. 7, no. 3, p. 403–443.

Mills, J., 2008, Fifty years of change in Wisconsin cedar 

glades: The American Midland Naturalist, v. 159, no. 1, 

p. 214.

Molano-Flores, B., and Bell, T., 2012, Projected population 

dynamics for a federally endangered plant under different 

climate change emission scenarios: Biological Conservation, v. 145, no. 1, p. 130–138.

Morris, L., Walck, J., and Hidayati, S., 2002, Reproduction 

of the invasive Ligustrum sinense and native Forestiera 

ligustrina (Oleaceae)—Implications for the invasion and 

persistence of a nonnative shrub: International Journal of 

Plant Sciences, v. 163, no. 6, p. 1001–1010.

Nelson, P., Fitzgerald, J., Larson, K., McCoy, R., Schotz, A., 

Taft, J., Witsell, T., and Yahn, B., 2013, Central hardwoods 

joint venture glade conservation assessment for the Interior 

Highlands and Interior Low Plateaus of the Central Hardwoods region: Central Hardwoods Joint Venture.

Niemiller, M., Reynolds, R., Glorioso, B., Spiess, J., and 

Miller, B., 2011, Herpetofauna of the cedar glades and 

associated habitats of the Inner Central Basin of Middle 

Tennessee: Herpetological Conservation and Biology, 

v. 6, no. 1, p. 135–149.

Nordman, C., 2004, Vascular plant community classification 

for Stones River National Battlefield: Durham, N.C., 

NatureServe. 

Norton, K., 2010, A floristic ecology study of seasonally wet 

limestone cedar glades of Tennessee and Kentucky: Clarksville, Tenn., Austin Peay State University, Masters Thesis.

Noss, R., 2013, Forgotten grasslands of the south: Washington, 

D.C., Island Press.

Noss, R., LaRoe, E., and Scott, J., 1995, Endangered ecosystems of the United States—A preliminary assessment of 

loss and degradation: Washington, D.C., U.S. Department 

of the Interior.

Quarterman, E., 1950a, Ecology of cedar glades. I. Distribution of glade flora in Tennessee: Bulletin of the Torrey 

Botanical Club, v. 77, no. 1, p. 1–9.

Quarterman, E., 1950b, Major plant communities of Tennessee 

cedar glades: Ecology, v. 31, no. 2, p. 234–254.

Quarterman, E., 1973, Allelopathy in cedar glade plant 

communities: Journal of the Tennessee Academy of 

Science, v. 48, no. 4, p. 147–150.

Quarterman, E., 1989, Structure and dynamics of the limestone cedar glade communities in Tennessee: Journal of the 

Tennessee Academy of Science, v. 64, no. 3, p. 155–158.

Quarterman, E., Burbanck, M., and Shure, D., 1993, Rock 

outcrop communities—Limestone, sandstone, and granite, 

in Martin, W., Boyce, S., and Echternacht, A., eds., 

Biodiversity of the southeastern United States—Upland 

terrestrial communities: New York, John Wiley and Sons, 

Inc., p. 35–86.

Rhoades, C., Miller, S., and Shea, M., 2004, Soil properties 

and soil nitrogen dynamics of prairie-like forest openings 

and surrounding forests in Kentucky’s Knobs Region: The 

American Midland Naturalist, v. 152, no. 1, p. 1–11.

Rollins, S., 1997, Calcareous glade communities in the 

Central Basin of Tennessee—The effects of scale on 

community classification: Chapel Hill, N.C., University 

of North Carolina, Masters Thesis.

Schultz, H., 1930, Birds of the cedar glades of Middle 

Tennessee: Nashville, Tenn., George Peabody College 

for Teachers, Thesis.

Shepherd, J., Pierce, H., and Negri, A., 2002, Rainfall modification by major urban areas—Observations from spaceborne rain radar on the TRMM satellite: Journal of Applied 

Meteorology, v. 41, p. 689–702.

Somers, P., Smith, L., Hamel, P., and Bridges, E., 1986, 

Preliminary analyses of plant communities and seasonal 

changes in cedar glades of Middle Tennessee: ASB Bulletin, 

v. 33, p. 178–192.

Sutter, R., Govus, T., Smyth, R., Nordman, C., Pyne, M., 

and Hogan, T., 2011, Monitoring change in a central 

U.S. calcareous glade—Resampling transects established

in 1993: Natural Areas Journal, v. 31, no. 2, p. 163–172.

Taylor, K., and Estes, D., 2012, The floristic and community 

ecology of seasonally wet limestone glade seeps of 

Tennessee and Kentucky: Journal of the Botanical 

Research Institute of Texas, v. 6, no. 2, p. 711–724.

Tennessee Exotic Pest Plant Council, 2009, TN-EPPC 

invasive exotic pest plants in Tennessee, accessed on 

November 18, 2015, at http://www.tneppc.org/invasive_

plants?sort_by=latin_name#invasive_plants_table.
38 Insular Ecosystems of the Southeastern United States

U.S. Army Corps of Engineers, 2010, Interim regional 

supplement to the Corps of Engineers wetland delineation manual—Eastern Mountains and Piedmont Region, 

in Wakeley, J., Lichvar, R., Noble, C., and Berkowitz, J., 

eds., ERDC/EL TR10-9: Vicksburg, Miss., Army Engineer 

Research and Development Center.

U.S. Department of Agriculture, 2015, Ecological subregions—Sections and subsections for the conterminous 

United States: Vector digital data, adapted from Fenneman 

and Johnson (1946) and ECOMAP (2007). Washington, 

D.C., U.S. Department of Agriculture, U.S. Forest Service. 

U.S. Geological Survey, National Gap Analysis Program, 

2011, Land cover data portal, accessed December 2015 at 

http://gapanalysis.usgs.gov/gaplandcover/data/download/.

Vandevender, J., 2006, Native plant restoration at Stones 

River National Battlefield, in People, Places and Parks: 

Proceedings of the 2005 George Wright Society Conference 

on Parks, Protected Areas and Cultural Sites: The George 

Wright Society, Inc., p. 146–148.

Walck, J., Baskin, J., and Baskin, C., 2001, Why is Solidago 

shortii narrowly endemic and S. altissima geographically 

widespread? A comprehensive comparative study of 

biological traits: Journal of Biogeography, v. 28, no. 10, 

p. 1221–1237.

Walther, G., 2010, Community and ecosystem responses to 

recent climate change: Philosophical Transactions of the 

Royal Society, B, v. 365, p. 2019–2024.

Walther, G., Post, E., Convey, P., Menzel, A., Paremesan, 

C., Beebee, T., Fromentin, J., Hoegh-Guldberg, O., and 

Bairlein, F., 2002, Ecological responses to recent climate 

change: Nature, v. 416, p. 389–395.
Chapter D. Xeric Limestone Prairies

Introduction

Xeric limestone prairies (fig. D –1) are open, moderately 

to steeply sloped nonforested areas, are dominated by C4

perennial grasses, and are present on relatively shallow, rocky 

soils derived from calcareous substrates (limestone and dolomite) that may include limited outcrops of bedrock (Laughlin 

and Uhl, 2003; Lawless and others, 2006a). Xeric limestone 

prairies commonly exist within a larger mosaic including 

woodland and shrub-dominated ecosystems—referred to as a 

“glade-woodland complex” or a “savanna-like ecosystem”—

and can grade into limestone cedar glades (Cipollini and 

others, 2005; Comer and others, 2011). 

Geographic Range and Differentiation 

from Related Ecosystems

Xeric limestone prairies are present in 6 physiographic 

provinces (the Ozark Plateaus, Central Lowland, Interior Low 

Plateaus, Appalachian Plateaus, Valley and Ridge, and Coastal 

Plain) in 12 States: Arkansas, Missouri, Illinois, Indiana, Ohio, 

Pennsylvania, West Virginia, Virginia, Kentucky, Tennessee, 

Alabama, and Georgia (fig. D–2; Lawless and others, 2006a). 

Within these States, xeric limestone prairies generally are 

present in unglaciated regions (Baskin and Baskin, 1999), such 

as in the extreme southern regions of Ohio and Indiana and 

in southern and western Illinois (Lawless and others, 2006a). 

Figure D–1. Grand Prairie, Coosa Valley Prairies, Floyd County, 

Georgia. Photograph by Alan M. Cressler, U.S. Geological Survey.
40 Insular Ecosystems of the Southeastern United States

Because of their relatively widespread geographic distribution, 

xeric limestone prairies develop in subtropical and temperate 

climatic regions and under a variety of precipitation regimes, 

with yearly precipitation amounts generally increasing from 

north to south and from west to east across their geographic 

range (Lawless and others, 2006a). 

Xeric limestone prairies have been referred to by a 

number of regional and colloquial names including barrens, 

alkaline barrens, prairies, dolomite prairies, prairie barrens, 

savannas, forest openings, grasslands, limestone glades, 

and cedar glades. This diversity of terminology has created 

considerable confusion, has slowed the recognition of 

fundamental similarities between related sites that are widely 

distributed geographically, and conversely has led to the conflation of ecosystems that are fundamentally distinct (Heikens 

and Robertson, 1994; Ludwig, 1999; Guyette and others, 2003; 

Cipollini and others, 2005; Lawless and others, 2006a). A 

considerable number of xeric limestone prairies have probably 

been mislabeled as other ecosystem types. For example, 

Copenheaver and others (2005) referred to a study site as “a 

typical southern Appalachian bald” despite the site having an 

elevation of less than 1,600 feet (about 500 meters) and having 

bedrock geology and vegetation composition that suggest it 

was actually a xeric limestone prairie. DeSelm and Murdock 

Figure D–2. General geographic distribution of xeric limestone prairies. 

95° 90° 85° 80° 75° 70°

45°

40°

35°

30°

25°

EXPLANATION

Physiographic province—From 

 Fenneman and Johnson (1946)

Appalachian Plateaus

Central Lowland

Coastal Plain

Interior Low Plateaus

Ozark Plateaus

Valley and Ridge

Selected occurrence of xeric 

 limestone prairies—Modified

 from Lawless and others (2006a)

0 100 200 300 400 KILOMETERS

0 100 200 300 400 MILES

Figure D-2

Base modifed from U.S. Department of Commerce, Census Bureau,

U.S. States, 2015, Mercator Auxiliary Sphere Projection

Horizontal datum is World Geodedic System of 1984

TENNESSEE

ALABAMA

GEORGIA

SOUTH

CAROLINA

NORTH CAROLINA

WEST

VIRGINIA

OHIO

PENNSYLVANIA

NEW YORK

MAINE

VERMONT

INDIANA

VIRGINIA

FLORIDA

KENTUCKY

DELAWARE

MARYLAND

MISSOURI

ARKANSAS

LOUISIANA

MISSISSIPPI

ILLINOIS

WISCONSIN

IOWA

MINNESOTA

MICHIGAN

NEW JERSEY

NEW HAMPSHIRE

CONNECTICUT

RHODE ISLAND

MASSACHUSETTS
Chapter D 41

(1993) noted that in Kentucky and Tennessee, grass-dominated 

communities of all types are generally referred to as “barrens.” 

Xeric limestone prairies have been referred to as “barrens” 

in Virginia (Ludwig, 1999), West Virginia (Bartgis, 1993), 

Illinois (Heikens and Robertson, 1995), and Indiana (Homoya, 

1994). However, xeric limestone prairies are easily differentiated from arable, deep-soil barrens, such as those in the 

“Big Barrens” region of southern and central Kentucky and 

northern Tennessee, by their endemism and thin soils (Baskin 

and others, 1999; Laughlin and Uhl, 2003).

Use of the terms “glades” and “cedar glades” in reference 

to xeric limestone prairies—for example Bartgis (1993), 

Kimmel and Probasco (1980), and Mills (2008)—has resulted 

in confusion in the literature related to the identification of 

both ecosystems. Recently, explicit attention to the problem of 

terminology has helped clarify which systems are true glades 

and which are actually xeric limestone prairies (Baskin and 

Baskin, 2004; Baskin and others, 2007a). Forest openings in 

the Ozark region especially have commonly been referred to 

as “glades” (Erickson and others, 1942; Heikens, 1999; Nelson 

and others, 2013), even though limestone glades—as defined 

by Baskin and Baskin (1999)—are not present west of the 

Mississippi River and Ozark “glades” are generally classified 

as xeric limestone prairies by those researchers who explicitly 

differentiate xeric limestone prairies from limestone cedar 

glades; see, for example, Lawless and others (2006a). This 

differentiation rests on a number of topographic, floristic, and 

ecological distinctions between the two ecosystem types (see 

table C–1 in chapter C). In some cases, however, limestone 

cedar glades can be found adjacent to or as microhabitats 

within xeric limestone prairies, as areas of relatively shallow 

soil depth and microsite dominance of C4 summer annual 

grasses such as Sporobolus vaginiflorus var. vaginiflorus

(Baskin and Baskin, 2000). 

The dolomite “glades” of Bibb County, Alabama 

(fig. D–3), which are present over the Ketona Dolomite, 

may represent a special sub-class of xeric limestone prairies 

(Baskin and Baskin, 2003). Referred to here as the Ketona 

dolomite glades, these forest openings correspond to the 

Alabama Ketona Glade and Woodland ecological system 

according to the International Terrestrial Ecological Systems 

Classification (ITESC). When these communities were first 

described, Allison (1994) and Allison and Stevens (2001) 

asserted that they were distinct from the limestone cedar 

glades of the Inner Central Basin of Middle Tennessee in a 

number of ways, including topography (the Ketona dolomite 

glades develop on steep slopes) and vegetation (dominance 

by Schizachyrium scoparium and relative scarcity of 

S. vaginiflorus var. vaginiflorus), despite the presence in the 

Ketona dolomite glades of two species otherwise believed to 

Figure D–3. Ketona dolomite glade, Kathy Stiles Freeland Bibb County Glades Preserve, Bibb County, Alabama. Photograph 

by Alan M. Cressler, U.S. Geological Survey.
42 Insular Ecosystems of the Southeastern United States

be endemics to limestone cedar glades: Leavenworthia exigua

var. lutea and Pediomelum subacaule. In addition, the Ketona 

dolomite glades contain several endemic species of their own 

(not present in limestone cedar glades or in any other outcrop 

community) and do not contain most of the limestone cedar 

glade endemics, leading Allison and Stevens (2001) and 

Baskin and Baskin (2003) to conclude that the Ketona dolomite glades are distinct from limestone cedar glades. Baskin 

and Baskin (2003) argued that the topographic and vegetation 

characteristics of the Ketona dolomite glades are consistent 

with a classification as xeric limestone prairies. Allison and 

Stevens (2001) acknowledged that the Ketona dolomite 

glades show greatest similarity to the xeric limestone prairie 

category; however, they pointed to three lines of evidence to 

argue that a new category should be created specifically to 

accommodate the Ketona dolomite glades: (1) they contain 

multiple endemics not found in (other) xeric limestone 

prairies, (2) they support two species of Leavenworthia, and 

(3) they develop over dolomite rather than limestone. Both 

arguments have merit, and certainly if the Ketona dolomite 

glades are categorized as xeric limestone prairies, then 

attention should be paid to their special characteristics and 

their suite of endemics that are not present on other xeric 

limestone prairies. 

Physical Geography

Throughout their geographic range, xeric limestone 

prairies share characteristics related to bedrock geology, soils, 

and topography. The calcareous bedrock substrates of xeric 

limestone prairies are of a wide range of ages, from Cambrian 

through Mississippian, and possibly Pennsylvanian and even 

Eocene (Lawless and others, 2006a). Xeric limestone prairies 

generally develop on rocky soils (alfisols, ultisols, mollisols, 

inceptisols, and vertisols) that are less than approximately 

3 feet (about 1 meter) thick and circumneutral to slightly 

alkaline (Homoya, 1994; Lawless and others, 2006a; Nelson 

and others, 2013). Xeric limestone prairies generally have 

southern to western aspects (Heikens and Robertson, 1995; 

Laughlin and Uhl, 2003; Nelson and others, 2013), although 

a few in the Ozark Plateaus are present on north-facing slopes 

(Theo Witsell, Arkansas Natural Heritage Commission, written 

commun., 2014). Rhoades and others (2004, 2005) asserted 

that xeric limestone prairies in the Knobs region of northern 

and central Kentucky (see fig. 1 in Rhoades and others, 2004) 

are confined to a narrow band of elevation, corresponding to a 

sedimentary layer of relatively soft calcareous shale between 

outcrops of harder dolomite. Edaphic conditions in xeric limestone prairies are controlled in part by the interaction of soil 

texture with topographic position. Soils tend to be fine-textured 

clay and silty clay loams. This texture, combined with the slope 

of xeric limestone prairies, which ranges from moderate to 

steep, makes these soils prone to erosion at rates that exceed 

soil genesis (Ware, 2002; Lawless and others, 2004, 2006a).

Stress and Disturbance Regimes

The factors most often cited as contributing to the 

origin and maintenance of xeric limestone prairies include 

geo-edaphic conditions (calcareous substrates, shallow 

soils influenced by erosion, limitation of certain important 

nutrients), topography (moderate to steep slopes with southern 

to western aspects), hydrologic conditions, and disturbance 

regimes based on fire and grazing (Homoya, 1994; Ludwig, 

1999; Baskin and Baskin, 2000; Laughlin and Uhl, 2003; 

Lawless and others, 2006a). Although the relative importance 

of these factors has been a subject of debate (Cipollini and 

others, 2005; Lawless and others, 2006a), in many cases, xeric 

limestone prairies likely have been maintained by complex, 

site-specific interactions of several different factors. 

Edaphic stress in xeric limestone prairies typically involves 

seasonally extreme hydrologic conditions (Kucera and Martin, 

1957). Seasonally xeric soil conditions are produced by the 

interaction of soil and landscape properties, including soil 

shallowness, abundance of rock fragments causing excessive 

drainage, relatively high insolation, and typically southern to 

western aspects (Laughlin and Uhl, 2003; Ware, 2010). The steep 

slope of some xeric limestone prairies also limits infiltration 

of precipitation because of fast runoff (Homoya, 1994). These 

factors suggest that seasonal water limitation and susceptibility 

to drought may help retard woody encroachment and thus 

maintain herbaceous habitat (Heikens and Robertson, 1995). 

On the other hand, many xeric limestone prairies experience seasonal saturation and frost-heaving of the thin soils 

from late fall to early spring, which are factors that may help 

exclude species vulnerable to winter rotting (Erickson and 

others, 1942; Hall, 1955; Kucera and Martin, 1957; Heikens 

and Robertson, 1994; Nelson and others, 2013). In some 

xeric limestone prairies, thinly bedded and relatively porous 

limestone and dolomite strata are underlain by impermeable 

strata, causing percolation to be impeded and redirected 

laterally, such that areas of rock outcrop may have abundant 

seepage during the winter and spring (Erickson and others, 

1942; Ludwig, 1999; Allison and Stevens, 2001; Ware, 2002). 

In many ways, these seasonal patterns of extreme hydrology 

mirror those occurring in limestone cedar glades (Quarterman, 

1950; Taylor and Estes, 2012). In contrast to limestone cedar 

glades, however, xeric limestone prairie communities contain 

relatively few winter-annual species—an adaptation to xeric 

summer conditions in which plants are dormant during the 

driest part of the year—suggesting that warm-season water 

stress may be less severe in xeric limestone prairies than in 

limestone cedar glades (Lawless and others, 2004). 

A number of soil properties display steep gradients from 

xeric limestone prairie interiors to adjacent forests, suggesting 

that edaphic conditions may help determine the geographic 

location and extent of xeric limestone prairies (Kucera and 

Martin, 1957; Allison and Stevens, 2001; Belden and others, 

2003; Rhoades and others, 2004, 2005). For example, bulk 

density, silt content, pH, and calcium availability were found 
Chapter D 43

to be elevated in xeric limestone prairie soils relative to 

surrounding forest soils (Rhoades and others, 2005). In one 

study, several soil taxonomic components, including presence 

of a mollic epipedon and soil thickness of less than about 

20 inches (roughly 50 centimeters), could be used to predict 

the spatial distribution of habitat for threatened taxa from 

calcareous grasslands (Mann and others, 1999). A growing 

body of research suggests that xeric limestone prairie soils are 

nutrient-poor relative to surrounding forests (Homoya, 1994) 

and relative to other grass-dominated ecosystems in North 

America (Rhoades and others, 2004). In particular, a number 

of researchers have noted low potassium availability in xeric 

limestone prairie soils (Ludwig, 1999; Allison and Stevens, 

2001; Rhoades and others, 2004; Trammell and others, 2004).

Disturbance regimes are also important in xeric limestone 

prairies. In particular, many investigations of successional 

dynamics have emphasized the importance of fire (Heikens 

and Robertson, 1994; Batek and others, 1999; Heikens, 1999; 

Baskin and Baskin, 2000; Guyette and others, 2003). Important 

components of fire regimes include seasonality, frequency, and 

intensity. High-intensity fires can result in loss of nutrients 

from ecosystems by a variety of mechanisms, including direct 

oxidation and volatilization, erosion, convection of ash, and 

nutrient leaching; however, small and low-intensity fires from 

prescribed burns have been shown to have minimal effect on 

leaching (Trammell and others, 2004). In addition to the direct 

effects on woody vegetation, fire may also slow the pace of 

woody encroachment indirectly by exacerbating water stress 

in soils, both by increasing radiant heating and by removing 

moisture-retaining litter (Homoya, 1994).

Historical fire regimes that regulated succession dynamics 

in xeric limestone prairies were probably a combination 

of natural and anthropogenic fires; however, the relative 

importance of the two fire types has been debated (Sauer, 1950; 

Batek and others, 1999; Laughlin, 2004; Noss, 2013). A number 

of researchers, including Guyette and others (2003), have 

emphasized the importance of anthropogenic fire and argued 

that natural fires were rare in the regions where xeric limestone prairies are present because lightning strikes—though 

frequent—were usually accompanied by heavy precipitation. 

By contrast, Foti and Glenn (1991) indicated a high prevalence 

of present-day lightning-set fires in the summer and early 

autumn in the Ouachita Mountains. Analysis of fire scars 

has consistently shown high fire incidence during the period 

of Native American occupation and much less frequent fire 

during the past 100 years (Jenkins and others, 1997; Guyette 

and others, 2003; Laughlin, 2004). Batek and others (1999) 

and Heikens (1999) concluded that grassland-shrub mosaics 

in the Ozarks, which include xeric limestone prairies, were 

maintained to a large extent by anthropogenic fire. On the 

other hand, the high degree of endemism present in some 

xeric limestone prairies argues strongly against origin or 

maintenance solely or primarily by anthropogenic disturbance 

because it is highly unlikely that so many endemics could 

have evolved during the relatively short timespan of human 

occupation (Noss, 2013). Also, historical accounts suggest 

that xeric limestone prairies were the last remaining types of 

forest openings in the Ozark Plateaus Physiographic Province 

by the mid-1800s, after many of the native barrens and 

savannas had been degraded or succeeded to forest (Heikens, 

1999). A similar history has been proposed in southern 

Indiana (Guyette and others, 2003). If this history is true, it 

suggests that other factors beyond fire disturbance, possibly 

including topographic and edaphic factors, also play a role 

in maintaining xeric limestone prairies (Erickson and others, 

1942; Hall, 1955; Homoya, 1994). The topographic and edaphic 

similarity of xeric limestone prairies contributes to this line 

of reasoning: if fire was the sole factor needed to create and 

maintain xeric limestone prairies, then they should be found 

with similar frequency on level ground, on north-facing slopes, 

and in sites with nutrient-rich soil (Homoya, 1994). 

Grazing is another important aspect of the historical 

disturbance regime for many xeric limestone prairies. Prior to 

the establishment of agriculture in North America, grasslands 

were grazed by large herbivores such as elk and bison and 

later by domesticated herbivores such as sheep, cows, and 

horses (Lawless and others, 2006a). Thus, grazing is an 

ancient phenomenon that has been altered in recent times by 

human intervention. Some Native Americans were known 

to pasture cattle before European settlement, and they could 

have used some xeric limestone prairies for this purpose 

(Webb and others, 1997). Many xeric limestone prairies 

also show evidence of more recent pasturing. In the Knobs 

region of Kentucky, Rhoades and others (2004) noted fence 

remnants near the edges of some xeric limestone prairies. In 

northwestern Alabama, Webb and others (1997) stated that 

most forest openings not suitable for tillage—including xeric 

limestone prairies because of their thin soils—were used for 

grazing until the 1940s. Describing barrens and “glades” 

(probably xeric limestone prairies based on the reported 

dominance of C4

 perennial grasses), Bartgis (1993) noted that 

most of these areas had been heavily used for sheep pasture. 

In eastern Missouri, Erickson and others (1942) noted that 

most “glades” had been occupied by cattle and, to a lesser 

extent, hogs. As with fire regimes, the ecological effects of 

grazing depend on frequency and intensity. For example, light 

to moderate grazing typically produces mosaics of shrub and 

grassland that support higher levels of species diversity than 

do grasslands that result from high-intensity grazing (DeSelm 

and Murdock, 1993).

Historically, xeric limestone prairies likely have 

been maintained by complex interactions of disturbance 

regimes with edaphic and hydrologic conditions. These 

interactions may have involved a number of feedback loops. 

For example, in addition to the direct effects of drought in 

slowing or reversing woody encroachment, drought is also 

intrinsically linked to fire regimes in the absence of fire 

suppression (DeSelm and Murdock, 1993; Quarterman and 

others, 1993). In turn, fires can exacerbate existing dry soil 

conditions (Homoya, 1994) and nutrient limitation (Trammell 

and others, 2004). Nutrient limitation and water stress may 

act synergistically as a stress regime experienced by plants 
44 Insular Ecosystems of the Southeastern United States

because nutrient absorption is impeded by xeric conditions 

(Heikens and Robertson, 1995). Prairies created by these 

processes would presumably become attractive grazing sites 

to native herbivores and to pastoralists, and grazing in turn has 

the potential to alter plant community composition, vegetation 

dynamics, and soil physical properties (DeSelm and Murdock, 

1993; Nelson and others, 2013). Depending on topographic 

setting, removal of woody vegetation by fire and (or) grazing 

can make soils more prone to erosion (Trammell and others, 

2004; Lawless and others, 2006a), and soil depletion from 

erosion might retard woody encroachment by contributing to 

water and nutrient stress. Additionally, disturbance to adjacent 

forests or woodlands may expand existing prairie sites 

and change vegetation dynamics (Harper, 1920; Laughlin, 

2004). Thus, topography, hydrology, edaphic conditions, and 

disturbance events have probably interacted in highly complex 

ways to produce and maintain xeric limestone prairies, and the 

nature of these interactions has undoubtedly varied from one 

site to another.

Community Types and Vegetation 

Dynamics

Xeric limestone prairies are typically characterized 

by a relatively stable matrix of graminoids and a highly 

variable—and less dominant—component of forbs (Lawless 

and others, 2006a). Figure D–4 shows such a graminoid 

and forb community. The characteristic dominant species is 

Schizachyrium scoparium, which has been suggested as one 

criterion for the differentiation of xeric limestone prairies 

from limestone cedar glades (Quarterman, 1989; Baskin 

and Baskin, 2000; Lawless and others, 2006a, b). Other 

C4 perennial grasses common in xeric limestone prairies 

include Bouteloua curtipendula, Andropogon gerardii, 

Sorghastrum nutans, and Sporobolus clandestinus (Kucera 

and Martin, 1957; Lawless and others, 2006a). Although 

less dominant, C3

 perennial forbs (including Silphium 

terebinthinaceum, Monarda fistulosa, and some Cyperaceae 

Figure D–4. Dry Prairie, Coosa Valley Prairies, Floyd County, Georgia. Photograph by Alan M. Cressler, U.S. Geological Survey.
Chapter D 45

taxa) and C4 summer annual grasses (such as Sporobolus 

vaginiflorus and S. neglectus) may be locally abundant at some 

sites (Kucera and Martin, 1957; Baskin and Baskin, 2000; 

Lawless and others, 2006a). Typically, xeric limestone prairies 

support assemblages containing geographically restricted 

taxa and geographically widespread disturbance-adapted taxa 

(Ware, 2010). Multiple microhabitats within xeric limestone 

prairies generally support characteristic plant assemblages, 

and community zonation may be strongly controlled by soil 

depth gradients (Ware, 2002). For example, C4 summer annual 

grasses are often restricted to areas with the shallowest soil 

(Baskin and Baskin, 2000; Rhoades and others, 2004; Lawless 

and others, 2006b), leading Baskin and Baskin (2000) to 

assert that “cedar glades of the type described for the southeastern United States... occur in microhabitats within xeric 

limestone prairies.” 

Xeric limestone prairies as an ecosystem category, as 

defined by Lawless and others (2004), do not correspond neatly 

to classification according to the ITESC and the U.S. National 

Vegetation Classification (USNVC) and include a number of 

component associations of the ecological systems Ozark Prairie 

and Woodland, Eastern and Western Highland Rim Prairie 

and Barrens, Pennyroyal Karst Plain Prairie and Barrens, 

Central Interior Highlands Calcareous Glade and Barrens, and 

Southern Ridge and Valley Calcareous Glade and Woodland 

(NatureServe, 2015). Differing terminologies and attitudes 

toward ecosystem change have existed among researchers 

studying these prairies. Some have described increased cover 

of woody species as “encroachment,” “invasion,” and “degradation” (Laughlin, 2004; Sutter and others, 2011), whereas others 

described it simply as succession (Baskin and others, 2007b). 

Regardless, a number of different methods (such as long-term 

monitoring based on field visits, comparison of modern and 

historical floristic lists, and methods using aerial photography 

and remote sensing) have documented the shrinking size of 

xeric limestone prairies across their geographic range, with a 

strong association between the increase of woody species and 

declines in species diversity and in populations of endemic 

species (Kimmel and Probasco, 1980; Annala and Kapustka, 

1983; Lowell and Astroth, 1989; Heikens and Robertson, 1995; 

Laughlin and Uhl, 2003; Mills, 2008; Young and others, 2009; 

Sutter and others, 2011). 

Some studies have also explicitly linked woody encroachment with changes in disturbance regimes. In the Ozarks, for 

example, fire intervals increased from 4 years in the 1870s 

to 22–24 years by the end of the 20th century, with the areal 

extent of xeric limestone prairies decreasing by roughly 

60 percent over the same time period (Heikens, 1999; Comer 

and others, 2011). Other investigations have found perimeters 

of xeric limestone prairies to be relatively stable over the 

course of decades. Rhoades and others (2004) attributed the 

relative lack of woody encroachment onto forest openings 

in the Knobs region of Kentucky over a 75-year interval 

(evidenced by aerial photography) to edaphic differences 

between these openings and the surrounding forest, but also 

noted evidence of past grazing. Also using aerial photography, 

Ludwig (1999) found limited encroachment onto 12 xeric 

limestone prairies in extreme southwestern Virginia over a 

33-year interval and attributed this stability to a combination 

of drought stress and fire. 

The relative importance of edaphic constraints versus 

disturbance to ecosystem maintenance has implications for 

the speed and geographic extent of woody encroachment 

near the edges of xeric limestone prairies (Heikens and 

Robertson, 1995; Trammell and others, 2004). In the absence 

of disturbance, succession in xeric limestone prairies may not 

be uniform given the coexistence of multiple heterogeneous 

microhabitats. Some areas within xeric limestone prairies may 

be edaphically constrained more than others and may succeed 

more slowly due to edaphic or hydrologic constraints (DeSelm 

and others, 1969; Trammell and others, 2004). At a regional 

scale, climatic differences may control succession rates as well. 

For example, succession in Ozark Plateaus xeric limestone 

prairies might be more constrained by dry summer conditions 

compared to Valley and Ridge prairies that experience greater 

precipitation during the growing season (Lawless and others, 

2006a; Sutter and others, 2011).

In the Ozark Plateaus, the primary woody species 

involved in encroachment onto xeric limestone prairies include 

Juniperus virginiana, Quercus marilandica, Carya texana, 

and Acer saccharum (Heikens, 1999). In southern Ohio, the 

primary woody species are J. virginiana, Pinus virginiana, 

Cercis canadensis, and various Quercus species (Annala and 

Kapustka, 1983). In their flow diagrams of succession pathways for limestone cedar glades, Baskin and others (2007b) 

conceptualized xeric limestone prairies as disturbancemediated alternatives to the seral stage of redcedar (J. virginiana) forests: the prior seral stage (shrub-redcedar) would be 

followed by xeric limestone prairie if the area was subjected to 

disturbance (grazing or fire) but in the absence of disturbance 

would be followed by redcedar forest, which would then be 

followed by redcedar/hardwood forest and would eventually 

become mixed hardwood forest. The presence of woody exotic 

invasive species such as Ligustrum sinense (Morris and others, 

2002; Sutter and others, 2011) complicates the succession 

pathways described by Baskin and others (2007b) because 

in the absence of disturbance, xeric limestone prairies may 

become dominated as much by L. sinense as by J. virginiana. 
46 Insular Ecosystems of the Southeastern United States

Contributions to Regional Biodiversity

Throughout their range, xeric limestone prairies provide 

habitat for globally rare taxa (see table 3–3 in appendix 3) 

(Webb and others, 1997; Laughlin and Uhl, 2003; Rhoades 

and others, 2005). A number of regionally distinct and globally 

rare plant associations also are present within these ecological 

systems, including Ozark Limestone Glades, Ozark Dolomite 

Glades, Ozark Annual Grass Glade, Ozark Ashe’s Juniper 

Glade Woodland, Moulton and Tennessee Valley Limestone 

Hill Barrens, Southern Limestone Glade Margin Shrubland, 

and Outer Bluegrass Dolomite Glade (Nelson and others, 

2013). Several federally listed plant taxa are found in xeric 

limestone prairies, including Dalea foliosa, which is federally listed as endangered and has been documented in xeric 

limestone prairies associated with limestone cedar glades, and 

Lesquerella filiformis, which is present in the Ozark Highlands 

and is listed as threatened (Thompson and others, 2006; Young 

and others, 2009; Nelson and others, 2013). The Ketona 

dolomite glades of Bibb County, Alabama, are exceptional as 

habitat for rare species. Spigelia gentianoides var. alabamensis

is federally listed as endangered and is present only in the 

Ketona dolomite glades; Marshallia mohrii is federally 

listed as threatened and also is present in other disturbancemaintained herbaceous habitats (Allison and Stevens, 2001). 

A total of more than 60 plant taxa of conservation concern are 

on or near these small and isolated forest openings that are 

believed to cover only about 250 acres (roughly 100 hectares) 

total, making them “one of the most significant reservoirs of 

botanical diversity in the eastern United States” (Allison and 

Stevens, 2001). Figure D–5 shows several rare plant taxa in 

Ketona dolomite glade habitat.

Endemic and Biogeographically Disjunct Taxa

Lawless and others (2006a) classified 13 taxa as being 

endemic or near-endemic to xeric limestone prairies, including 

8 in the Ketona dolomite glades (Allison and Stevens, 2001; 

Alabama Natural Heritage Program, 2007), 4 from the Ozark 

Plateaus of Arkansas and Missouri, and 1 from the Valley and 

Ridge of Virginia and West Virginia. In addition to these taxa, 

Collins and Wieboldt (1992) described a new clover species 

(Trifolium calcaricum) found at only a few locations in xeric 

limestone prairies in southwestern Virginia and in Middle 

Tennessee. Kral and Moffett (2009) described a new speces, 

Xyris spathifolia, which has so far been documented only 

in Ketona dolomite glades. The Ketona dolomite glades also 

support populations of Solanum carolinense var. hirsutum, 

which was previously believed to be extinct (Allison and 

Stevens, 2001).

Near-endemic species of Appalachian shale barrens 

and of limestone cedar glades (such as Dalea gattingeri, 

Echinacea tennesseensis, and several varieties of Leavenworthia exigua) have also been found in xeric limestone prairies, with dispersal from cedar glades often facilitated by their 

geographic proximity to xeric limestone prairies (Lawless and 

others, 2006a). Other factors likely affecting the distribution 

of rare and endemic plants in xeric limestone prairies include 

topography, soil depth, size of forest openings (Skinner and 

others, 1983), competition with mesic species, and herbivory 

by insects (VanZandt and others, 2005). Although characteristic 

plant taxa of xeric limestone prairies do not necessarily show 

marked disjunction in their distributions, they do tend to have 

affinities with midwestern, western, and southwestern taxa, 

and for many species, xeric limestone prairies represent range 

extensions far to the east (Lawless and others, 2006a). 

Animal Taxa of Conservation Concern

Xeric limestone prairies also provide eastern range 

extensions for a number of western and southwestern animal 

species, including Aphonopelma hentzi8 (Texas brown tarantula), Crotaphytus collaris (collared lizard), and Geococcyx 

californianus (greater roadrunner) (Kimmel and Probasco, 

1980). Such range extensions contribute to regional faunal 

biodiversity by supporting taxa that are of local or regional 

conservation concern even though they are abundant in 

western regions. For example, whereas the collared lizard is 

globally secure within its range, it is ranked S3 in Arkansas 

(NatureServe, 2015) and is considered a species of conservation concern by the Arkansas Natural Heritage Commission 

(2014). Xeric limestone prairies provide important habitat to 

a number of other reptile and amphibian species, including 

Aspidoscelis sexlineata (six-lined racerunner), Plethodon 

glutinosus (northern slimy salamander), Plestiodon anthracinus (southern coal skink), Lampropeltis getula (eastern 

kingsnake), L. triangulum syspila (red milksnake), Sonora 

semiannulata (groundsnake), Lithobates sphenocephalus

(southern leopard frog), Terrapene carolina (eastern box 

turtle), and Sceloporus undulatus (eastern fence lizard) 

(Kimmel and Probasco, 1980; Angert and others, 2002; 

VanZandt and others, 2005; Nelson and others, 2013). A 

number of grassland- and shrubland-associated birds have 

also been documented in xeric limestone prairies, including 

Peucaea aestivalis (Bachman’s sparrow), Spizella pusilla

(field sparrow), Setophaga discolor (prairie warbler), 

Setophaga cerulea (cerulean warbler), Vermivora cyanoptera

(blue-winged warbler), Geothlypis formosa (Kentucky 

warbler), Passerina ciris (painted bunting), and Toxostoma 

rufum (brown thrasher) (Kimmel and Probasco, 1980; Hardin 

and others, 1982; Comer and others, 2011; Nelson and others, 

2013). In the Ozarks, xeric limestone prairies support several 

rare invertebrates, including Trimerotropis saxatilis (lichen 

grasshopper). The Ketona dolomite glades support populations 

of Argyria rufisignella 9

, a rare moth species (Martinez and 

Brown, 2007), as well as several species of grasshoppers 

normally found in sandy habitats (Hill, 2007).

8

Species name according to the Integrated Taxonomic Information 

System (2015).

9

Species name according to Encyclopedia of Life (2016).
Chapter D 47

Figure D–5. Rare plant taxa in Ketona dolomite glades at the Kathy Stiles Freeland Bibb County Glades Preserve, Bibb County, 

Alabama: A, Dalea cahaba; B, Erigeron strigosus var. dolomiticola; C, Onosmodium decipiens; D, Spigelia gentianoides. Photographs 

by Alan M. Cressler, U.S. Geological Survey.

Conservation Considerations

In several states, xeric limestone prairies have been 

identified as key priorities for conservation and restoration 

(Laughlin, 2004; Cipollini and others, 2005; Nelson and 

others, 2013). Conservation of these ecosystems involves 

several considerations, including assessment of current threats, 

evaluation of various conservation strategies, and identification of gaps in scientific knowledge.

Threats to Ecosystem Integrity

Woody encroachment into xeric limestone prairies, 

largely attributable to fire suppression, is regarded by many 

researchers to be the primary threat to ecosystem health 

(Harper, 1920; Kimmel and Probasco, 1980; Annala and 

Kapustka, 1983; Skinner and others, 1983; Lowell and 

Astroth, 1989; Bartgis, 1993; Ware, 2002; Comer and 

others, 2011). Many researchers have recommended the 

reintroduction of fire into xeric limestone prairies in the 

form of prescribed burns to prevent woody encroachment, 

as well as the targeted removal of invasive exotic species and 

(or) the woody overstory if necessary (Lowell and Astroth, 

1989; Laughlin, 2004; Jenkins and Jenkins, 2006; Duncan 

and others, 2008; Comer and others, 2011; Sutter and others, 

2011). Woody invasive exotics mentioned as possible targets 

for removal include Ligustrum sinense, L. vulgare, Elaeagnus 

umbellata, Lonicera maackii, L. tatarica, and Rhamnus 

A B

C D
48 Insular Ecosystems of the Southeastern United States

cathartica, and herbaceous exotics include Coronilla varia, 

Daucus carota, Lonicera japonica, Lespedeza cuneata, 

Bromus species, and Verbascum thapsus (Laughlin and Uhl, 

2003; Laughlin, 2004; Sutter and others, 2011; Nelson and 

others, 2013). The shrinking of xeric limestone prairies by 

woody encroachment has implications not just for herbaceous 

endemic plant species, but also for grassland-associated birds, 

some of which are species of conservation concern (Kimmel 

and Probasco, 1980; Hardin and others, 1982; Comer and 

others, 2011; Nelson and others, 2013). For example, fire 

disturbance appears to be important in maintaining habitat 

for Bachman’s sparrow, which has experienced noticeable 

population declines (Hardin and others, 1982; Meyer, 2006). 

Land-use change, including conversion to agriculture 

and overgrazing, has been mentioned as a threat to some 

xeric limestone prairies (Hardin and others, 1982; Lowell and 

Astroth, 1989; Webb and others, 1997; Heikens, 1999; Rhoades 

and others, 2004; Young and others, 2009). In addition to 

direct effects on plant community composition, overgrazing 

can have indirect effects that ripple through the ecosystem, 

including soil compaction and disruption of interspecific interactions between plants and insects (Nelson and others, 2013). 

Some forms of resource extraction may also pose threats, 

including quarrying and logging (Bartgis, 1993; Webb and 

others, 1997; Heikens, 1999; Laughlin, 2004). Some sites have 

also been damaged by recreational use (Webb and others, 1997; 

Nelson and others, 2013). Increasingly, conservation officials 

have noted feral hogs to be a threat to some xeric limestone 

prairies in the Ozarks, causing erosion and reduction in plant 

and animal biodiversity (Nelson and others, 2013).

Conservation Strategies

Throughout their geographic range, a number of xeric 

limestone prairies are present on protected land. In southern 

Ohio, prairies are protected on the Richard and Lucile Durell 

Edge of Appalachia Preserve (Annala and Kapustka, 1983; 

Cipollini and others, 2005), and in Indiana they are within 

the Hoosier National Forest (Guyette and others, 2003). In 

Kentucky, the Crooked Creek Barrens Preserve (Rhoades and 

others, 2005), the Jim Scudder and Thompson Creek Glade 

nature preserves, the Fort Knox Military Reservation (Baskin 

and others, 1997), and the Bernheim Arboretum and Research 

Forest (Rhoades and others, 2004) all contain xeric limestone 

prairies. Most of the Ketona dolomite glades of Bibb County, 

Alabama, are now protected by the Nature Conservancy on its 

Kathy Stiles Freeland Bibb County Glades Preserve (Allison 

and Stevens, 2001). In the Ozarks of Missouri and Arkansas, 

xeric limestone prairies are present on numerous publicly 

protected sites, including Wilson’s Creek National Battlefield 

(Young and others, 2009), Mark Twain National Forest (Kucera 

and Martin, 1957), Washington University’s Tyson Research 

Center, Meramec State Park (VanZandt and others, 2005), the 

Lower Buffalo Wilderness Area of the Buffalo National River 

(Jenkins and Jenkins, 2006), Taum Sauk Mountain State Park, 

Butler Hollow Glade Natural Area (Nelson and others, 2013), 

and other protected sites.

Because woody encroachment is widely recognized as a 

major threat to xeric limestone prairies, many conservation 

programs involve attempts to preserve herbaceous habitat by 

mimicking natural disturbance regimes by using prescribed 

burns or other means of removing woody vegetation. A 

number of successful restoration projects have been described 

and empirically examined. Experimental prescribed burns in 

the Ketona dolomite glades and surrounding longleaf pine 

forest in Alabama produced considerable reduction in woody 

encroachment without any long-term detrimental effects 

on rare or endemic herbaceous species (Duncan and others, 

2008). Prescribed burns in xeric limestone prairies in northern 

Arkansas also resulted in increased cover of prairie grasses 

and increased diversity of ground flora species, with increased 

diversity driven by increased evenness rather than increased 

richness (Jenkins and Jenkins, 2006). Following prescribed 

burns in Ozark xeric limestone prairies in Missouri, Comer 

and others (2011) noted a substantial shift in bird community 

structure toward grass-shrubland species, despite relatively 

subtle changes in vegetation structure. After the mechanical 

removal of Juniperus virginiana, Young and others (2009) 

documented a rapid response of the rare winter annual 

Lesquerella filiformis. In addition to removal of woody and 

exotic species, some xeric limestone prairie restoration projects 

have also involved planting of glade endemics, a practice that 

is most effective when juvenile plants are transplanted because 

seedling mortality is generally high and germination success 

low (VanZandt and others, 2005).

Although understanding of pre-settlement fire regimes 

is incomplete, evidence from tree-scar analysis suggests 

that prescribed burn practices will best mimic historical fire 

regimes if they incorporate spatially heterogeneous burning, 

low- to moderate-intensity fires, and variability in fire intervals 

(Guyette and others, 2003). Noting that some rare herbaceous 

species are able to colonize multiple microhabitats within 

xeric limestone prairies, Young and others (2009) recommended that management actions should be microhabitatspecific, for example by targeting burning to areas of heavy 

woody cover while focusing on reducing trampling at locations of shallower soil near outcrops. Different microhabitats 

for different rare xeric limestone prairie herbaceous species, as 

documented by Skinner and others (1983), also highlight the 

need to focus on microhabitat diversity as a part of management and conservation strategies. In assessing ecological 

effects of prescribed burning, Jenkins and Jenkins (2006) point 

out that unburned plots need to be included in experimental 

designs and analysis, rather than relying solely on pre-burn 

versus post-burn comparisons.

Additionally, restoration efforts to promote herbaceous 

habitat in xeric limestone prairies need to be based in a 

thorough understanding of trophic interactions, including the 

effects of herbivores and carnivores within the ecosystem. 

Specifically, Van Zandt and others (2005) recommended that 

when woody species are removed, downed logs, brush piles, 
Chapter D 49

and other similar structures should be preserved as refugia for 

carnivorous lizards that indirectly protect herbaceous plants 

by feeding upon insect herbivores. This type of conservation 

planning based on expected trophic interactions can be 

complicated, however, involving an accounting for multiple 

factors governing the population structures of insectivorous 

lizards in xeric limestone prairies, such as competition, 

predation, thermal biology, and differential microhabitat use 

(Angert and others, 2002).

Knowledge Gaps

The conservation of xeric limestone prairies is rarely 

straightforward, for a number of reasons. The origins of these 

prairies and the relative importance of factors involved in their 

maintenance are still debated. The ambiguous definition of 

ecological integrity for xeric limestone prairies complicates 

conservation decisionmaking and prioritization. Effective 

management and conservation of fire-maintained ecosystems 

require thorough understanding of the relevant historical 

disturbance regimes (Guyette and others, 2003); however, 

the timing, frequency, and intensity of fire regimes predating 

European settlement are not fully understood. Research efforts 

are often complicated by problems such as the disappearance 

of fire evidence from decomposition in humid climates, the 

difficulty in finding undisturbed lake sediments for analysis, 

and inherent biases in fire-scar analysis of trees (Guyette 

and others, 2003; Laughlin and Uhl, 2003; Laughlin, 2004; 

Cipollini and others, 2005). Also lacking is a thorough 

understanding of how fire in xeric limestone prairies affects 

population dynamics of animal species (Heikens, 1999) and 

of rare and endemic plants (Duncan and others, 2008), or 

of the interaction between fire regime and topography and 

substrate (Batek and others, 1999). 

Possible Ecological Effects from Climate Change

A systematic climate-change vulnerability analysis across 

the full range of xeric limestone prairie community types has 

yet to be conducted; however, some species-level analyses 

have included taxa associated with xeric limestone prairies. 

For example, Davenport (2007) predicted that Dalea foliosa

and Spigelia gentianoides “would suffer greatly under drier 

conditions and perhaps disappear entirely.” Molano-Flores and 

Bell (2012) used count-based population viability analysis to 

project population dynamics of D. foliosa based on various 

climate-change scenarios. Warmer winter temperatures 

and associated reductions in snow cover were projected to 

contribute to population declines, which was attributed to 

diminished insulation under cold conditions and premature 

snowmelt-triggered dehardening, leading to root injury and 

seedling mortality. This study used approximately one decade’s 

worth of population data, such that long-term climate-related 

ecosystem-level effects on habitat (such as changes in 

disturbance regime, changes in woody cover, or changes in 

soil physical or chemical properties) likely were not accounted 

for. Also, the data were obtained from a single site near the 

northern limit of the species range, so applicability of findings 

to other D. foliosa populations is unclear.

Such analyses may be useful in forecasting population 

dynamics based on the interactions of climate change and 

species-level physiology (for example, seedling mortality 

increased by early snowmelt), but cannot account for species 

vulnerability to climate change that derives primarily from 

habitat degradation. Habitat-based vulnerability may be 

important, however, because analysis of microhabitat characteristics of D. foliosa suggests this species is sensitive to woody 

encroachment and to competition with other herbaceous 

plants (Thompson and others, 2006). Thus, an analysis 

of climate-related habitat alteration would likely improve 

predictive power of models for climate-change effects on 

populations of this rare plant—and others with similar habitat 

needs—present in xeric limestone prairies.

Predicting ecosystem-level climate-change effects for 

xeric limestone prairies is complicated by ambiguity over the 

relative importance of disturbance regimes and of topo-edaphic 

stress factors, as well as by the complex interactions between 

the two. In the absence of human interference, climate is 

generally a strong control on fire intervals (Guyette and others, 

2012). If climate change produces higher summer temperatures, drier summer conditions, and (or) more frequent or 

longer droughts, then natural fire frequency and intensity would 

be expected to increase. Below certain temperature-dependent 

precipitation thresholds, however, dry conditions can limit fuel 

availability (Guyette and others, 2012). If drought intensification reduced woody and (or) herbaceous cover within xeric 

limestone prairies or their immediate surroundings, reduced 

fuel loads could limit fire frequency and intensity. Additionally, 

changes in seasonal storm frequency could be an important 

factor to the degree that lightning strikes are an ignition source. 

Thus, fire-regime effects from climate change—even under 

an assumption of drought intensification—might be difficult 

to predict. Of course, for many xeric limestone prairies, this 

would be a moot point because fire regimes have already been 

highly modified through fire suppression and (or) controlled 

burning (Kucera and Martin, 1957; Ware, 2002; Comer and 

others, 2011; Nelson and others, 2013), implying that the 

possibility for climate change to alter ecosystem dynamics 

through fire-regime alteration has been greatly diminished. 

Even in the absence of fire-regime effects, regional 

changes in precipitation timing, frequency, and intensity 

could produce ecological changes in xeric limestone prairies. 

As with limestone cedar glades, plant taxa associated with 

seasonally wet microhabitats (Taylor and Estes, 2012) and 

those with life histories timed to historical seasonal patterns 

in soil moisture availability (Quarterman, 1950; Ware, 2002) 

might be especially sensitive to changes in the seasonal 

timing of precipitation inputs. Additionally, changes in 

storm intensity, duration, or seasonality could have implications for soil erosion dynamics in xeric limestone prairies. 

Thornberry-Ehrlich (2012) discussed the potential for greater 
50 Insular Ecosystems of the Southeastern United States

erosion in Stones River National Battlefield in Rutherford 

County, Tennessee—a site containing xeric limestone prairies 

closely associated with limestone cedar glades—resulting 

from predicted increases in storm frequency and intensity. 

The degree to which changes in storm dynamics might be 

associated with increased erosion in xeric limestone prairies 

would likely be based on a number of interrelated and sitespecific factors including soil texture, slope, and vegetation 

cover (Ware, 2002; Lawless and others, 2006a) as well as 

disturbance history (Trammell and others, 2004). 

References

Alabama Natural Heritage Program, 2007, Alabama inventory list—The rare, threatened and endangered plants and 

animals of Alabama: Montgomery, Ala., Alabama Natural 

Heritage Program.

Allison, J., 1994, A botanical “lost world” in central Alabama, 

in Proceedings of the North American Conference on 

Savannas and Barrens, Environmental Protection Agency, 

Great Lakes National Program Office, Chicago, Ill., 

p. 323–327.

Allison, J., and Stevens, T., 2001, Vascular flora of Ketona 

dolomite outcrops in Bibb County, Alabama: Castanea, 

v. 66, no. 1, p. 154–205.

Angert, A., Hutchison, D., Glossip, D., and Losos, J., 2002, 

Microhabitat use and thermal biology of the collared lizard 

(Crotaphytus collaris collaris) and the fence lizard (Sceloporus undulatus hyacinthinus) in Missouri glades: Journal 

of Herpetology, v. 36, no. 1, p. 23–29.

Annala, A., and Kapustka, L., 1983, Photographic history of 

forest encroachment in several relict prairies of the Edge 

of Appalachia preserve system, Adams County, Ohio: 

Ohio Journal of Science, v. 83, no. 3, p. 109–114.

Arkansas Natural Heritage Commission, 2014, Rare species 

search engine—Find Arkansas species of conservation 

interest, accessed November 18, 2015, at http://www.natural

heritage.com/Research-and-Data/rare-species-search.

Bartgis, R., 1993, The limestone glades and barrens of West 

Virginia: Castanea, v. 58, no. 2, p. 69–89.

Baskin, J., and Baskin, C., 1999, Cedar glades of the southeastern United States, in Anderson, R., Fralish, J., and 

Baskin, J., eds., Savannas, barrens, and rock outcrop plant 

communities of North America: New York, Cambridge 

University Press, p. 206–219.

Baskin, J., and Baskin, C., 2000, Vegetation of limestone and 

dolomite glades in the Ozarks and Midwest regions of the 

United States: Annals of the Missouri Botanical Garden, 

v. 87, no. 2, p. 286–294.

Baskin, J., and Baskin, C., 2003, The vascular flora of cedar 

glades of the southeastern United States and its phytogeographical relationships: Journal of the Torrey Botanical 

Society, v. 130, no. 2, p. 101–118.

Baskin, J., and Baskin, C., 2004, History of the use of 

“cedar glades” and other descriptive terms for vegetation 

on rocky limestone soils in the Central Basin of Tennessee: 

The Botanical Review, v. 70, no. 4, p. 403–424.

Baskin, J., Baskin, C., and Chester, E., 1999, The Big Barrens 

Region of Kentucky and Tennessee, in Fralish, J., and 

Baskin, J., eds., Savannas, barrens, and rock outcrop plant 

communities of North America: New York, Cambridge 

University Press, p. 190–205.

Baskin, J., Baskin, C., and Lawless, P., 2007a, Calcareous rock 

outcrop vegetation of eastern North America (exclusive of 

the Nashville Basin), with particular reference to use of the 

term “cedar glades”: The Botanical Review, v. 73, no. 4, 

p. 303–325.

Baskin, J., Baskin, C., and Quarterman, E., 2007b, Flow 

diagrams for plant succession in the Middle Tennessee 

cedar glades: Journal of the Botanical Research Institute 

of Texas, v. 1, no. 2, p. 1131–1140.

Baskin, J., Chester, E., and Baskin, C., 1997, Special paper—

Forest vegetation of the Kentucky karst plain (Kentucky 

and Tennessee)—Review and synthesis: Journal of the 

Torrey Botanical Society, v. 124, no. 4, p. 322–335.

Batek, M., Rebertus, A., Schroeder, W., Haithcoat, T., 

Compas, E., and Guyette, R., 1999, Reconstruction of 

early nineteenth-century vegetation and fire regimes in the 

Missouri Ozarks: Journal of Biogeography, v. 26, no. 2, 

p. 397–412.

Belden, A., Chazal, A., and Hobson, C., 2003, A natural 

heritage inventory of fourteen headwater sites in the Dragon 

Run Watershed: Richmond, Va., Virginia Department of 

Conservation and Recreation.

Cipollini, K., Maruyama, A., and Zimmerman, C., 2005, 

Planning for restoration—A decision analysis approach to 

prioritization: Restoration Ecology, v. 13, no. 3, p. 460–470.

Collins, J., and Wieboldt, T., 1992, Trifolium calcaricum

(Fabaceae), a new clover from limestone barrens of eastern 

United States: Castanea, v. 57, no. 4, p. 282–286.
Chapter D 51

Comer, C., Bell, A., Oswald, B., Conway, W., and Burt, D., 

2011, Vegetation and avian response to prescribed fire 

on glade habitats in the Missouri Ozarks: The American 

Midland Naturalist, v. 165, no. 1, p. 91–104.

Copenheaver, C., Kyle, K., Stevens, G., and Kamp, M., 2005, 

Comparing Juniperus virginiana tree-ring chronologies 

from forest edge vs. forest interior positions in the Cedars 

Natural Area Preserve in Virginia, USA: Dendrochronologia, v. 23, no. 1, p. 39–45.

Davenport, L., 2007, Climate change and its potential effects 

on Alabama’s plant life: Birmingham, Ala., Samford 

University, Vulcan Materials Center for Environmental 

Stewardship and Education.

DeSelm, H., and Murdock, N., 1993, Grass-dominated 

communities, in Martin, W., Boyce, S., and Echternacht, A.,

eds., Biodiversity of the southeastern United States—

Upland terrestrial communities: New York, John Wiley 

and Sons, Inc., p. 87–141.

DeSelm, H., Whitford, P., and Olson, J., 1969, The barrens

of the Oak Ridge area, Tennessee: American Midland 

Naturalist, v. 81, no. 2, p. 315–330.

Duncan, R., Anderson, C., Sellers, H., and Robbins, E., 2008, 

The effect of fire reintroduction on endemic and rare plants 

of a southeastern glade ecosystem: Restoration Ecology, 

v. 16, no. 1, p. 39–49.

Encyclopedia of Life, 2016, Encyclopedia of Life, accessed 

February 18, 2016, at http://www.eol.org. 

Erickson, R., Brenner, L., and Wraight, J., 1942, Dolomitic 

glades of east-central Missouri: Annals of the Missouri 

Botanical Garden, v. 29, no. 2, p. 89–101.

Fenneman, N., 1938, Physiography of the Eastern 

United States: New York, McGraw-Hill.

Fenneman, N., and Johnson, D., 1946, Physical divisions of 

the United States: U.S. Geological Survey map prepared 

in cooperation with the Physiographic Commission, 

scale 1:7,000,000. 

Foti, T., and Glenn, S., 1991, The Ouachita Mountain landscape at the time of settlement, in Restoration of old growth 

forests in the interior highlands of Arkansas and Oklahoma: 

Proceedings of the 1990 Conference: Morrilton, Ark., 

Winrock International Institute for Agricultural Development, p. 49–66.

Guyette, R., Dey, D., and Stambaugh, M., 2003, Fire and 

human history of a barren-forest mosaic in southern 

Indiana: American Midland Naturalist, v. 149, p. 21–34.

Guyette, R., Stambaugh, M., Dey, D., and Muzika, R., 2012, 

Predicting fire frequency with chemistry and climate: 

Ecosystems, v. 15, no. 2, p. 322–335.

Hall, M., 1955, Comparison of Juniper populations on an 

Ozark glade and old fields: Annals of the Missouri Botanical Garden, v. 42, no. 2, p. 171–194.

Hardin, K., Baskett, T., and Evans, K., 1982, Habitat of 

Bachman’s Sparrows breeding on Missouri glades: The 

Wilson Bulletin, v. 94, no. 2, p. 208–212.

Harper, R., 1920, The limestone prairies of Wilcox County, 

Alabama: Ecology, v. 1, no. 3, p. 198–203.

Heikens, A., 1999, Savanna, barrens, and glade communities 

of the Ozark Plateaus Province, in Anderson, R., Fralish, J.,

and Baskin, J., eds., Savannas, barrens, and rock outcrop 

plant communities of North America: New York, Cambridge 

University Press, p. 220–230.

Heikens, A., and Robertson, P., 1994, Barrens of the 

Midwest—A review of the literature: Castanea, v. 59, 

no. 3, p. 184–194.

Heikens, A., and Robertson, P., 1995, Classification of 

barrens and other natural xeric forest openings in southern 

Illinois: Bulletin of the Torrey Botanical Club, v. 122, no. 3, 

p. 203–214.

Hill, J., 2007, The grasshoppers (Orthoptera: Acrididae) 

of the Bibb County Glades, Alabama: Marginalia Insecta, 

v. 2, no. 1, p. 1–3.

Homoya, M., 1994, Indiana barrens—Classification and 

description: Castanea, v. 59, no. 3, p. 204–213.

Integrated Taxonomic Information System, 2015, The 

Integrated Taxonomic Information System, accessed 

October 5, 2015, at http://www.itis.gov.

Jenkins, S., Guyette, R., and Rebertus, A., 1997, Vegetationsite relationships and fire history of a savanna-glade-woodland mosaic in the Ozarks: Proceedings of the 11th Central 

Hardwood Forest Conference.

Jenkins, S., and Jenkins, M., 2006, Effects of prescribed fire 

on the vegetation of a savanna-glade complex in northern 

Arkansas: Southeastern Naturalist, v. 5, no. 1, p. 113–126.

Kimmel, V., and Probasco, G., 1980, Change in woody cover 

on limestone glades between 1938 and 1975: Transactions 

of the Missouri Academy of Science, v. 14, p. 69–74.

Kral, R., and Moffett, J., 2009, Xyris spathifolia (Xyridaceae), 

a new xyrid from the Ketona dolomite-limestone glades 

of Alabama: Journal of the Botanical Research Institute of 

Texas, v. 3, no. 2, p. 469–478.
52 Insular Ecosystems of the Southeastern United States

Kucera, C., and Martin, S., 1957, Vegetation and soil relationships in the glade region of the southwestern Missouri 

Ozarks: Ecology, v. 38, no. 2, p. 285–291.

Laughlin, D., 2004, Woody plant invasion and the importance of anthropogenic disturbance within xeric limestone 

prairies: Journal of the Pennsylvania Academy of Science, 

v. 78, no. 1, p. 12–28.

Laughlin, D., and Uhl, C., 2003, The xeric limestone prairies 

of Pennsylvania: Castanea, v. 68, no. 4, p. 300–316.

Lawless, P., Baskin, J., and Baskin, C., 2004, The floristic 

ecology of xeric limestone prairies in Kentucky, and a 

comparison to limestone cedar glades and deep-soil barrens: 

Sida, v. 21, no. 70, p. 1055–1079.

Lawless, P., Baskin, J., and Baskin, C., 2006a, Xeric limestone 

prairies of eastern United States: review and synthesis: The 

Botanical Review, v. 72, no. 3, p. 235–272.

Lawless, P., Baskin, J., and Baskin, C., 2006b, Scaledependent classification of xeric limestone prairies—

Annual or perennial grasslands?: Annals of the Missouri 

Botanical Garden, v. 93, no. 3, p. 455–464.

Lowell, K., and Astroth, J., 1989, Vegetative succession 

and controlled fire in a glades ecosystem: a geographical 

information system approach: International Journal of 

Geographical Information Systems, v. 3, no. 1, p. 37–41.

Ludwig, J., 1999, The flora of dolomite and limestone barrens 

in southwestern Virginia: Castanea, v. 64, no. 3, p. 209–230.

Mann, L., King, A., Dale, V., Hargrove, W., WashingtonAllen, R., Pounds, L., and Ashwood, T., 1999, The role 

of soil classification in geographic information system 

modeling of habitat pattern—Threatened calcareous 

ecosystems: Ecosystems, v. 2, p. 524–538.

Martinez, E., and Brown, R., 2007, Argyriini (Lepidoptera: 

Crambidae) of Mississippi and Alabama with a rediscription 

of Argyria rufisignella (Zeller): Journal of the Lepidopterists’ Society, v. 61, no. 2, p. 78–83.

Meyer, R., 2006, Aimophila aestivalis: Fire Effects Information System (Online), U.S. Department of Agriculture, 

Forest Service, Rocky Mountain Research Station, 

Fire Sciences Laboratory, accessed August 27, 2015, at 

http://www.fs.fed.us/database/feis/animals/bird/aiae/all.html. 

Mills, J., 2008, Fifty years of change in Wisconsin cedar 

glades: The American Midland Naturalist, v. 159, no. 1, 

p. 214.

Molano-Flores, B., and Bell, T., 2012, Projected population 

dynamics for a federally endangered plant under different 

climate change emission scenarios: Biological Conservation, v. 145, no. 1, p. 130–138.

Morris, L., Walck, J., and Hidayati, S., 2002, Reproduction 

of the invasive Ligustrum sinense and native Forestiera 

ligustrina (Oleaceae)—Implications for the invasion and 

persistence of a nonnative shrub: International Journal of 

Plant Sciences, v. 163, no. 6, p. 1001–1010.

NatureServe, 2015, NatureServe Explorer—An online 

encyclopedia of life [web application], ver. 7.1, accessed 

October 7, 2015, at http://www.natureserve.org/explorer.

Nelson, P., Fitzgerald, J., Larson, K., McCoy, R., Schotz, A., 

Taft, J., Witsell, T., and Yahn, B., 2013, Central hardwoods 

joint venture glade conservation assessment for the Interior 

Highlands and Interior Low Plateaus of the central hardwoods region: Central Hardwoods Joint Venture.

Noss, R., 2013, Forgotten grasslands of the south: Washington, 

D.C., Island Press.

Quarterman, E., 1950, Ecology of cedar glades. I. Distribution 

of glade flora in Tennessee: Bulletin of the Torrey Botanical 

Club, v. 77, no. 1, p. 1–9.

Quarterman, E., 1989, Structure and dynamics of the limestone cedar glade communities in Tennessee: Journal of the 

Tennessee Academy of Science, v. 64, no. 3, p. 155–158.

Quarterman, E., Burbanck, M., and Shure, D., 1993, Rock 

outcrop communities—Limestone, sandstone, and granite, 

in Martin, W., Boyce, S., and Echternacht, A., eds., 

Biodiversity of the southeastern United States—Upland 

terrestrial communities: New York, John Wiley and Sons, 

Inc., p. 35–86.

Rhoades, C., Miller, S., and Shea, M., 2004, Soil properties 

and soil nitrogen dynamics of prairie-like forest openings 

and surrounding forests in Kentucky’s Knobs Region: 

The American Midland Naturalist, v. 152, no. 1, p. 1–11.

Rhoades, C., Miller, S., and Skinner, D., 2005, Forest 

vegetation and soil patterns across glade-forest ecotones 

in the Knobs region of northeastern Kentucky, USA: The 

American Midland Naturalist, v. 154, no. 970, p. 1–10.

Sauer, C., 1950, Grassland climax, fire, and man: Journal of 

Range Management, v. 3, no. 1, p. 16–21.

Skinner, B., Probasco, G., and Samson, F., 1983, Environmental requirements of three threatened plants on limestone 

glades in southern Missouri: Biological Conservation, v. 25, 

p. 63–73.
Chapter D 53

Sutter, R., Govus, T., Smyth, R., Nordman, C., Pyne, M.,

 and Hogan, T., 2011, Monitoring change in a central 

U.S. calcareous glade—Resampling transects established

in 1993: Natural Areas Journal, v. 31, no. 2, p. 163–172.

Taylor, K., and Estes, D., 2012, The floristic and community

ecology of seasonally wet limestone glade seeps of 

Tennessee and Kentucky: Journal of the Botanical Research 

Institute of Texas, v. 6, no. 2, p. 711–724.

Thompson, J., Walck, J., and Hidayati, S., 2006, Microhabitat 

requirements of the federally endangered Dalea foliosa, 

with recommendations on establishment of new populations: Castanea, v. 71, no. 2, p. 94–104.

Thornberry-Ehrlich, T., 2012, Stones River National Battlefield—Geological resources inventory report: Fort Collins, 

Colo., Natural Resource Report NPS/NRSS/GRD/NRR.

Trammell, T., Rhoades, C., and Bukaveckas, P., 2004, Effects 

of prescribed fire on nutrient pools and losses from glades 

occurring within oak-hickory forests of Central Kentucky: 

Restoration Ecology, v. 12, no. 4, p. 597–604.

Van Zandt, P., Collins, E., Losos, J., and Chase, J., 2005, 

Implications of food web interactions for restoration of 

Missouri Ozark glade habitats: Restoration Ecology, v. 13, 

no. 2, p. 312–317.

Ware, S., 2002, Rock outcrop plant communities (glades) 

in the Ozarks—A synthesis: The Southwestern Naturalist, 

v. 47, no. 4, p. 585–597.

Ware, S., 2010, Croton monanthogynus and Crotonopsis 

elliptica (Euphorbiaceae) in Ozark rock outcrop communities—Abundance, soil depth, and substrate tolerance: 

Northeastern Naturalist, v. 17, no. 4, p. 659–666.

Webb, D., DeSelm, H., and Dennis, W., 1997, Studies of 

prairie barrens of northwestern Alabama: Castanea, v. 62, 

no. 3, p. 173–184.

Young, C., Morrison, L., and Haack, J., 2009, Habitat 

relationships and management implications for Lesquerella 

filiformis Rollins (Missouri bladderpod) on a xeric limestone prairie: The Journal of the Torrey Botanical Society, 

v. 136, no. 2, p. 233–241.

Chapter E. Mid-Appalachian Shale Barrens

Introduction

Shale barrens of the mid-Appalachian region (fig. E–1) 

are forest openings with scattered shrubs and patches of 

herbaceous vegetation interspersed with bedrock outcrops 

supporting bryophyte and lichen communities (Norris and 

Sullivan, 2002). Stunted trees, including Quercus, Pinus, 

and Carya species and Juniperus virginiana, are sparsely 

distributed, resulting in shale barrens typically having a 

moderately open canopy with mostly woody vegetation 

and a relatively sparse herbaceous understory (Clarkson, 

1966; Braunschweig and others, 1999). These shale barrens 

support a number of taxa of conservation concern, including 

several endemic plant taxa that display specialized adaptations to tolerate the abiotic stress regime of shale barren 

environments. 

Geographic Range

Mid-Appalachian shale barrens are present within 

the Valley and Ridge Physiographic Province of the 

Appalachian Highlands (Fenneman and Johnson, 1946) 

in south-central Pennsylvania, western Maryland, eastern 

West Virginia, and southwestern Virginia (Keener, 1983; 

Morse, 1983; Braunschweig and others, 1999). Norris and 

Sullivan (2002) estimated the number of shale barrens to be 

25 in Pennsylvania, 35 in Maryland, 800 in Virginia, and 

100 in West Virginia, although they acknowledged that the 

accuracy of these counts might be affected by ambiguities 

over the definition of shale barrens as an ecosystem type. 

The greatest concentration of shale barrens is within a zone 

lying along the southern part of the border between Virginia 

and West Virginia (Norris and Sullivan, 2002). Figure E–2 

shows a map of the general geographic distribution of 

mid-Appalachian shale barrens.

Within this region of the Appalachian Highlands, shale 

barrens are primarily associated with Upper Devonian strata 

of shale including the Brallier Shale, the Romney Shale, and 

the Trimmers Rock Formation (Artz, 1948; Platt, 1951; Core, 

1952; Keener, 1983; Morse, 1983), although they have also 

been described on the Martinsburg Formation of Ordovician 

strata in the Massanutten Mountains of Virginia (Artz, 1937) 

and on the Rose Hill Formation of Silurian strata in Maryland (Morse, 1983). 

Rawinski and others (1996) described shale barrens with 

a similar suite of endemics in the Blue Ridge Physiographic 

Province and argued that these constitute a subtype known 

as Blue Ridge shale barrens. Although mid-Appalachian 

shale barrens share some physical and physiognomic features 

(such as shallow soil, open canopy, and prevalence of 

herbaceous vegetation) with shale barrens in other parts of 

the United States such as in Illinois, they are clearly distinct 

from Midwest shale barrens at a floristic level, as evidenced 

by a suite of species endemic only to the mid-Appalachian 

barrens (Heikens and others, 1994; McCall and Gibson, 1999; 

West and others, 2009).

Figure E–1. Shale barren above a cowpasture, Bath County, 

Virginia. Photograph by Gary Fleming, Virginia Department of 

Conservation and Recreation.
56 Insular Ecosystems of the Southeastern United States

Figure E–2. General geographic distribution of mid-Appalachian shale barrens. 

85° 80° 75° 70°

40°

35°

30°

25°

EXPLANATION

Appalachian Shale Barrens 

 ecological system—Modified from

 U.S. Geological Survey (2011)

Mid-Appalachian Shale Barrens—

 Modified from Keener (1983)

Appalachian Plateaus

Blue Ridge

Piedmont

Valley and Ridge

Figure E-2

Physiographic province—From 

 Fenneman and Johnson (1946)

0 100 200 300 400 KILOMETERS

0 100 200 300 400 MILES

Base modifed from U.S. Department of Commerce, Census Bureau,

U.S. States, 2015, Mercator Auxiliary Sphere Projection

Horizontal datum is World Geodedic System of 1984

TENNESSEE

ALABAMA GEORGIA

SOUTH

CAROLINA

NORTH

CAROLINA

WEST

VIRGINIA

OHIO

PENNSYLVANIA

NEW YORK

VERMONT

NEW

HAMPSHIRE

INDIANA

VIRGINIA

FLORIDA

KENTUCKY

MISSISSIPPI

ILLINOIS

MICHIGAN

NEW JERSEY

DELAWARE

MARYLAND

MASSACHUSETTS

MAINE

CONNECTICUT

RHODE ISLAND
Chapter E 57

Physical Geography

Mid-Appalachian shale barrens typically are present at 

elevations ranging from approximately 1,000 to 1,900 feet 

(about 300 to 575 meters) on steep (20 – 40 percent) slopes 

(Platt, 1951; Keener, 1983; Braunschweig and others, 1999). 

Figure E–3 shows a shale barren with characteristically steep 

slope. Widespread agreement exists that the spatial distribution of mid-Appalachian shale barrens is highly influenced by 

topographic position on the landscape (especially slope and 

aspect), which is in turn determined largely by geophysical 

and geographic factors (Allard and Leonard, 1946; Platt, 

1951; Keener, 1983; Braunschweig and others, 1999). Within 

areas of varying topography, shale barrens are present more 

frequently on east-, south-, and west-facing slopes, while 

adjacent areas of level terrain or north-facing slopes are more 

likely to be forested, implying that slope and aspect are important—though probably not exclusive—determinants of shale 

barren formation (Braunschweig and others, 1999). Allard 

and Leonard (1946) argued that the steep slopes upon which 

shale barrens typically develop are produced by a combination 

of erosion-resistant sandstone ridge-caps upslope and stream 

action undercutting the hillside downslope, indicating that—

in addition to bedrock geology—proximity to streams and the 

hydrologic and geomorphic characteristics of those streams 

are additional factors helping to create the topographic 

contexts in which shale barrens may develop. An example of a 

shale barren undercut by a stream is shown in figure E–4. 

Platt (1951) questioned this conceptualization of shale 

barren development, noting that (1) softer shales in the 

region with similar geographic and geologic conditions do 

not produce barrens, (2) more barrens develop away from 

streams than near them, for example on the flanks of ridges, 

and (3) the perception that undercutting stream action is 

necessary for shale barren formation might be attributable 

to observational bias because barrens are often viewed from 

roads, which commonly are parallel to streams. Despite these 

assertions, recent publications—such as Braunschweig and 

others (1999)—have continued to maintain that undercutting 

stream action is an important factor in creating and maintaining shale barrens.

Shale barren soils are typically overlain by a surface layer 

of weather-resistant shale fragments and typically have little 

organic litter at the soil surface (Keener, 1983; Braunschweig 

and others, 1999). Platt (1951) suggested that this surface 

Figure E–3. Shale barren, Edinburg Gap, George Washington and Jefferson National Forests, Shenandoah County, 

Virginia. Photograph by Gary Fleming, Virginia Department of Conservation and Recreation.
58 Insular Ecosystems of the Southeastern United States

layer of shale fragments provided insulation to the soil below, 

partially shielding it from the drying effects of wind and direct 

sunlight. Below this surface layer, he also noted poor differentiation of the horizons within the soil profile of shale barrens 

on the Brallier Shale and the general lack of a B horizon. 

Although these characteristics of shale barren soils make them 

atypical for the region, their moderately acidic pH and the 

availability of certain inorganic nutrients is comparable to the 

soils of nearby forested sites (Braunschweig and others, 1999).

Stress and Disturbance Regimes

The exact nature of the stress regime characteristics of 

mid-Appalachian shale barrens has been the subject of some 

debate. Several generations of researchers have attributed 

the existence of shale barren communities to a lack of soil 

nutrients resulting from an absence of true soil due to erosion 

(Allard and Leonard, 1946; Keener, 1983). Platt (1951), 

however, concluded that shale barrens “are not the subject of 

extreme erosion” and discounted erosion as a major factor 

in forming or maintaining shale barren ecosystems, citing 

evidence such as the lack of erosional gullies and observations 

from dye experiments of sheet flow moving primarily over 

resistant rock fragments rather than over soil. Relatively recent 

publications, including Braunschweig and others (1999) and 

the U.S. Forest Service (2011), have asserted that surficial 

shale fragments continually migrate downslope such that the 

ground surface of shale barrens is “constantly shifting” (Nott, 

2006), but they did not present empirical evidence to support 

these claims. Platt (1951), however, did empirically investigate 

this possibility. Through a simple experiment involving stripes 

spray-painted along contour lines across shale barrens, he 

found the stripes to be almost wholly intact after 10 months, 

with the exception of a few small washes, indicating minimal 

downslope migration of shale fragments over the course of 

his study. He also pointed to the lack of accumulated rock 

fragments at the downslope bases of shale barrens and to the 

presence of lichen communities atop shale fragment surfaces 

as further evidence in contradiction to the idea of a highly 

unstable ground surface. It appears that although Platt (1951) 

presented empirical evidence against the premise of erosion 

(either of soil or of surficial rock fragments) as a primary 

component of the shale barren stress regime, this concept 

has nonetheless been perpetuated by subsequent generations 

of investigators.

Similarly, the degree and type of water stress experienced by shale barren vegetation have been debated. It has 

been noted that the geographic area in which shale barrens 

are clustered experiences relatively less rainfall than the 

surrounding region because it lies in the rain shadow of the 

Allegheny Front, which extends from central Pennsylvania to 

eastern West Virginia (Clarkson, 1966; Norris and Sullivan, 

2002). Platt (1951), however, observed that hillsides adjacent 

Figure E–4. Shale barren undercut by a stream, Bath County, Virginia. Photograph by Irvine Wilson, Virginia 

Department of Conservation and Recreation.
Chapter E 59

to shale barrens receiving comparable precipitation are 

forested, indicating that lack of precipitation alone cannot 

account for barren formation and maintenance. Platt (1951) 

also empirically discredited earlier notions that low levels 

of soil nutrients or soil moisture throughout the soil profile 

are primary determinants of the structure of shale barren 

communities. Platt (1951) demonstrated that for shale barren 

soils, “the depth of the C horizon, as well as its moisture and 

nutrient content, is essentially similar to that of the vegetated 

slopes,” and asserted that the C horizon maintains sufficient 

moisture (largely due to the insulating effects of the shale 

fragments at the soil surface) to sustain species adapted to 

exploiting it. However, he also demonstrated that the evaporative power of air over shale barrens was more than twice that 

for adjacent north-facing slopes, with temperature having a 

disproportionately high effect on evaporation for shale barren 

sites. Thus, Platt (1951) concluded that surficial hydrologic 

conditions were indeed xeric in shale barrens for at least part 

of the growing season.

Generally, more recent publications (for example, 

Keener, 1983; Braunschweig and others, 1999; Norris and 

Sullivan, 2002) have rejected the earlier assumption that 

below-ground resources—subsurface moisture, geochemistry, 

and nutrients—are the primary limiting factors in midAppalachian shale barren habitats (Allard and Leonard, 1946), 

and have instead focused on the microenvironment at the 

soil surface. Braunschweig and others (1999) suggested that 

the nutrient limitation that might exist could take the form of 

reduced organic matter or nitrate availability in the A horizon. 

In particular, high insolation, high temperature (up to 63 °C), 

and xeric conditions at the soil surface—resulting from 

topography (slope and aspect) and lack of canopy cover—are 

the primary components of the stress regime experienced by 

shale barren species, with this stress most acutely affecting 

germinating seeds and seedlings (Platt, 1951; Keener, 1983; 

Braunschweig and others, 1999; Norris and Sullivan, 2002).

Certain forms of disturbance may contribute to shale 

barren maintenance. Anthropogenic disturbance, such as from 

logging or grazing, is thought to be of minimal importance to 

the maintenance of ecosystem integrity overall, though some 

shale barrens may show evidence of previous grazing (Braunschweig and others, 1999). Fire is not believed to be a major 

factor in creating or maintaining shale barrens, largely because 

little fuel exists to sustain large fires (Allard and Leonard, 

1946; Platt, 1951; Keener, 1983; Braunschweig and others, 

1999). However, fire regimes in nearby pine-oak forests 

appear, based on dendrochronology, to have been historically 

important in suppressing later successional species such as 

Acer saccharum, Fagus grandifolia, and Tsuga canadensis

(Abrams and others, 1995). Disturbance from fire within 

adjacent pine-oak communities possibly played a role in 

limiting encroachment of woody species onto shale barrens 

(U.S. Forest Service, 2011). The possible role of fire disturbance in maintaining shale barrens deserves further research 

so as to inform appropriate management policies (Norris and 

Sullivan, 2002).

Community Types and Vegetation 

Dynamics

Natural Heritage Programs in Virginia, West Virginia, 

Maryland, and Pennsylvania have all used different sets of 

names for community classification in shale barrens (Norris 

and Sullivan, 2002). Although these differing naming 

conventions were justified in part because of regional variation 

in community occurrence, they also complicated efforts to 

assess shale barren habitat throughout its full range. A national 

classification, developed and standardized by NatureServe 

and the U.S. National Vegetation Classification, identifies one 

prairie community and several woodland community types 

(Fleming and Patterson, 2012), as described next.

Central Appalachian Shale Ridge Prairie 

This community type is ranked G2, with fewer than 

25 documented occurrences. This community has been 

referred to as “shale ridge balds,” based on characteristic topographic position along spur ridge crests. In these locations, 

associations with more stable habitat, higher soil fertility, and 

dominance of graminoids (such as Schizachyrium scoparium) 

are distinguishable from shale woodland community types by 

having a more open physiognomy and “prairie-like” appearance. This community type provides optimal habitat for the 

endemics Arabis serotina, Clematis viticaulis, Eriogonum 

allenii, Oenothera argillicola, and Trifolium virginicum.

Chestnut Oak–Virginia Pine/Ragwort Acidic 

Shale Woodland (Southern Type)

This community type is ranked G3, with the number of 

occurrences estimated at several hundred within the State of 

Virginia alone. This community is located on steep, unstable 

slopes with canopy coverage ranging from almost closed to 

mostly open. Shrubs are sparsely distributed, and herbaceous 

coverage is patchy and typically less than 50 percent. 

Thus, this community type can range in physiognomy from 

scrubby-herbaceous to open stunted woodland. As with other 

community types, outcroppings of shale bedrock are common, 

as are unvegetated areas with exposed shale fragments at the 

soil surface. Compared to other shale woodland community 

types, species richness and the concentration of shale barren 

endemic taxa in this community are both high.

North-Central Appalachian Acidic 

Shale Woodland

This woodland community type is also ranked G3, 

with fewer than 100 occurrences estimated rangewide (in 

Maryland, Pennsylvania, Virginia, and West Virginia). This 

community typically forms on moderate to steep slopes, 
60 Insular Ecosystems of the Southeastern United States

including convex slopes and ridge spurs. Areas of exposed 

bedrock may be present, and soils are typically acidic. The 

canopy ranges from almost closed to mostly open, and is 

often dominated by Pinus virginiana, Juniperus virginiana, 

Quercus rubra, and Quercus prinus. The herbaceous layer 

is typically less than 25 percent and is patchy in distribution 

along with sparse or scattered shrubs. This community type 

has a more northerly distribution than some of the other 

community types and characteristically lacks the species 

richness—as well as several endemic or near-endemic taxa 

such as Arabis serotina and Clematis coactilis—found in more 

southerly community types.

Central Appalachian Virginia Pine/Sparse Herbs 

Shale Woodland

This community type is also ranked G3, with a few 

hundred occurrences in west-central Virginia and eastern 

West Virginia. This community typically occupies divide 

crests and strongly convex slopes. Unlike other shale woodland community types, this type typically lacks Juniperus 

virginiana, is species poor in the herbaceous layer, and has a 

greater dominance of lichens such as Cladina species. Overall 

species richness is low, and the canopy is strongly dominated 

by Pinus virginiana and Quercus prinus. Lack of fire in this 

community type is believed to contribute to canopy dominance 

by P. virginiana as well as to relatively low herbaceous cover, 

with edaphic conditions stunting the growth of hardwood 

successors such as Quercus alba and Q. prinus.

Central Appalachian Chestnut Oak/Mixed Herbs 

Shale Woodland

This G3-ranked woodland community commonly borders 

shale prairies and typically has an open canopy dominated by 

Quercus prinus. There is a patchy shrub layer, and although 

herbaceous coverage is generally sparse, there is a relatively 

high richness of drought-tolerant forbs and graminoids. Soils 

are typically shallow and acidic. Compared to the Central 

Appalachian Virginia Pine/Sparse Herbs Shale Woodland, this 

community type has greater hardwood representation in the 

canopy and greater herbaceous species richness, differences 

that might be attributable to edaphic differences or to greater 

incidence of fire in the past.

Several other related community types also exist within 

the southeastern United States, including two shale woodland 

types that are distributed within the Blue Ridge in the 

Southern Appalachians and a Central Appalachian Circumneutral Barrens community type that may form on calcareous 

shale or on other substrates such as sandstone. 

Contributions to Regional Biodiversity

Mid-Appalachian shale barrens support a number of plant 

taxa of conservation concern, including several strict endemics 

as well as near-endemics that also are present in related 

contexts, such as outcrops or barrens, on other substrates, or 

in other regions (table 3–4 in appendix 3). Figure E–5 shows 

several rare taxa supported by shale barren habitat. 

Endemic and Biogeographically Disjunct Taxa

Braunschweig and others (1999) observed that, although 

at least seven lists have been published of species endemic to 

shale barrens, no two lists are in complete agreement and they 

vary in the number of species included. Several taxa have been 

renamed, or have undergone taxonomic clarification or recategorization, making the comparison of lists across publications 

difficult. For example, Frye (1942) mentioned Celtis pumila as 

a shale barren near-endemic; however, this species is no longer 

recognized, and its supposed occurrences are now considered 

to be simply dwarf individuals of Celtis occidentalis (Thomas 

Wieboldt, Virginia Polytechnic Institute, written commun., 

2014). Similarly, Allard (1957) asserted that a shale barren 

lichen named Cetraria islandica was disjunct from boreal and 

arctic populations; however, this lichen has been redetermined 

as Cetraria arenaria, which has a nondisjunct distribution 

including the Appalachian occurrences (Thomas Wieboldt, 

Virginia Polytechnic Institute, written commun., 2014).

Some of the ambiguity in itemizing the shale barren 

endemics is also likely due to the presence of several taxa 

that have at times been referred to as “strict endemics” but 

have also been documented on nonshale substrates and in 

other ecosystems. These include Solidago arguta var. harrisii

(found on limestone cliffs and barrens), Trifolium virginicum

(on diabase and ultramafic substrates), and Clematis coactilis

(on limestone and dolomite) (Braunschweig and others, 

1999). Wherry (1953) also described occurrences of some 

near-endemics on other bedrock types, including sandstone 

and limestone. Bartgis (1993) noted several shale barren 

“endemics” in limestone glades in West Virginia. These 

findings indicate that these taxa might appropriately be labeled 

near-endemics, and also that they are not highly specialized in 

terms of their adaption to shale substrate. Indeed, a number of 

investigators have concluded that many shale barren endemics 

are geographically restricted less by geologic substrate than 

by their inability to compete in more shaded, mesic habitats 

because they are obligate heliophytes (Wherry, 1953; Baskin 

and Baskin, 1988; Braunschweig and others, 1999; Byers and 

Norris, 2011).

Of the shale barren endemics, only Arabis serotina is 

a federally listed endangered species, with possibly fewer 
Chapter E 61

Figure E–5. Characteristic plant taxa 

of mid-Appalachian shale barrens: 

A, Clematis viticaulis, Bath County, Virginia; 

B, Eriogonum allenii, Woodstock Gap, 

George Washington and Jefferson National 

Forests, Shenandoah County, Virginia; 

C, Oenothera argillicola, Alleghany County, 

Virginia; D, Trifolium virginicum, Bath 

County, Virginia. Photographs A, C, and 

D by Irvine Wilson, Virginia Department of 

Conservation and Recreation; Photograph B

by Gary Fleming, Virginia Department of 

Conservation and Recreation.

A

B

C D
62 Insular Ecosystems of the Southeastern United States

than 4,000 individuals remaining (U.S. Forest Service, 2011). 

However, several other endemic and near-endemic plant taxa 

are globally rare, including Clematis coactilis, Oenothera 

argillicola, Taenidia montana, and Trifolium virginicum (all 

ranked G3); Allium oxyphilum and Phlox buckleyi (G2); and 

Clematis viticaulis (G1).

Many shale barren endemics and near-endemics share 

common characteristics believed to represent adaptations to 

the stress regime of their habitat. For example, adaptations to 

drought stress may provide a competitive advantage to some 

shale barren plants in an environment where surface soil moisture is limited but water is available lower in the soil profile. 

Certain species, particularly geophytes and hemicryptophytes, 

are dependent primarily on the C horizon for moisture and 

nutrients and can thus better tolerate low water availability at 

upper soil horizons (Platt, 1951; Braun, 1955). Not surprisingly, 

many of these species possess unusually high root-to-shoot 

ratios (Platt, 1951; Keener, 1983; Norris and Sullivan, 2002). 

In addition, some adaptations appear to aid in germination and recruitment under conditions of harsh abiotic 

stress. Some shale barren species time their life cycles to the 

seasonal availability of soil moisture, such that plants are 

dormant or engaged in seed maturation during the driest times 

of the year (Platt, 1951). Also, experiments have shown the 

seeds, seedlings, and leaves of some endemics to have relatively 

high tolerance for heat. Platt (1951) found bud mortality for 

several shale barren plants to occur between 42 and 45 °C. 

Braunschweig (1993) found membrane thermotolerance for 

the mature leaves of the shale barren endemic Eriogonum 

allenii to be up to 52 °C. Tolerance for high-heat and highlight conditions is facilitated in some cases by morphological 

adaptations such as waxy leaf coatings and steep leaf angles 

to minimize light interception (Nilsen and Orcutt, 1996). 

Additionally, some shale barren taxa have been shown to have 

considerable intraspecific variation and genetic diversity (Platt, 

1951; Baskin and Baskin, 1988). Bayer and Crawford (1986), 

for example, reported genetic diversity for the shale barren 

endemic Antennaria virginica that was comparable to that of 

its more geographically widespread congeners. 

Biogeographic disjunction has also been noted for some 

plant taxa found in shale barrens. Several generations of shale 

barren explorers have noted an affinity between certain shale 

barren plants and western taxa, highlighted by disjunct populations of western species growing on shale barrens (Frye, 

1942; Braun, 1955; Keener, 1983; Braunschweig and others, 

1999). E. allenii and Packera antennariifolia are members 

of widespread genera, with closest relatives that are typically 

associated with the Rocky Mountains (Core, 1932; Artz, 1948; 

Braun, 1955; Noss, 2013). Core (1932, 1952) described the 

Appalachian presence of Astragalus distortus as being disjunct 

from its primary range in the Midwest prairies. Knobloch 

and Lellinger (1969) described occurrences of Cheilanthes 

eatonii on Virginia and West Virginia shale barrens, whereas 

the central distribution of this species is in the southwestern 

United States and Mexico. Some disjunctions occur with large 

distances of complete separation: approximately 1,200 miles 

(about 2,000 kilometers) in the case of Asplenium septentrionale, which ranges from Oregon to Texas with the exception 

of disjunct occurrences in West Virginia on shale barrens. A 

few other disjunct taxa have populations geographically intermediate between their western ranges and the Appalachian 

shale barren populations (Braunschweig and others, 1999). 

Animal Taxa of Conservation Concern

Relatively little research has been done on the fauna of 

shale barrens, including the interactions between shale barren 

plants and insects, such as herbivory or pollination (Wheeler, 

1995). Some evidence does suggest that shale barrens offer 

habitat to rare arthropod taxa. Two rare butterfly species 

are associated with shale barren habitat: Pyrgus wyandot

(Appalachian grizzled skipper) and Euchloe olympia (Olympia 

marble) (Braunschweig and others, 1999). P. wyandot is 

ranked G1 and has been extirpated across much of its original 

range in the United States, whereas E. olympia has secure 

populations throughout the Midwest, Rocky Mountains, and 

in Canada but is State-rare in the Appalachian region. Several 

rare spider species, including one previously undescribed 

species (Hahniidae), another possible new species (Araniidae), 

and four species previously unrecorded in West Virginia, were 

collected by the West Virginia Natural Heritage Program in a 

pit trap survey at one shale barren site (Norris and Sullivan, 

2002). Chinaola quercicola, a microphysid insect with 

specialized habitat on granite outcrops and pitch pine-scrub 

oak barrens, has also been found in a shale barren (Wheeler, 

1992). Some evidence suggests that shale barrens may be 

ecologically important to certain insects by providing specialized habitat or access to resources. For example, Kalhorn 

and others (2003) reported greater activity of bees on shale 

barren interiors than in the surrounding forest or ecotone. In a 

survey of plant bugs (Heteroptera: Miridae), Wheeler (1995) 

noted that certain species showed a distinct affinity to midAppalachian shale barrens.

Conservation Considerations

Mid-Appalachian shale barrens have been recognized as 

important conservation priorities, primarily because of their 

contributions to regional plant biodiversity (Braunschweig 

and others, 1999; Norris and Sullivan, 2002). A number of 

threats and conservation options have been identified. Gaps in 

scientific knowledge have been highlighted, indicating needs 

for additional research in this ecosystem.

Threats to Ecosystem Integrity

Mid-Appalachian shale barrens may be more protected 

from anthropogenic destruction than some other types of 
Chapter E 63

natural forest openings because of their lack of commercial 

value and because their steep slopes and scree make them 

treacherous to hikers and unattractive sites for development 

(Braunschweig and others, 1999; Norris and Sullivan, 2002). 

Trampling from recreational use may still be a problem in 

some cases, however, and some shale barrens have been 

damaged by the construction of hiking trails (U.S. Forest 

Service, 2011). A few shale barrens have been degraded or 

destroyed by road construction (Braunschweig and others, 

1999; Norris and Sullivan, 2002), and shale barrens near 

roads have also been damaged by quarrying for road material 

(Platt, 1951; Norris and Sullivan, 2002). Shale barren sites 

located downslope from developed areas have also been 

subjected to trash dumping (Norris and Sullivan, 2002; 

U.S. Forest Service, 2011). Although shale barrens typically 

do not produce adequate timber to attract large-scale logging 

operations, certain sites have been affected by the cutting of 

trees for fenceposts (Braunschweig and others, 1999; Norris 

and Sullivan, 2002). A few shale barrens with low slope 

near pastures have been grazed by livestock (Platt, 1951). 

At least one shale barren has been destroyed by flooding 

from an upstream impoundment (Dix, 1990; U.S. Forest 

Service, 2011).

The application of pesticides may be a threat to the rare 

flora and fauna of some shale barrens, which are maximally 

exposed to aerial spraying due to their relatively open canopies. 

The regional use of insecticides to control Lymantria dispar

(gypsy moth) has inadvertently affected rare lepidoptera such 

as the Appalachian grizzled skipper in shale barrens (Norris 

and Sullivan, 2002). Dimilin is a broad-spectrum insecticide 

that can persist for up to several years in leaf litter and may 

pose a lasting threat in shale barrens. Bacillus thuringiensis

(Bt), a biological pesticide, persists for only a week or so 

but can be a threat to shale barren lepidoptera if applied 

during larval development (Dix, 1990). These pesticides may 

also threaten the globally rare shale barren endemic Arabis 

serotina by harming its insect pollinators, which include the 

Appalachian grizzled skipper (Dix, 1990; U.S. Forest Service, 

2011). Lack of cross-pollination in A. serotina is likely to 

result in inbreeding depression and a corresponding reduction in germination; thus, attempts to conserve A. serotina

populations must necessarily incorporate steps to conserve 

the Appalachian grizzled skipper and other rare pollinators in 

shale barrens (Nott, 2006). 

Shale barren endemics such as A. serotina may also be 

threatened by browsing from native herbivores, primarily deer 

(Nott, 2006), because several A. serotina populations show 

evidence of moderate to severe damage from deer browsing 

(U.S. Forest Service, 2011). Although the stress regime of 

shale barrens may offer some protection against invasion by 

exotic species, grazing by sheep and goats and recreational 

use can introduce exotic weedy plants that may pose a threat 

to certain endemics (Braunschweig and others, 1999; Keech, 

1999; Nott, 2006). Some invasive species can be allelopathic 

to native taxa and distasteful to native herbivores (Nott, 2006). 

The presence of exotics, including Centaurea biebersteinii, 

Bromus sterilis, and B. japonicus, has been noted and 

mentioned as a possible threat to A. serotina, which is a poor 

competitor (Dix, 1990; Norris and Sullivan, 2002; U.S. Forest 

Service, 2011). 

Conservation Strategies

Shale barrens are present on protected public land in 

three national forests: George Washington and Jefferson 

National Forests in Virginia and Monongahela National Forest 

in West Virginia (DeSelm and Murdock, 1993; Norris and 

Sullivan, 2002). A few other shale barren sites are present 

on land owned by the U.S. Army Corps of Engineers and 

The Nature Conservancy (Norris and Sullivan, 2002). The 

great majority of mid-Appalachian shale barrens (possibly 

up to 80 percent) are believed to be on privately owned, 

nonprotected land (Norris and Sullivan, 2002), however, no 

comprehensive map of the location of shale barrens exists. 

Even on protected land, active management of shale barrens is 

sporadic and is generally limited to the removal or attempted 

control of exotic species (Norris and Sullivan, 2002). Some 

past and current monitoring efforts have focused on rare 

plants (especially Arabis serotina and Eriogonum allenii) and 

lepidoptera; however, the methods and intensity of monitoring 

activities vary between managing agencies.

Knowledge Gaps

Norris and Sullivan (2002) stated that “despite a plethora 

of shale barren literature, there has been far more speculation 

than substantial ecological research and observation.” Very 

little is known about pollination or dispersal strategies of 

shale barren endemics (Braunschweig and others, 1999), and 

as a result, virtually nothing is known about how endemics 

disperse between shale barrens (Norris and Sullivan, 2002). 

Also, few publications have described shale barren species 

other than vascular plants. Ammons (1943) presented a list 

of 21 mosses and 2 liverworts observed on a single trip to 

a single shale barren; however, he noted that many of these 

bryophyte species were not necessarily confined to shale 

barren habitat. The ecology, physiology, and biogeography 

of lichens and bryophytes in mid-Appalachian shale 

barrens deserve further study because they appear to be 

important constituents of several shale barren communities. 

Shale barrens have also received limited attention from 

entomologists, despite the discovery of rare and disjunct 

lepidoptera that use shale barren endemic plants as 

larval hosts (Norris and Sullivan, 2002). Although some 

information is available on the interspecific interactions of 

the more well-studied shale barren plants such as Arabis 

serotina (Nott, 2006), many of the trophic interactions 

and food-web dynamics in shale barrens have yet to be 

examined empirically.
64 Insular Ecosystems of the Southeastern United States

Possible Ecological Effects from Climate Change

Although harsh microclimatic conditions are frequently 

described in the literature concerning mid-Appalachian shale 

barrens, little research has been done on their vulnerability 

to climate change. Recent reviewers of this ecosystem, such 

as Braunschweig and others (1999) and Norris and Sullivan 

(2002), discussed a number of anthropogenic threats to shale 

barrens but did not explicitly discuss climate change. Byers 

and Norris (2011) assessed the climate-change vulnerability 

of several West Virginia species and determined that Arabis 

serotina and Trifolium virginicum would be highly vulnerable, 

Phlox buckleyi would be moderately vulnerable, and Pyrgus 

wyandot was presumed stable; however, the confidence 

associated with these predictions was categorized as “low” for 

A. serotina and T. virginicum. Furthermore, these assessments 

were for individual species, and Byers and Norris (2011) did 

not include an ecosystem-level vulnerability assessment nor 

a detailed explanation of the mechanisms by which climate 

change might threaten each species. For example, increased 

vulnerability for A. serotina was attributed to factors named 

“dispersal/movement,” “historical hydrological niche,” 

“disturbance,” “physical habitat,” and “genetic bottlenecks,” 

but Byers and Norris (2011) did not explain the physical 

or biological mechanisms by which climate change would 

interact with these factors to increase vulnerability. 

Certain climate-related stress factors in mid-Appalachian 

shale barrens—specifically seasonally high temperatures and 

seasonal moisture limitation at the soil surface—are often 

credited with the exclusion of mesic competitors and the 

maintenance of ecosystem integrity, including habitat for 

endemic taxa (Platt, 1951; Keener, 1983; Braunschweig and 

others, 1999; Norris and Sullivan, 2002; Nott, 2006). Higher 

temperatures at a regional level would presumably elevate 

soil surface temperatures in shale barrens on a seasonal basis. 

Some experimental evidence suggests that certain plant taxa 

adapted to heat stress in shale barren habitat are already near 

their heat-tolerance thresholds when subjected to the hottest 

and driest conditions in mid to late summer (Platt, 1951; 

Braunschweig, 1993). Thus, further elevation in seasonally 

high soil surface temperatures might surpass even the capacities of these specially adapted taxa to cope with heat stress. 

An evaluation of the likelihood of this scenario would require 

better identification of the heat-tolerance thresholds of shale 

barren taxa coupled with specific predictions of temperature 

change at the soil surface. 

Hydrologic change would presumably be another avenue 

by which climate change might affect mid-Appalachian 

shale barren ecosystems. As part of a species-level climate 

vulnerability assessment in West Virginia, Byers and Norris 

(2011) predicted that “all habitats are likely to face increased 

drought stress, especially during the summer and early fall.” 

Given that temperature increases have a disproportionate 

effect on evaporative drying in shale barren habitat relative to 

surrounding forests (Platt, 1951), increased seasonal drought 

stress in shale barrens appears to be a likely consequence of 

regional climate change. This increased stress might help 

maintain shale barrens by slowing the encroachment of 

mesic plant species and boosting the competitive advantage 

of drought-adapted endemic taxa. However, the droughttolerance thresholds of most shale barren taxa have not been 

systematically identified, and it is unclear if exacerbated 

seasonal dryness might stress these taxa beyond their coping 

abilities. Existing knowledge of the physiology and life 

cycles of some shale barren taxa does suggest that the timing 

of precipitation will continue to be important because some 

plants time their reproduction to minimize the drought stress 

experienced during the vulnerable germination period (Platt, 

1951). Also, because some shale barren taxa are adapted to 

exploit moisture within the C horizon of soil (Platt, 1951; 

Keener, 1983), hydrologic changes affecting moisture availability at this depth might have more relevance to these taxa 

than would those changes limited to the soil surface. Other 

possible climate-change effects on mid-Appalachian shale 

barrens that could be biologically mediated are even more 

difficult to predict. The possible interactive effects of stress 

regime alteration coupled with climate-driven changes in 

biotic processes—such as exotic species invasions, herbivory, 

and pollination—are largely unknown.

References

Abrams, M., Orwig, D., and Demeo, T., 1995, Dendroecological

analysis of successional dynamics for a presettlement-origin 

white-pine-mixed-oak forest in the southern Appalachians, 

USA: Journal of Ecology, v. 83, p. 123–133.

Allard, H., 1957, Occurrence of the lichen Cetraria islandica

Ach. in Virginia and West Virginia: Castanea, v. 22, no. 2, 

p. 106–109.

Allard, H., and Leonard, E., 1946, Shale barren associations 

on Massanutten Mountain, Virginia: Castanea, v. 11, no. 3, 

p. 71–124.

Ammons, N., 1943, Bryophytes of the Appalachian shale 

barrens: Castanea, v. 8, no. 7/8, p. 128–131.

Artz, L., 1937, Plants of the shale banks of the Massanutten 

Mountains of Virginia: Claytonia, v. 3, p. 45–50.

Artz, L., 1948, Plants of the shale barrens of the tributaries 

of the James River in Virginia: Castanea, v. 13, no. 4, 

p. 141–145.

Bartgis, R., 1993, The limestone glades and barrens of 

West Virginia: Castanea, v. 58, no. 2, p. 69–89.
Chapter E 65

Baskin, J., and Baskin, C., 1988, Endemism in rock outcrop 

plant communities of unglaciated eastern United States—

An evaluation of the roles of the edaphic, genetic and light 

factors: Journal of Biogeography, v. 15, no. 5, p. 829–840.

Bayer, R., and Crawford, D., 1986, Allozyme divergence 

among five diploid species of Antennaria (Asteraceae: 

Inuleae) and their allopolyploid derivatives: American 

Journal of Botany, v. 73, no. 2, p. 287–296.

Braun, E., 1955, The phytogeography of unglaciated eastern 

United States and its interpretation: The Botanical Review, 

v. 21, no. 6, p. 297–375.

Braunschweig, S., 1993, The acclimation ability of the 

shale barren endemic Erigonum allenii to light and heat: 

Blacksburg, Va., Virginia Polytechnic Institute and State 

University, Ph.D. Dissertation.

Braunschweig, S., Nilson, E., and Wieboldt, T., 1999, The 

mid-Appalachian shale barrens, in Anderson, R., Fralish, J.,

and Baskin, J.M., eds., Savannas, barrens, and rock 

outcrop plant communities of North America: New York, 

Cambridge University Press, p. 83–98.

Byers, E., and Norris, S., 2011, Climate change vulnerability 

assessment of species of concern in West Virginia: Elkins, 

W.Va., West Virginia Divison of Natural Resources. 

Clarkson, R., 1966, The vascular flora of the Monongahela 

national forest, West Virginia: Castanea, v. 31, no. 1, 

p. 1–119.

Core, E., 1932, Some aspects of the phytogeography of 

West Virginia: Torreya, v. 32, no. 3, p. 65–71.

Core, E., 1952, The ranges of some plants of the Appalachian 

shale barrens: Castanea, v. 17, no. 3, p. 105–116.

DeSelm, H., and Murdock, N., 1993, Grass-dominated 

communities, in Martin, W., Boyce, S., and Echternacht, A.,

eds., Biodiversity of the southeastern United States—

Upland terrestrial communities: New York, John Wiley

and Sons, Inc., p. 87–141.

Dix, E., 1990, Element stewardship abstract for Appalachian 

shale barrens: Arlington, Va., The Nature Conservancy.

Fenneman, N., 1938, Physiography of the Eastern 

United States: New York, McGraw-Hill.

Fenneman, N., and Johnson, D., 1946, Physical divisions of 

the United States: U.S. Geological Survey map prepared 

in cooperation with the Physiographic Commission, 

scale 1:7,000,000. 

Fleming, G., and Patterson, K., 2012, Natural communities 

of Virginia—Ecological groups and community types: 

Richmond, Va., Virginia Department of Conservation and 

Recreation, Division of Natural Heritage.

Frye, W., 1942, Notes on the flora of the South Branch Valley: 

Castanea, v. 7, no. 8, p. 118–122.

Heikens, A., West, K., and Robertson, P., 1994, Short-term 

response of chert and shale barrens vegetation to fire in 

southwestern Illinois: Castanea, v. 59, no. 3, p. 274–285.

Kalhorn, K., Barrows, E., and LaBerge, W., 2003, Bee 

(Hymenoptera: Apoidea: Apiformes) diversity in an 

Appalachian shale barrens: Journal of the Kansas 

Entomological Society, v. 76, no. 3, p. 455–468.

Keech, D., 1999, Managing invasive weeds on shale barrens: 

Chevy Chase, Md., The Nature Conservancy.

Keener, C., 1983, Distribution and biohistory of the endemic 

flora of the mid-Appalachian shale barrens: The Botanical 

Review, v. 49, no. 1, p. 65–115.

Knobloch, I., and Lellinger, D., 1969, Cheilanthes castanea

and its allies in Virginia and West Virginia: Castanea, v. 34, 

no. 1, p. 59–61.

McCall, R., and Gibson, D., 1999, The regeneration potential 

of a threatened southern Illinois shale barren: Journal of the 

Torrey Botanical Society, v. 126, no. 3, p. 226–233.

Morse, L., 1983, A shale barren on Silurian strata in Maryland: 

Castanea, v. 48, no. 3, p. 206–208. 

NatureServe, 2015, NatureServe Explorer—An online 

encyclopedia of life [web application], ver. 7.1, accessed 

October 7, 2015, at http://www.natureserve.org/explorer.

Nilsen, E., and Orcutt, D., 1996, The physiology of plants 

under stress: New York, John Wiley.

Norris, S., and Sullivan, R., 2002, Conservation assessment 

for the mid-Appalachian shale barrens: Elkins, W. Va., 

West Virginia Divison of Natural Resources.

Noss, R., 2013, Forgotten grasslands of the south: Washington, 

D.C., Island Press. 

Nott, M., 2006, Shale barren rock cress (Arabis serotina)—

A literature review and analysis of vegetation data: Sugar 

Grove, W.V., Navy Information and Operations Command 

and the Institute for Bird Populations.

Platt, R., 1951, An ecological study of the mid-Appalachian 

shale barrens and of the plants endemic to them: Ecological 

Monographs, v. 21, no. 4, p. 269–300.
Rawinski, T., Hickmain, K., Waller-Elling, J., Fleming, G., 

Austin, C., Helmick, S., Huber, C., Kappesser, G., Huber, F.,

Bailey, T., and Collins, T., 1996, Plant communities and 

ecological land units of the Glenwood Ranger District, 

George Washington and Jefferson National Forests: 

Richmond, Va., Virginia Department of Conservation 

and Recreation.

U.S. Geological Survey, National Gap Analysis Program, 

2011, Land cover data portal, accessed December 2015 at 

http://gapanalysis.usgs.gov/gaplandcover/data/download/.

U.S. Forest Service, 2011, Draft environmental impact 

statement for the revised land and resource management 

plan, George Washington National Forest: Roanoake, Va., 

U.S. Department of Agriculture, Forest Service.

West, N.M., Gibson, D.J., and Minchin, P.R., 2009, Characterizing the microhabitats of exotic species in Illinois shale 

barrens: Plant Ecology, v. 200, no. 2, p. 255–265.

Wheeler, A., 1992, Chinaola quercicola rediscovered in 

several specialized plant communities in the southeastern 

United States (Heteroptera: Microphysidae): Proceedings 

of the Entomological Society of Washington, v. 94, no. 2, 

p. 249–252.

Wheeler, A., 1995, Plant bugs (Heteroptera: Miridae) of 

Phlox subulata and other narrow-leaved phloxes in eastern 

United States: Proceedings of the Entomological Society of 

Washington, v. 97, p. 435–451.

Wherry, E., 1953, Shale-barren plants on other geological 

formations: Castanea, v. 18, no. 2, p. 64–65.

66 Insular Ecosystems of the Southeastern United States
Chapter F. High-Elevation Outcrops and Balds of the

Southern Appalachians

Introduction

The summits of even the highest peaks of the southern 

Appalachian Mountains are forested—estimates of a theoretical treeline based on current climate suggest it would be about 

1,300 feet (approximately 400 meters) above the highest peaks 

in the region—yet a number of treeless openings exist within 

the generally forested landscape (Gersmehl, 1973; Cogbill and 

others, 1997; White and Sutter, 1999). These high-elevation 

forest openings, above 4,000 feet (about 1,200 meters), 

include grass and heath balds (fig. F–1) and high-elevation 

rock outcrops (Wiser and White, 1999). Grass balds are 

high-elevation communities dominated by grasses and sedges, 

with a sub-dominant component of forbs, shrubs, and scattered 

trees. Grass balds vary in size, from roughly 2 to 20 acres 

(about 1 to 8 hectares) in Great Smoky Mountains National 

Park to more than 500 acres (more than 200 hectares) on 

Roan Mountain in eastern Tennessee (White and Sutter, 1999). 

Heath balds are high-elevation communities dominated by 

shrubs, typically by ericaceous, acid-tolerant species from 

genera including Rhododendron, Kalmia, and Vaccinium

(Pittillo and others, 1998), and in some cases also including 

Photinia melanocarpa, Viburnum nudum var. cassinoides, 

Clethra acuminata, and Gaylussacia baccata (Whittaker, 

1963). At elevations above 4,000 feet (about 1,200 meters), 

high-elevation rock outcrop communities are forest openings 

characterized by exposed bedrock with sparse vegetation, a 

preponderance of lichens, mosses, and low-growing forbs, 

and characteristic assemblages of endemic species (Wiser and 

others, 1996; Wiser and White, 1999; Schafale, 2012). 

As a set of ecosystems, high-elevation outcrops and balds 

of the southern Appalachians are all rare, resulting in their 

Figure F–1. Round Bald, Pisgah and Cherokee National Forests, 

Mitchell County, North Carolina, and Carter County, Tennessee. 

Photograph by Alan M. Cressler, U.S. Geological Survey.
68 Insular Ecosystems of the Southeastern United States

being ranked G1 to G3. Grass and heath balds are categorized 

together under the Southern Appalachian Grass and Shrub 

Bald Ecological System based on the International Terrestrial 

Ecological Systems Classification (ITESC) used by NatureServe 

(2015), with this ecological system containing several natural 

communities that are grass bald types (all ranked G1) and 

several heath-bald types (ranked G2 or G1) (Fleming and 

Patterson, 2012; Schafale, 2012). The natural communities 

corresponding to high-elevation rock outcrops are ranked G1, 

G2, or G3 (Schafale, 2012).

Geographic Range

High-elevation outcrops and balds are present in the 

Blue Ridge Physiographic Province (Fenneman and Johnson, 

1946) throughout the southern Appalachian region (fig. F–2). 

The southern Appalachian region includes the Blue Ridge 

Mountains, the Unaka Mountains, and the system of basins 

and traverse mountain chains (including the Black Mountains 

and the Great Craggy Mountains) that extend between them, 

covering about 10,000 square miles (about 26,000 square kilometers) in southern Virginia, western North Carolina, eastern 

Tennessee, and northern Georgia (Mark, 1958; Ramseur, 1960). 

Grass balds are present in parts of Virginia, North Carolina, 

and eastern Tennessee, but their distribution is believed to be 

clustered primarily within Great Smoky Mountains National 

Park and on Roan Mountain (White and Sutter, 1999). Within 

Virginia, the geographic extent of grass balds is extremely 

limited and is restricted primarily to an area near the summit 

of Whitetop Mountain (Fleming and Patterson, 2012). Similar 

grass balds also are present in the central Appalachian region 

of West Virginia and Virginia, and although they share many 

similarities to southern Appalachian grass balds in terms of 

climate, disturbance history, and soil characteristics, they are 

floristically different. For example, the balds of West Virginia 

studied by Rentch and Fortney (1997) had a more diverse 

herbaceous layer and a more highly developed shrub component than those in the southern Appalachians. 

Like grass balds, heath balds are also patchily distributed 

throughout the southern Appalachian region, with most 

occurrences in Tennessee and North Carolina. In Virginia, 

heath balds are confined to only a few sites in the Mount 

Figure F–2. Geographic distribution of 

high-elevation outcrops and balds in the 

southern Appalachians. 

Area of map

at right

0 200 KILOMETERS 100

Base modifed from U.S. Department 0 100 200 MILES

of Commerce, Census Bureau,

U.S. States, 2015, Mercator

Auxiliary Sphere Projection

85° 80°

40°

35°

Southern Appalachian Granitic Dome

Figure F-2

TENNESSEE

TENNESSEE

ALABAMA

GEORGIA

GEORGIA

SOUTH

CAROLINA

SOUTH CAROLINA

NORTH

CAROLINA NORTH

CAROLINA

WEST

VIRGINIA

INDIANA OHIO

VIRGINIA

VIRGINIA

KENTUCKY

KENTUCKY

N

EXPLANATION

Southern Appalachian Grass and

 Shrub Bald (Herbaceaous)

Ecological System—Modified from

 U.S. Geological Survey (2011)

Southern Appalachian Grass and

 Shrub Bald (Shrub)

Southern Appalachian Rocky Summit

Blue Ridge Physiographic Province—

 From Fenneman and Johnson (1946)

Great Smoky Mountains National Park

Selected occurrences of high-elevation

 rock outcrops—Based on locations

 sampled by Wiser (1994)
Chapter F 69

Rogers–Whitetop Mountain area and in the Iron Mountains 

(Fleming and Patterson, 2012). Within Great Smoky Mountains National Park, heath balds are distributed primarily in 

the northern and eastern regions of the park (White and others, 

2001). Heath balds have also been documented in the Great 

Craggy Mountains (Thomas and Pittillo, 1987) and in the 

Amphibolite Mountains (Lee, 2004) of North Carolina and on 

Roan Mountain in eastern Tennessee, among other locations.

Generally, rock outcrops above 4,000 feet (about 

1,200 meters) are geographically restricted to the southern 

Appalachians south of the southern border of Virginia 

(Wiser and others, 1996). Within this zone of high-elevation 

rock outcrop distribution, physical habitat conditions show 

geographic variablility. For example, mafic parent material 

is more common in the northern part of the area (north of 

Grandfather Mountain) and generally weathers to form more 

fertile soil with higher magnesium and iron concentrations 

than does the felsic parent material more common in the 

southern part of the area (Wiser and others, 1996).

Physical Geography

Grass balds are present at elevations ranging from approximately 4,600 to 6,200 feet (about 1,400 to 1,900 meters), 

with the majority between 5,200 to 5,800 feet (about 

1,600 to 1,780 meters); heath balds are present from roughly 

3,900 to 6,500 feet (about 1,200 to 2,000 meters) (Mark, 1958). 

Grass and heath balds generally have southern to western 

aspects (Whittaker, 1956). Grass balds form on rounded 

ridgetops, domes, gaps, and southern slopes, which range from 

nearly level to about 15° and in some places greater than 25° in 

slope and are commonly located along the elevational ecotone 

between hardwood forest below and spruce-fir (Picea rubens

and Abies fraseri) forest above (Wells, 1936b; Billings and 

Mark, 1957; Mark, 1958). Heath balds are generally confined 

to exposed ridgetops and steep slopes (Jenkins, 2007).

High-elevation rock outcrops and balds typically develop 

on parent material of Precambrian origin, which can be igneous 

(felsic or mafic), sedimentary, or metamorphic, on substrates 

including granite, gneiss, schist, gabbro, amphibolite, sandstone, sand conglomerates, and quartzites (Cain, 1931; Wiser 

and others, 1996). High-elevation rock outcrop soils are 

generally classified as Entisols or Lithic Haplumbrepts, with 

relatively high organic matter content and thickness of typically 

less than 12 inches (about 30 centimeters) (Wiser and others, 

1996; Wiser and White, 1999). Soils are generally acidic, with 

pH ranging from below 3.5 to circumneutral (Wiser, 1993). 

Soil pH, nutrient levels, and water-holding capacity in these 

outcrops can be related to factors including bedrock type, 

topographic position, and the presence or absence of seepage 

(Wiser and others, 1996; Wiser and White, 1999).

In general, the soil of grass balds is considered to be 

similar to that of the surrounding forest soil (Wiser and White, 

1999); however, individual studies comparing the two have 

produced somewhat mixed results. Mark (1958) reported pH 

values ranging from 3.5 to 5.0 for several grass balds, with 

similar values for surrounding forest soils. Organic matter 

ranged from 8.1 percent to 28.6 percent in the A1 horizon 

for grass balds and from 15.1 percent to 70.8 percent for 

surrounding forest. Chemical analysis of soil from grass balds 

and surrounding forest indicated that both soils are highly 

leached (Mark, 1958). Comparing sites in the Great Smoky 

Mountains, McCracken and others (1962) found that soil from 

a grass bald was largely similar in morphology and chemistry 

to soil from spruce-fir and beech gap sites, but the grass bald 

soil had higher pH, lower exchangeable aluminum and free 

iron, and wider carbon:nitrogen ratios.

Heath bald soils have been classified as podzols 

(McCracken and others, 1962) and as folist histosols (Conkle 

and others, 2003). These soils may be covered with peat 

deposits ranging from thin crusts to several inches in thickness 

and are often covered with a layer of Sphagnum and associated mosses where sufficient moisture exists (Cain, 1931). 

Relative to grass balds, heath bald soils are highly acidic and 

highly organic (Cain, 1931; McCracken and others, 1962; 

Whittaker, 1963). For example, heath bald soils studied by 

Conkle and others (2003) had organic matter content greater 

than 80 percent, had pH below 3.0, and had high aluminum 

saturation and high cation exchange capacity. 

The combination of extreme acidity and generally 

low levels of calcium and potassium in heath bald soils 

suggests excessive leaching (McCracken and others, 1962; 

Whittaker, 1963), a property that may be attributable in part 

to topographic position. Precipitation is abundant at these 

high elevations but is drained quickly through the subsurface 

and down the steep slopes (Whittaker, 1956). Heath balds 

commonly occupy ridgetops, which have been associated 

with acidic soil. For example, Cain (1931) demonstrated 

a pH correlation with topography in the Great Smoky 

Mountains, with ridgetops having the lowest pH (2.9 –3.7), 

valleys having the highest pH (3.7–5.1), and intervening 

slopes having intermediate values, a correlation he attributed 

to increased leaching with elevation. Other factors that likely 

contribute to the acidity of heath bald soils include microclimatic conditions (high moisture, low temperatures, and 

poor aeration) promoting the formation of peat, as well as the 

decomposition of ericad and conifer leaves and the influence 

of Sphagnum and related mosses, which are known to decrease 

the pH of soils in which they grow (Cain, 1931).

Throughout the southern Appalachians, precipitation 

gradients generally increase from northeast to southwest 

(Mark, 1958; Wiser and others, 1996). Regionally, precipitation derives from winter cyclones, summer and autumn 

hurricanes, and summer thunderstorms, such that rainfall 
70 Insular Ecosystems of the Southeastern United States

is reasonably well distributed throughout the year with a 

slight increase during the summer and early autumn (Mark, 

1958). Soil moisture availability typically increases with 

elevation because of increased precipitation and because 

lower temperatures at higher elevations result in lower 

rates of evapotranspiration (Mark, 1958). At the highest 

elevations in the Appalachians—generally those supporting 

spruce-fir forest, above the elevational ecotone with deciduous 

forest—cloud immersion is an important part of ecosystem 

dynamics, occurring on approximately 2 out of every 3 days 

and contributing approximately 45 percent of the annual water 

input (Johnson and Smith, 2008). Cloud immersion affects 

the hydrologic cycle of montane ecosystems not only through 

direct water inputs but also indirectly through effects on 

temperature, relative humidity, and evapotranspiration (Foster, 

2001). Cloud immersion also alters solar irradiance, diffusion, 

and spectral quality of sunlight. These effects are not constant 

across ecosystems; whereas cloud cover decreases irradiance 

in open areas, it enables greater penetration of light into 

forest understories (Johnson and Smith, 2008). Leaf wetness 

resulting from cloud immersion can also have a range of 

consequences for photosynthetic activity and carbon assimilation (Johnson and Smith, 2008). Fog is another important 

environmental factor for these high-elevation ecosystems, 

with effects similar to those of cloud immersion: increased 

humidity, diffusion of solar irradiation, reduced evapotranspiration, and—in winter—possible formation of rime ice 

(Mark, 1958).

The unique water regime of high-elevation balds and rock 

outcrops distinguishes them in part from their counterparts 

at lower elevations. For example, high-elevation rock outcrops 

generally have greater (and less seasonally variable) water 

availability than outcrops at lower elevations due to increased 

precipitation, cooler temperatures, and cloud or fog immersion 

(Schafale and Weakley, 1990; Wiser, 1994). Moisture availability from seepage is also a component of the water regime 

in many high-elevation rock outcrops; however, moisture availability is highly variable even at small spatial scales and can 

range from continuous to only sporadic (Schafale and Weakley, 

1990). Sites perennially wet from seepage have considerably 

different soil chemistry than dry sites, including higher pH, 

sodium, and water-holding capacity (Wiser and White, 1999).

Stress and Disturbance Regimes

High-elevation outcrops, grass balds, and heath balds 

generally share certain environmental conditions, including 

high irradiation, dessicating winds, high rainfall, frequent 

fog, and an important fraction of water input from cloud 

immersion (Crawford, 2007; Johnson and Smith, 2008). Frost 

typically occurs between late September and mid-May, and 

wind speeds in winter can exceed 20 miles per hour (about 

30 kilometers per hour) (Wiser and others, 1996). Whittaker 

(1956) published a schematic diagram of vegetation communities in the Great Smoky Mountains based on elevation 

and topographic position and situated grass balds and heath 

balds in one extreme corner—the highest elevations and most 

exposed and xeric topographic sites. The extent to which 

abiotic stress factors contribute to ecosystem maintenance is 

unclear, however. Furthermore, any discussion of the stress 

or disturbance regimes that currently maintain grass and 

heath balds invariably raises the question of their origins, 

a subject that has been controversial for decades (Lindsay 

and Bratton, 1979b; Wiser and White, 1999; Crawford and 

Kennedy, 2009). 

Both types of balds occupy only a small fraction of 

seemingly appropriate sites within their ranges based on 

topographic and environmental factors (White and Sutter, 

1999; White and others, 2001). Grass balds appear not to 

be edaphically determined or constrained, and their soil is 

generally comparable to that of the surrounding forest (Mark, 

1958; Wiser and White, 1999). Isolated natural disturbances 

such as wind damage from storms, ice or frost damage, and 

insect herbivory may kill individual trees and thus slow 

forest encroachment somewhat, but they are not of sufficient 

magnitude or extent to play a major role in the origin or 

maintenance of grass balds (White and Sutter, 1999). Natural 

fire frequencies are low at this elevation due to moist and 

cool conditions, suggesting that natural fire is unlikely to 

be important in grass bald origin or maintenance (Barden, 

1978; White and Sutter, 1999). Furthermore, Lindsay and 

Bratton (1979a) argued that grass balds do not resemble 

pyric disclimax communities floristically, noting that recently 

burned areas are generally not dominated by Danthonia 

compressa and instead have vegetation typical of secondary 

successional communities at high elevations. Wiser and White 

(1999) also noted that fire does not eliminate woody plants, 

but rather can stimulate the sprouting of Rubus, Vaccinium, 

and other woody species known to encroach onto grass balds. 

Although some early explorers of grass balds, such 

as Harshberger (1903), explained their formation based on 

topographic vulnerability to ice storms, topographic position 

alone cannot explain the origin or maintenance of balds 

because the vast majority of topographically appropriate 

sites within their range are forested (White and Sutter, 1999). 

Likewise, exposure to storms and drying from winds are 

unlikely causative factors for bald formation or maintenance 

because wind velocities are greater near many forested gaps 

and many of the most exposed and windy peaks are forested 

(Gersmehl, 1973). The fact that topographic or environmental 

factors alone cannot maintain grass balds has been empirically 

confirmed by experiments in which transplanted conifers grew 

successfully in grass bald sites (Brown, 1953).

Noting that grass balds are commonly located along 

elevational ecotones between beech–buckeye–red oak forest 

below and spruce–fir forest above, Billings and Mark (1957) 

and Mark (1958) postulated that grass balds are “primarily 

ecotonal or forest margin phenomena.” They maintained that 
Chapter F 71

grass balds were created by past climatic changes that eliminated low-elevation ecotypes of spruce-fir species, followed 

by subsequent cooling in which the deciduous treeline 

retreated downslope faster than the spruce-fir margin could 

invade. Billings and Mark (1957) cited the stress regime of 

grass balds once they have become established (wind, drought, 

temperature extremes, and snow removal) as factors restricting 

the (re)establishment of tree seedlings. Weaknesses of this 

explanation, highlighted by Wells (1961), White and Sutter 

(1999), and others, include the relative scarcity of actual grass 

balds within the “bald susceptible” zone described by Billings 

and Mark (1957), and that it necessarily rests on an inability 

of spruce-fir forest to colonize suitable habitat over the course 

of thousands of years. Furthermore, stands of mixed hardwood 

forest have been found as high as approximately 1,300 feet 

(about 400 meters) above the ecotone, and stands of spruce-fir 

as much as 1,900 feet (about 600 meters) below (Gersmehl, 

1973). Woody encroachment on many nongrazed grass balds 

also contradicts the idea that previous climate change could 

have created permanently deforested zones (Ramseur, 1960; 

Gersmehl, 1973). 

Weigl and Knowles (1995) offered a hypothesis that grass 

balds were originally created during glacial episodes of the 

Pleistocene and were maintained for millennia by grazing 

from large herbivores. Prior to the great megafaunal extinction 10,000 years ago, these herbivores might have included 

mastodon, mammoth, musk ox, and ground sloth; thereafter, 

the primary grazers would have been bison, elk, and deer until 

European settlement, and then livestock thereafter. Weigl and 

Knowles (1995) cited Native American and European historical accounts and the presence of rare, relict, and “endemic” 

plants, although they acknowledged that the grass bald 

endemics also grow on related habitats such as high-elevation 

outcrops. They also pointed to archaeological excavations of a 

variety of Pleistocene megaherbivores less than about 60 miles 

(roughly 100 kilometers) from Roan Mountain and within 

about 180 miles (roughly 300 kilometers) of most grass balds, 

and they highlighted the landscape-level effects of grazing by 

megaherbivores, such as elephants and hippopotamuses, on 

other continents.

Anecdotal evidence about possible anthropogenic origins 

of the grass balds is mixed and contradictory. For example, 

some long-time residents living near grass balds have claimed 

that their ancestors had cleared those areas; however, some 

Cherokee legends also mention the grass balds, or at least 

treeless areas at high elevations (Gersmehl, 1973; Lindsay and 

Bratton, 1979b; Weigl and Knowles, 1995). Wells (1936a, b, 

1961) argued that grass balds were summer camp sites and 

game lures maintained by Native Americans who systematically removed seedlings of fire cherry (Prunus pensylvanica) 

that otherwise would have initiated succession to forest. 

Others, including Brown (1941), Mark (1958), and Lindsay 

and Bratton (1979a), disputed this argument and cited the 

lack of Native American artifacts at grass bald sites among 

their critiques. Gersmehl (1973) asserted that grass balds are 

“19th century cultural relics” that were deliberately created 

by European settlers and whose persistence is dependent 

upon frequent grazing and fires, a hypothesis supported by 

empirical work by Barden (1978). Most authors acknowledge 

that at least some grass balds were cleared by settlers and 

that livestock grazing took place on most grass balds up 

until grazing cessation was initiated by land protection at 

various points in the early to mid-20th century, though this 

may not necessarily mean that all grass balds are of recent 

anthropogenic origin (Weigl and Knowles, 1995; White and 

Sutter, 1999). Evidence for grazing includes early photographs 

(Lindsay and Bratton, 1979b), early botanical accounts 

(White and Sutter, 1999), and fencepost remnants (Crawford 

and Kennedy, 2009).

Additionally, authors including Whittaker (1956) have 

noted general floristic similarities between grass balds and 

cleared fields at similar elevations. Based on analysis of 

vegetation data, Lindsay and Bratton (1979a) found that 

grass bald sites within the Great Smoky Mountains tended 

to separate in intermediate positions between continually 

disturbed communities (such as roadsides and campgrounds) 

and burn scars and mature forest stands. Despite disagreement 

over the origins of grass balds, researchers generally agree 

that grass balds are not currently being formed, indicating 

that whatever processes created the grass balds are probably 

different from the factors or processes responsible for their 

maintenance (Mark, 1958; Lindsay and Bratton, 1979b; 

Rentch and Fortney, 1997; White and Sutter, 1999).

Although a robust literature exists on the competing 

theories regarding origins and maintenance of grass balds, 

comparatively little attention has been paid to the creation 

and maintenance of heath balds. However, at least two 

analytical studies have provided evidence that heath balds 

predated European arrival by at least 500 years, possibly by 

several thousand years. Radiocarbon dating of base organic 

deposits in heath balds of Great Smoky Mountains National 

Park has revealed ages ranging from 1,000 to 3,000 years 

(Jenkins, 2007). At another heath bald site, Flat Laurel Gap 

near Mount Pisgah, analysis of plant macrofossils and pollen 

grains indicated the presence of heath shrubs for at least the 

past 3,000 years (Shafer, 1986). Whittaker (1956) favored 

an explanation for heath bald formation based on primary 

succession from bare rock, suggesting that succession had 

slowed on exposed ridges and that once heath species were 

well established, they effectively excluded tree establishment. 

White and others (2001) discounted this idea and argued 

that heath balds were more likely the result of localized 

disturbance to spruce-fir forest with an existing heath species 

understory. Some studies have suggested that heath balds may 

be created and maintained by the interaction of disturbance 

and chronic abiotic stress. For example, using a GIS-based 

analytical approach, White and others (2001) found that heath 

balds were associated with upper slopes, convex topography, 

and acidic rock type and with sites that had experienced 

burning following commercial logging activities. 
72 Insular Ecosystems of the Southeastern United States

Community Types and Vegetation 

Dynamics

The highest elevation regions of the southern Appalachians are occupied by spruce-fir forest, which extends down 

to elevations of about 5,200 feet (roughly 1,600 meters) on 

southern and southwestern slopes and approximately 4,750 feet 

(roughly 1,450 meters) on northern and northeastern slopes, 

with a downslope gradient from fir to spruce dominance 

(Mark, 1958). At the lowest elevations of spruce-fir forest 

is an ecotone to high-elevation deciduous forest (Ramseur, 

1960). Near this ecotone are trees of the deciduous forest at 

the upper reaches of their elevation tolerance, which exhibit 

dwarfing, stunting, and a widely spaced distribution. This 

physiognomy led some observers to designate these areas 

as “sub-alpine orchard associations” (Mark, 1958) or simply 

“orchards” (Cain, 1931), corresponding to the woodland 

community dominated by Crataegus punctata and C. flabellata

and termed the Southern and Central Appalachian Hawthorn 

Orchard by NatureServe (2015). Grass balds, heath balds, and 

high-elevation rock outcrops generally are present within or 

above the elevational band of this ecotone, such that the forests 

surrounding them typically include spruce-fir forests, northern 

hardwood forests, and oak forests (Wiser and others, 1996).

According to ITESC classification, grass balds and heath 

balds are grouped together under the Southern Appalachian 

Grass and Shrub Bald ecological system; however, in addition 

to their contrasting physiognomy, grass balds and heath balds 

differ in a number of important ways. For example, species 

richness is generally much higher in grass balds than in heath 

balds (Ramseur, 1960; Wiser and White, 1999), and the two 

bald types are governed by differing successional dynamics, 

as explained below. Grass and heath balds, as well as highelevation rock outcrops, commonly form patches and compositional gradients from one to another such that community 

delineation at a particular site may not be straightforward 

(Schafale and Weakley, 1990). Some sites contain mosaics 

of heath vegetation interspersed with grass bald vegetation 

(fig. F–3). Some grass and heath balds are completely 

surrounded by hardwood or spruce-fir forest, whereas 

others grade into other high-elevation forest openings such 

as cliff faces, seepage areas, burn scars, and man-made 

clearings (Tucker, 1972; White and Sutter, 1999). Similarly, 

high-elevation rock outcrop communities in some places are 

surrounded by forest (generally spruce-fir at higher elevations 

and oak or mixed hardwood forest at lower elevations), but 

may also be embedded within, and grade into, grass balds, 

heath balds, and other forest openings (Tucker, 1972; Johnson, 

1996; Wiser and others, 1996). 

Grass Balds

As the name suggests, grass balds are graminoiddominated forest openings (fig. F–4). Important and 

characteristic species include Danthonia compressa, 

D. spicata, Potentilla canadensis, Deschampsia flexuosa, 

Rumex acetosella, Rubus canadensis, Vaccinium vacillans, and 

several Aster, Solidago, Carex, and Viola species (Mark, 1958; 

Lindsay and Bratton, 1979a, b; Pittillo and others, 1998; White 

and Sutter, 1999). Three associations described by Schafale 

(2012) are grass bald subtypes: the grass subtype (distinguishable by the dominance of Danthonia compressa), the sedge 

subtype (dominated by Carex species), and the alder subtype 

(a unique association that is confined to the highlands near 

Roan Mountain, consisting of Alnus viridis ssp. crispa, Carex

species, Danthonia compressa, and other herbaceous taxa). All 

three associations are ranked G1. Grass balds support relatively 

high species richness for high-elevation communities, with 

researchers often reporting more than 150 (sometimes more 

than 200) vascular plant species for a single site, most of which 

are grasses and sedges (White and Sutter, 1999; Wiser and 

White, 1999). Adding to their species richness is the presence 

of ecotypic populations of many forest herb species, in addition 

to other taxa that prefer or require open habitat (Whittaker, 

1956). In his floristic survey of high-elevation communities, 

Ramseur (1960) documented more than twice as many species 

in grass balds (236) as in heath balds (102), and grass balds 

had greater species richness than any other high-elevation 

community he surveyed, although they accounted for less than 

5 percent of total land area.

Within grass balds, the relative dominance of species has 

been shown to fluctuate seasonally (Ramseur, 1960), but the 

relative coverage of herbaceous versus woody species has also 

been shown to change directionally over the course of years, 

suggesting successional processes that may be attributable to 

changes in disturbance patterns such as grazing or drought 

(Mark, 1958). Investigations of vegetation dynamics in grass 

balds—and specifically the processes of succession and forest 

encroachment—have used a variety of methods, including 

aerial photography, historical records, dendrochronology, and 

geographic information systems (Crawford, 2007). Although 

not all findings have been in agreement, a general pattern 

has emerged that grass balds are highly vulnerable to woody 

encroachment and that, following the cessation of disturbance 

(grazing, mowing, or cutting), they typically succeed to forest 

(Mark, 1958; Ramseur, 1960; White and Sutter, 1999). Using 

aerial photography and tree core sampling on two grass 

balds in Great Smoky Mountains National Park, Lindsay and 

Bratton (1980) determined that nearly all trees had become 

established since the cessation of grazing, and they calculated 

approximately linear decreases in nonforested areas ranging 

from about 0.1 to 0.25 acre (about 0.04 to 0.1 hectare) per year. 

Primary species responsible for encroachment onto grass balds 

in Great Smoky Mountains National Park include Crataegus, 

Amelanchier, Vaccinium, Rhododendron, Picea, Abies, Fagus, 

Rubus, and Quercus species (Lindsay and Bratton, 1979b, 1980; 

Weigl and Knowles, 1995). Grass balds on Roan Mountain 

were encroached upon by Rhododendron, Picea, Abies, and 

Alnus species (Brown, 1941). As part of a long-term monitoring project at a grass bald site in the Balsam Mountains 

in North Carolina, researchers observed a bald convert from 
Chapter F 73

Figure F–3. Grass bald and heath bald mosaic at Round Bald, Pisgah and Cherokee National Forests, Mitchell 

County, North Carolina, and Carter County, Tennessee. Photograph by Alan M. Cressler, U.S. Geological Survey.

Figure F–4. Grass bald at Round Bald, Pisgah and Cherokee National Forests, Mitchell County, North Carolina, 

and Carter County, Tennessee. Photograph by Alan M. Cressler, U.S. Geological Survey.
74 Insular Ecosystems of the Southeastern United States

predominantly Danthonia compressa to primarily tree and 

shrub cover (Abies fraseri and Picea rubens) in only a few 

decades (Sullivan and others, 2010), with tree cover increasing 

88 percent and shrub cover increasing 239 percent in the 

first 5 years of the study (Sullivan and Pittillo, 1988). Using 

tree-ring analysis, Crawford and Kennedy (2009) noted that a 

dramatic increase in establishment of Quercus rubra coincided 

temporally with the cessation of grazing at Craggy Gardens, 

North Carolina. 

In addition to disturbance regimes, factors influencing 

succession trajectories and speeds in grass balds may include 

environmental conditions created by local topography. For 

example, Brown (1941) noticed that invasion of a grass bald on 

Roan Mountain by Rhododendron was more rapid in depressions and slower in more elevated parts of the bald, suggesting 

that soil moisture might play a role in determining succession 

speed. Physiological and life-history traits of invading species 

also shape succession dynamics. Woody invasion of grass balds 

by root-shooting and clonal expansion may be more efficient 

than by seed dispersal because seedling densities in grass balds 

are low and establishment from seed is inhibited by the dense 

herbaceous vegetation (White and Sutter, 1999).

Brown (1941) observed substantial encroachment into a 

grass bald on Roan Mountain by Rhododendron catawbiense

and proposed a model of succession from bare rock to grass 

bald to heath bald to spruce-fir climax. Later generations of 

researchers, including White and others (2001), provided 

evidence that this model is highly oversimplified and have 

disputed the notion that succession of grass balds produces 

heath balds. However, encroachment of ericaceous shrubs 

into grass balds can produce what Ramseur (1960) referred 

to as shrub balds: communities dominated by heath shrubs 

but with a more diverse herbaceous component than heath 

balds, at least in the initial stages of encroachment. White and 

Sutter (1999) suggested that nonheath shrub balds may indeed 

represent a possible seral stage in the succession of grass balds 

to forest. 

Heath Balds

Ramseur (1960) distinguished two primary types of heath 

balds: heath slicks—found on steep ridges with exposed slate 

and dominated by Rhododendron carolinianum—and rhododendron gardens—found on gaps, slopes, and summits and 

dominated by R. catawbiense. According to current ITESC 

classification, several similar and related high-elevation associations are classified as heath bald subtypes, all of which are 

ranked G1 or G2 (Schafale, 2012). These include heath balds 

categorized primarily by dominant vegetation (the Catawba 

rhododendron subtype, the Carolina rhododendron subtype, 

the southern mixed subtype, and the sand myrtle subtype) and 

by occurrence on slate substrate (the slate subtype). In general, 

heath balds tend to be species poor, dominated by relatively 

few species of ericaceous shrubs with a sparse herbaceous 

component (Whittaker, 1956; Wiser and White, 1999; Jenkins, 

2007). Many ericaceous species can exist as an understory 

in the surrounding spruce-fir forest, such that heath balds 

may grade into a forest community type with a heath shrub 

understory—what Whittaker (1963) called a “forest heath.” 

Disturbance to trees—whether by wind, debris slides, fire, 

or logging—may thus produce heath bald communities by 

removing canopy and exposing the understory (Peter White, 

University of North Carolina, written commun., 2014).

Compared to grass balds, succession in heath balds 

is considerably slower and is limited by a specific suite of 

environmental factors. Whittaker (1956, 1963) postulated that 

heath balds were stable “topographic climax” communities, 

a hypothesis supported by White and others (2001) who 

concluded based on aerial photography that heath balds in 

the Great Smoky Mountains National Park had not changed 

in area between the 1930s and 1980s. Other investigators, 

however, have documented forest encroachment into heath 

balds, albeit at rates less readily noticeable than for grass 

balds (Brown, 1941; Thomas and Pittillo, 1987). It is now 

generally accepted that succession to forest does occur in 

heath balds but is very slow due to extremely low pH (which 

may result in aluminum toxicity to nonadapted species), thick 

leaf litter, shallow soil, and shady conditions at the soil surface 

resulting from the very thick shrub cover (Pittillo and others, 

1998; Wiser and White, 1999; Jenkins, 2007). Using dendrochronology at a bald in Craggy Gardens, Buncombe County, 

North Carolina, Crawford (2007) demonstrated that Quercus 

rubra encroachment was confined primarily to an area free of 

heath species, suggesting that heath species serve as a barrier 

that slows tree establishment. On the other hand, Thomas and 

Pittillo (1987), also working at Craggy Gardens, documented 

the encroachment of Fagus grandifolia into a heath bald 

dominated by Rhododendron catawbiense at a rate of approximately 10 feet (about 3 meters) per year, enabled primarily by 

root sprouting believed to circumvent the allelopathic defenses 

of Rhododendron. 

High-Elevation Outcrops

High-elevation outcrops include communities characterized by exposed rock surfaces colonized by lichen and mosses 

(fig. F–5) as well as communities that also include graminoids, 

forbs, and shrubs (fig. F–6). Community composition and 

the presence of rare plant species in high-elevation rock 

outcrops depend in part on elevation, surrounding vegetation, 

and the age of the outcrops. Wiser and others (1996) used 

hierarchical clustering methods based on floristic data to 

distinguish nine plant community types of rock outcrops, 

with these community types correlating strongly to elevation, 

bedrock type, and surrounding vegetation composition. Peter 

White (University of North Carolina, written commun., 2014) 

distinguished between very steep outcrops that were likely 

never forested due to their slope and those created by debris 

avalanches, which tend to occur at steep, concave locations 

at the headwaters of first-order streams. According to ITESC 
Chapter F 75

Figure F–5. Rock outcrop at Sassafrass Mountain, Nantahala National Forest, Jackson County, North Carolina. 

Photograph by Alan M. Cressler, U.S. Geological Survey.

Figure F–6. High-elevation rock outcrop community, Craggy Dome, Blue Ridge Parkway, Buncombe County, 

North Carolina. Photograph by Alan M. Cressler, U.S. Geological Survey.
76 Insular Ecosystems of the Southeastern United States

classification, associations belonging to the High Elevation 

Rocky Summit include the typic subtype (which describes 

most outcrops within the ecosystem range), the little bluestem 

subtype (present on a mafic or felsic substrate with perennial 

seepage and sparse graminoid and forb cover), the high-peak 

subtype (present at only the highest elevations and generally 

lacking lower-elevation species such as Kalmia latifolia and 

Krigia montana), and the Anakeesta subtype (documented 

only on slate in the Anakeesta Formation in the Great Smoky 

Mountains, and having up to 80 percent exposed bedrock). 

Associations within this group generally are found in 

geophysical contexts characterized by irregular and fragmented rock with numerous crevices (Schafale and Weakley, 

1990) and are all ranked G1 with the exception of the High 

Elevation Rocky Summit typic subtype, which is ranked G2. 

Two associations of the Southern Appalachian Granitic Dome 

ecological system are also high-elevation outcrop types, 

forming on smooth continuous rock surfaces: the highelevation granitic dome and the high-peak lichen subtype. The 

ninebark subtype of the Southern Appalachian Montaine Cliff 

and Talus ecological system may also be included as a highelevation outcrop, and is characterized by sparse vegetation on 

vertical amphibolite cliffs (Schafale, 2012). The high-elevation 

outcrop barrens (black chokeberry igneous/metamorphic 

type) is a component of the Southern Appalachian Grass 

and Shrub Bald ecological system present in Virginia and 

possibly North Carolina and is characterized by shrub thickets, 

herbaceous mats, and lichen-covered rock surfaces on strongly 

convex upper slopes and summits (Schafale, 2012). Vegetation 

dynamics in high-elevation outcrops appear to be characterized primarily by cyclic primary succession (Schafale, 2012) 

because slowly accumulating soil mats supporting patches of 

woody species are sloughed off or washed away by localized 

disturbance events such as tree falls, storms, exfoliation, and 

loss of individual boulders (Wiser and White, 1999).

Contributions to Regional Biodiversity

Wiser and White (1999) stated that high-elevation 

outcrops of the southern Appalachians “support one of the 

richest floras of rare species of any regional habitat.” These 

outcrops and associated balds sustain many rare plant taxa 

(table 3–5 in appendix 3) and provide habitat to a number of 

animal taxa of conservation concern.

Endemic and Biogeographically Disjunct Taxa

Rare plant taxa in high-elevation outcrops and balds 

include taxa that are primarily restricted to high-elevation 

outcrops, as well as those primarily restricted to the suite of 

high-elevation open habitats including grass balds, heath balds, 

outcrops, seepage areas, and landslide scars (Wiser, 1994; 

Wiser and White, 1999). These habitats sustain a number of 

taxa that are State-listed as endangered within the southern 

Appalachian region, several of which are G1- or G2-ranked 

narrow endemics of high-elevation outcrops: Geum radiatum

(fig. F–7A), Calamagrostis cainii, Houstonia purpurea var. 

montana (fig. F–7B), and Solidago spithamaea (Wiser and 

White, 1999). Additionally, Liatris helleri was considered 

to be a narrow endemic to North Carolina outcrops prior to 

a broadening of the concept by Nesom (2005) to include 

occurrences recognized as Liatris turgida and L. pilosa in 

West Virginia and Virginia. Houstonia purpurea var. montana

and Geum radiatum are federally listed endangered taxa, and 

Liatris helleri and Solidago spithamaea are federally listed as 

threatened taxa. In addition to vascular plants, high-elevation 

rock outcrops also support at least 15 globally rare or regionally rare bryophyte and lichen taxa, including Gymnoderma 

lineare (fig. F–7C ), Marsupella emarginata var. latiloba, 

Metzgeria furcata var. setigera, Polytrichum appalachianum, 

Xanthoparmelia monticola, and Cephaloziella obtusilobula

(Wiser and White, 1999). As pointed out by Wiser and 

others (1998), high-elevation outcrops are often relatively 

inaccessible to botanists, resulting in “woefully incomplete” 

knowledge of rare-species distributions. Thus, new species 

have been described relatively recently, such as by Anderson 

(1987), and others may exist that are yet to be described.

Grass and heath balds also support taxa of conservation 

concern. No known taxa are strict endemics to grass balds, but 

grass balds do provide habitat for a number of regionally rare 

plant species and species that might be called “near endemics” 

because they are endemic to the southern Appalachian region 

and are present in open habitats (rock outcrops, cliffs, and 

seeps) as well as in grass balds (Lindsay and Bratton, 1979b; 

Weigl and Knowles, 1995; White and Sutter, 1999). Heath and 

shrub balds also have no strict endemics, though in some cases 

they provide habitat for regionally and globally rare taxa such 

as Geum geniculatum and Alnus viridis ssp. crispa (Whittaker, 

1956; Clarkson, 1960). Evidence suggests that rare plants 

endemic to high-elevation outcrops may have invaded grass 

balds during the process of bald formation, which provided 

newly suitable habitat to the outcrop endemics (Wiser and 

White, 1999). Thus, the richness of rare taxa supported by 

individual grass balds may be influenced by proximity to 

high-elevation rock outcrop sites (Wiser and White, 1999) 

and by geographic size. The grass balds of Roan Mountain 

are the largest and most geographically expansive in the 

region, supporting 19 state or federally listed or candidate 

taxa (Weigl and Knowles, 1995). Applying concepts from 

island biogeography to isolated high peaks in the southern 

Appalachians, White and others (1984) produced results 

consistent with the species-area relationship for rare plants in 

these ecosystems (larger “island” areas had greater numbers 

of rare plant taxa) and the habitat-heterogeneity hypothesis 

(topographic diversity was related to rare-species richness). 

Several authors have noted biogeographically disjunct 

distributions for a number of species of high-elevation 

outcrops and balds, with population centers either in northern 

regions (such as alpine zones of northern New England 

and artic or boreal zones of Canada) or else in the Coastal 
Chapter F 77

Figure F–7. Characteristic taxa of high-elevation outcrops and 

balds: A, Geum radiatum, Roan Mountain, Pisgah National Forest, 

Mitchell County, North Carolina; B, Houstonia purpurea var. 

montana, Roan Mountain, Pisgah National Forest, Mitchell County, 

North Carolina; C, Gymnoderma lineare, Chattahoochee National 

Forest, Rabun County, Georgia. Photographs by Alan M. Cressler, 

U.S. Geological Survey.

A

C

B
78 Insular Ecosystems of the Southeastern United States

Plain (Tucker, 1972). Taxa with arctic or alpine affinities 

present in high-elevation outcrops and balds in the southern 

Appalachian region include Alnus viridis ssp. crispa, Huperzia 

appalachiana, Juncus trifidus, Minuartia groenlandica, 

Trichophorum caespitosum, Sibbaldiopsis tridentata, Agrostis 

mertensii, and Trisetum spicatum (Mark, 1958; Ramseur, 

1960; Wiser, 1994; Wiser and others, 1996; Wiser and White, 

1999). While all of these taxa are secure within their northern 

ranges, they are State-rare within the southern Appalachian 

region and thus contribute to regional biodiversity. For some 

species, their presence in the Appalachian region represents 

a profound disjunction. For example, the population of Alnus 

viridis ssp. crispa on Roan Mountain is disjunct by more than 

740 miles (approximately 1,200 kilometers) from a northern 

range that extends from the mountains of New England to 

Labrador and Alaska (Clarkson, 1960). In addition to vascular 

plants, many lichens and bryophytes also display disjunction, 

with population centers as far away as Japan (Dey, 1978). 

The presence of so many plant and lichen taxa in 

southern Appalachian outcrops, balds, and adjacent forests 

that are disjunct from treeless arctic or alpine habitats far to 

the north suggests that these mountains served as refugia for 

relic populations derived from Pleistocene alpine flora that 

moved southward in advance of the Wisconsin ice sheet and 

later retreated north during deglaciation (Ramseur, 1960; Dey, 

1978; Wiser and others, 1998). Support for this hypothesized 

dynamic between vegetation and climate derives primarily 

from paleoecological evidence, including pollen and macrofossil deposition, together with the mapping of periglacial 

geomorphic features such as stone polygons and stone stripes 

(Clark, 1968; Delcourt and Delcourt, 1985, 1988). Based 

on these lines of evidence, climatic treelines in the southern 

Appalachian region were estimated to have been present 

near 4,600 to 4,900 feet (about 1,400 to 1,500 meters) above 

sea level during the last glacial maximum (approximately 

18,000 years before present), such that areas at higher 

elevations were alpine tundra (Delcourt and Delcourt, 1985, 

1988; Cogbill and others, 1997). Pleistocene relic taxa with 

arctic or alpine affinities are believed to have survived in 

high-elevation areas due to a competitive advantage in cool, 

exposed microclimates as regional climatic conditions warmed 

(Dey, 1978; Pittillo and others, 1998). This hypothesis is 

supported by findings that the minimum elevation for most 

relict species populations in the southern Appalachians was 

roughly 1,100—3,600 feet (about 350–1,100 meters) higher 

than for relict species in the northern Appalachian region 

(Wiser, 1998). 

Animal Taxa of Conservation Concern

High-elevation outcrops and balds provide habitat 

to a number of rare vertebrates. Rare mammals include 

Corynorhinus townsendii virginianus (Virginia big-eared bat), 

which is federally listed as endangered, Neotoma magister

(Allegheny woodrat), and N. floridana (eastern woodrat). 

Aneides aeneus (green salamander) is State-rare throughout 

the southern Appalachian region and typically is present 

at lower elevations, but has been documented on outcrops 

above 3,900 feet (about 1,200 meters) (Wiser and White, 

1999). High-elevation balds and outcrops are important 

to regional avian diversity in the southern Appalachians 

because they contribute to regional habitat diversity in terms 

of microclimate, vegetation structures, and food supplies 

(Kendeigh and Fawver, 1981). Roan Mountain, for example, 

which contains the largest sized grass balds in the region, also 

has exceptionally high avian diversity and supports State-rare 

bird species including Plectrophenax nivalis (snow bunting) 

(Lee, 2004). Other avian taxa of conservation concern 

supported by high-elevation balds and outcrops include 

Falco peregrinus (peregrine falcon) and Aquila chrysaetos

(golden eagle) (Wiser and White, 1999). However, some 

evidence indicates that woody encroachment is beginning 

to degrade the special avian habitat provided by grass balds. 

McConnell and McConnell (1983) found differences in bird 

community structure in the Unicoi Mountains over a 36-year 

interval, and Haney and others (2001) noted declines in 

particular species associated with open habitat over a 50-year 

interval in the same area. Both research teams cited habitat 

changes as possible explanatory factors, specifically taller 

grass, increased shrub cover, and forest encroachment in the 

absence of grazing.

Information about diversity of invertebrate species 

associated with high-elevation balds and outcrops offers varied 

perspectives. Whittaker (1952) found grass and heath balds 

to support relatively low insect productivity and diversity 

compared to forested communities in the Great Smoky Mountains, and he noted a similarity in insect species composition 

between grass balds and alpine meadows of Eurasia. Dourson 

and Langdon (2012) found that species richness of land snails 

decreased with elevation in the Great Smoky Mountains, 

although heath balds were relatively species-rich for their 

elevation, a somewhat surprising finding given their acidic soil 

conditions and low calcium availability. Conversely for spiders, 

Toti and others (2000) observed much greater species richness 

in grass bald sites than in heath bald sites. Aiken and Coyle 

(2000), who documented microhabitat segregation between two 

Tetragnatha spider species in grass balds, suggested that finescale habitat heterogeneity within grass balds can be important 

to arthropod distribution. Van Pelt (1963) suggested that grass 

and heath balds are important in the distribution of many ant 

species within the southern Blue Ridge. Globally or regionally 

rare arthropod taxa supported by high-elevation outcrops and 

balds include insects such as Trimerotropis saxatilis (lichen 

grasshopper) and Spharagemon saxatile (ledge grasshopper) 

and spiders such as Hypochilus coylei and H. sheari (two 

lampshade weaver species) (Wiser and White, 1999).
Chapter F 79

Conservation Considerations

High-elevation outcrops and balds have been foci for 

conservation and restoration efforts because of their biodiversity contributions and cultural importance (Johnson, 1996; 

White and Sutter, 1999; Wiser and White, 1999). These efforts 

are complicated by uneven levels of scientific understanding 

across ecosystem types and, in some cases, by ideological 

differences concerning appropriate management practices.

Threats to Ecosystem Integrity

Damage from recreational use is an often-mentioned 

threat to high-elevation outcrop communities (Tucker, 1972; 

Wiser, 1994; Wiser and others, 1996; Collins and others, 

2001; Fleming and Patterson, 2012). High-elevation outcrops 

offering scenic views are popular destinations. As a result, 

trampling by hikers and rock climbers is a major threat to 

these communities. Even some protected sites that allow 

public access have been heavily damaged by trampling, with 

concomitant declines in rare plant populations (Johnson, 1996; 

Wiser and White, 1999). Through an intentional program of 

overlook design and trail rerouting, the National Park Service 

was able to effect a 97-percent reduction in visitor trampling at 

rock outcrops in Craggy Gardens in North Carolina (Johnson, 

1996); however, the rate of vegetation recovery in previously 

trampled areas is unclear (Bart Johnson, University of Oregon, 

written commun., 2014). While some high-elevation outcrops 

and balds are on protected land, others lack legal protection. 

Thus, depending on their degree of isolation and accessibility, 

some outcrops and balds may be threatened by development 

(Wiser and White, 1999; Collins and others, 2001). 

Air pollution may also threaten some species of highelevation outcrop communities. In particular, vegetation can 

be affected by exposure to ozone, smog haziness that alters 

irradiance and spectral quality of light, pH alteration from acid 

precipitation (which may lead to aluminum toxicity and cation 

leaching), and the deposition of nitrate, sulfate, and particulate 

matter (Lefohn and others, 1997; Wiser and White, 1999; 

Jenkins, 2007).

Exotic and invasive species also threaten the integrity 

of some outcrops and balds. Within Great Smoky Mountains 

National Park, staff are currently managing approximately 

50 exotic plant species that threaten natural communities 

through hybridization or competitive displacement (Jenkins, 

2007). Invasive weedy species mentioned as a threat to 

high-elevation rock outcrop communities in Virginia include 

Poa compressa and Rumex acetosella (Fleming and Patterson, 

2012). Floristic analysis by Wiser (1994) and Wiser and 

White (1999), however, indicated exotic species to be only 

a relatively minor component of rock outcrop vegetation in 

Tennessee and North Carolina. In some cases, exotic pest 

species can pose a threat to high-elevation ecosystems such as 

grass balds. In response to ecosystem damage from the rooting 

activities of Sus scrofa (wild hogs), a hog-control program 

was begun in 1960 in Great Smoky Mountains National Park 

in which animals were trapped and killed (Howe and Bratton, 

1976; Peine and Farmer, 1990). Adelges piceae10 (balsam 

woolly adelgid), an exotic insect species, began causing 

extensive Fraser fir mortality as early as the 1970s (Lindsay 

and Bratton, 1979a). Little is known about the effects of this 

Fraser fir mortality on nearby communities including highelevation outcrops and balds, although fir mortality is expected 

to contribute to soil erosion and increased soil drying through 

changes to microclimate and localized wind patterns (Wiser 

and White, 1999).

Although thin-to-absent soils generally exclude woody 

vegetation in high-elevation outcrops, woody encroachment 

and succession to forest are considered to be major threats 

for grass balds, and to a lesser extent for heath balds. Woody 

encroachment onto grass balds acquired in the early 1930s by 

Great Smoky Mountains National Park became apparent as 

early as the late 1930s and was recognized as a management 

problem in the 1950s (Lindsay and Bratton, 1979b). Active 

maintenance is underway at 2 grass balds in Great Smoky 

Mountains National Park, out of the 20–30 balds believed to 

exist or have existed within the park, and at the 6 grass balds 

on Roan Mountain (White and Sutter, 1999). 

Conservation Strategies

High-elevation outcrop and bald communities in the 

southern Appalachians are present on protected land in three 

States. Great Smoky Mountains National Park, Cherokee 

National Forest, Roan Mountain State Park, and land owned 

by the Southern Appalachian Highlands Conservancy are in 

Tennessee; Jefferson National Forest and Grayson Highlands 

State Park are in Virginia; and Nantahala National Forest, 

Pisgah National Forest, and land owned by The Nature Conservancy are in North Carolina (DeSelm and Murdock, 1993). 

Some concerns have arisen as to whether the number of grass 

balds currently under protected land ownership is sufficient. 

Because of the distribution patterns of rare species across grass 

balds within their range, a relatively large number of grass 

balds would need to be preserved if the full suite of rare species 

they sustain is to be conserved (Wiser and White, 1999). 

Given the well-documented patterns of woody 

encroachment into grass balds, the question of whether they 

are of natural or anthropogenic origin is important not only 

for scientific reasons but also because of the management 

implications for sites on protected land. If these ecosystems 

are anthropogenic in origin and maintenance, then they do not 

10Species name according to Encyclopedia of Life (2016).
80 Insular Ecosystems of the Southeastern United States

qualify as natural ecosystems according to the National Park 

Service, and their management would more appropriately be 

as “historical areas,” necessitating a different set of management practices than those for natural ecosystems (Lindsay and 

Bratton, 1979b; Weigl and Knowles, 1995). However, even the 

management of grass balds as “historical areas” is complicated 

by uncertainty surrounding their age, and the question of what 

historical period to use as a management goal (White and 

Sutter, 1999). Mounting evidence that grass balds have been 

maintained (if not also created) primarily through anthropogenic disturbance has created some ideological division 

among conservationists, with some favoring preservation of 

the grass balds through continued human interference because 

of their aesthetic and historical value, habitat for rare species, 

and species richness, while others argue that the balds should 

be allowed to undergo succession to forest because they 

are essentially human artifacts on the landscape (White and 

Sutter, 1999).

If grass balds are to be actively managed and preserved, 

managers face a number of considerations in combatting 

woody encroachment. Best practices include hand cutting 

of woody species, targeted herbicide application to control 

sprouting in Vaccinium and Rubus species, and seeding with 

native plants (Wiser and White, 1999). Lindsay and Bratton 

(1979b) evaluated a number of management options for maintaining grass balds, such as burning, hand cutting, mowing, 

and grazing, and concluded that although cutting and grazing 

would be the most “historically correct” management strategy, 

a more practical approach would involve cutting followed by 

mowing and (or) burning. However, Jenkins (2007) noted that 

burning may actually accelerate woody encroachment in grass 

balds because it stimulates sprouting from the root collar of 

burned trees; he suggested that grazing and cutting be used 

instead. If prescribed burns are to be used to maintain balds, 

the appropriate burn frequency depends in part on the invading 

species that are being controlled and the type of bald being 

maintained. Rapidly sprouting species such as Vaccinium 

pallidum and Diervilla sessilifolia in grass balds require 

frequent burning, whereas heath balds require less frequent 

burning (Barden, 1978). Although grazing and browsing 

by goats have been widely suggested as a mechanism for 

management of grass balds to prevent succession, the effects 

of grazing and browsing on rare and regionally endemic plant 

populations have largely not been investigated. One exception 

is Powell (2011), who found that goat browsing did not affect 

morphology, reproduction, or vigor of Lilium grayi on Roan 

Mountain, but did result in greater numbers of juvenile plants. 

A further consideration for managers of grass balds is that 

certain woody taxa may actually play important ecological 

roles. In particular, the disjunct Alnus viridis ssp. crispa

is a keystone species and prominent nitrogen-fixer in the 

G1-ranked alder subtype of grass bald that is unique to Roan 

Mountain (Donaldson and others, 2014; NatureServe, 2015). 

For these reasons, Donaldson and others (2014) argued that 

conservation of Alnus viridis ssp. crispa should be considered 

as a component of grass bald conservation on Roan Mountain 

rather than simply targeting it for removal along with other 

woody taxa. 

In some cases, conservation efforts for high-elevation 

outcrop and bald ecosystems have included attempts to 

reintroduce populations of rare and endemic plants. Habitat 

restoration and rare-plant reintroduction projects have been 

undertaken by the National Park Service and the U.S. Forest 

Service for several rare species endemic to high-elevation 

outcrops (Johnson, 1996). Wiser and others (1998) asserted 

that effective management and reintroduction of rare plant 

populations hinge on thorough understanding of habitat 

requirements for individual species, and they evaluated 

predictive models for the occurrence of several high-elevation 

outcrop plants. Such predictive models have special value in 

reintroduction efforts because they can potentially identify 

suitable habitat for a species that it does not currently 

occupy—presumably due to factors such as past extinction 

events or dispersal limitation—and these suitable but unoccupied sites can be targeted for reintroduction. Predictive 

habitat models may also be used to guide searches for new 

populations of rare species (Wiser and others, 1998).

Knowledge Gaps

Although much has been published about grass balds, 

research has generally been restricted to only a few individual 

balds—mostly on publicly managed land such as in the Great 

Smoky Mountains National Park—leaving many other balds 

unstudied (Crawford, 2007). Rugged and inaccessible terrain 

has restricted efforts to catalog biodiversity in high-elevation 

rock outcrops and balds, such that existing species lists may 

well be incomplete (Wiser and others, 1998; Dourson and 

Langdon, 2012). In particular, species distributions, life 

histories, and interspecific interactions are poorly understood 

for nonvascular plants and most animals inhabiting these 

high-elevation ecosystems (Wiser and White, 1999). Understanding of the factors determining habitat suitability for rare 

plants is also still limited, as is knowledge of the full ranges 

of these taxa (Wiser and others, 1998). Additionally, there 

have been few empirical investigations of the genetic effects 

of small population size and habitat fragmentation for rare 

taxa, although population genetic theory suggests these effects 

might include founder effects and genetic drift resulting in 

inbreeding depression (Godt and others, 1996).
Chapter F 81

Possible Ecological Effects From 

Climate Change

Globally, upslope range-shifting in response to recent 

warming has been documented for a number of plant and 

animal species across a range of ecosystems (Walther, 2010). 

Species with arctic or alpine affinities in high-elevation 

habitats in the southern Appalachians, including disjunct 

and regionally rare populations derived from Pleistocene 

alpine flora, are likely to be growing near their tolerance 

limits for warm temperature. For example, Cogbill and 

others (1997) estimated that the current distributions of these 

plant populations are about 1,300 to 3,300 feet (roughly 

400 to 1,000 meters) lower than would be expected based on 

their preferred temperature regimes. If this is the case, these 

populations could be especially sensitive to climate warming 

and could be vulnerable to competitive displacement by lowerelevation species migrating upslope. Warmer temperatures 

may be especially threatening to high-elevation outcrop taxa 

if higher-elevation habitats are not readily available for them 

to colonize (Wiser and White, 1999). The possible effects of 

climate change on grass balds may be mediated in large part 

by woody encroachment, indicating the need for research on 

the climate-growth relationships of the species responsible 

for tree encroachment onto balds (Crawford and Kennedy, 

2009). Asymmetrical rates of upslope range-shifting have been 

documented in other regions, such that taxa from lower elevations invade more quickly than existing taxa recede upward 

(Walther and others, 2002; Loarie and others, 2009). Interspecific differences in dispersal abilities and climate-tolerance 

thresholds might also be expected to result in changes to 

community composition in high-elevation outcrops and balds 

(Walther and others, 2002). Because grass balds are associated 

with the elevational ecotone between high-elevation hardwood 

and spruce-fir forest (Mark, 1958), an understanding of the 

dynamics of upslope ecotone migration resulting from warmer 

temperatures may have direct implications for management of 

grass balds (Crawford, 2007).

Another important mechanism by which climate change 

might affect high-elevation Appalachian communities is 

through reduced cloud immersion, a climate-change-linked 

phenomenon already causing ecosystem destruction in tropical 

mountain cloud forests as cloud-base elevations rise (Pounds 

and others, 1999; Foster, 2001). Because cloud immersion 

affects a broad array of environmental factors including water 

input, temperature, quantity and quality of solar irradiance, 

and evapotranspiration, decreased cloud immersion would 

likely have profound but also complicated effects on highelevation Appalachian communities (Johnson and Smith, 

2008). Few studies have addressed this potential threat to 

the southern Appalachians, either at the species level or the 

ecosystem level.

References

Aiken, M., and Coyle, F., 2000, Habitat distribution, life 

history and behavior of Tetragnatha spider species in the 

Great Smoky Mountains National Park: Journal of Arachnology, v. 28, no. 1, p. 97–106.

Anderson, L., 1987, A new species of Polytrichum from the 

Southern Appalachians: Bryologist, v. 90, no. 1, p. 1–5.

Barden, L., 1978, Regrowth of shrubs in grassy balds of the 

Southern Appalachians after prescribed burning: Castanea, 

v. 43, no. 4, p. 238–246.

Billings, W., and Mark, A., 1957, Factors involved in the 

persistence of montane treeless balds: Ecology, v. 38, 

no. 1, p. 140–142.

Brown, D., 1941, Vegetation of Roan Mountain—A phytosociological and successional study: Ecological Monographs, v. 11, no. 1, p. 61–97.

Brown, D., 1953, Conifer transplants to a grassy bald on 

Roan Mountain: Ecology, v. 34, no. 3, p. 614–617.

Cain, S., 1931, Ecological studies of the vegetation of the 

Great Smoky Mountains of North Carolina and Tennessee. 

I. Soil reaction and plant distribution: Botanical Gazette, 

v. 91, no. 1, p. 22–41.

Clark, G., 1968, Sorted patterned ground—New Appalachian localities south of the glacial border: Science, v. 161, 

p. 355–356.

Clarkson, R., 1960, Notes on the distribution of Alnus crispa

in Eastern North America: Castanea, v. 25, no. 2, p. 83–86.

Cogbill, C., White, P., and Wiser, S., 1997, Predicting treeline 

elevation in the southern Appalachians: Castanea, v. 62, 

no. 3, p. 137–146.

Collins, B., White, P., and Imm, D., 2001, Introduction to 

ecology and management of rare plants of the southeast: 

Natural Areas Journal, v. 21, no. 1, p. 4–11.

Conkle, L., Young, R., Bochicchio, C., and Khiel, A., 2003, 

Why go bald? Understanding the age and origin of southern 

Appalachian heath balds in Great Smoky Mountains 

National Park, in Proceedings of the Geological Society 

of America 2003 Annual Meeting: Seattle, Wash.

Crawford, C., 2007, A dendrochronological analysis of 

vegetation change and climate history in two southern 

appalachian balds, Craggy Gardens and Big Bald, N.C.: 

Blacksburg, Va., Virginia Polytechnic Institute and State 

University, Masters Thesis.
82 Insular Ecosystems of the Southeastern United States

Crawford, C., and Kennedy, L., 2009, Spatial and temporal 

patterns of tree encroachment into a southern Appalachian 

grass/heath bald using tree rings: Natural Areas Journal, 

v. 29, no. 4, p. 367–375.

DeSelm, H., and Murdock, N., 1993, Grass-dominated 

communities, in Martin, W., Boyce, S., and Echternacht, A.,

eds., Biodiversity of the southeastern United States—

Upland terrestrial communities: New York, John Wiley 

and Sons, Inc., p. 87–141.

Delcourt, H., and Delcourt, P., 1985, Quaternary palynology 

and vegetation history of the southern United States, in

Bryant, V., and Holloway, R., eds., Pollen records of lateQuaternary North American sediments: Dallas, Texas, 

American Association of Stratigraphic Paylonologists 

Foundation, p. 1–37.

Delcourt, H., and Delcourt, P., 1988, Quaternary landscape 

ecology—Relevant scales in space and time: Landscape 

Ecology, v. 2, p. 23–44.

Dey, J., 1978, Fruticose and foliose lichens of the highmountain areas of the southern Appalachians: The 

Bryologist, v. 81, no. 1, p. 1–93.

Donaldson, J., Dinkins, Z., Levy, F., and Nandi, A., 2014, 

Surface-soil properties of alder balds with respect to 

grassy and rhododendron balds on Roan Mountain, North 

Carolina-Tennessee: Southeastern Naturalist, v. 13, no. 2, 

p. 277–395.

Dourson, D., and Langdon, K., 2012, Land snails of selected 

rare high elevation forests and heath balds of the Great 

Smoky Mountains National Park: Journal of North Carolina 

Academy of Science, v. 128, no. 2, p. 27–32.

Encyclopedia of Life, 2016, Encyclopedia of Life, accessed 

February 18, 2016, at http://www.eol.org.

Fenneman, N., and Johnson, D., 1946, Physical divisions of 

the United States: U.S. Geological Survey map prepared 

in cooperation with the Physiographic Commission, 

scale 1:7,000,000. 

Fleming, G., and Patterson, K., 2012, Natural communities 

of Virginia—Ecological groups and community types: 

Richmond, Va., Virginia Department of Conservation and 

Recreation, Division of Natural Heritage.

Foster, P., 2001, The potential negative impacts of global 

climate change on tropical montane cloud forests: EarthScience Reviews, v. 55, no. 1–2, p. 73–106.

Gersmehl, P., 1973, Pseudo-timberline—The southern 

Appalachian grassy balds: Arctic and Alpine Research, 

v. 5, no. 3, p. 137–139.

Godt, M., Johnson, B., and Hamrick, J., 1996, Genetic 

diversity and population size in four rare southern Appalachian plant species: Conservation Biology, v. 10, no. 3, 

p. 796–805.

Haney, J., Lee, D., and Wilbert, M., 2001, A half-century 

comparison of breeding birds in the southern Appalachians: 

The Condor, v. 103, no. 2, p. 268–277.

Harshberger, J., 1903, An ecologic study of the flora of 

mountainous North Carolina: Botanical Gazette, v. 36, 

no. 5, p. 368–383.

Howe, T., and Bratton, S., 1976, Winter rooting activity of the 

European wild boar in the Great Smoky Mountains National 

Park: Castanea, v. 41, no. 3, p. 256–264.

Jenkins, M., 2007, Vegetation communities of Great Smoky 

Mountains National Park: Southeastern Naturalist, v. 6, 

Special Issue 1, p. 35–56.

Johnson, B., 1996, Southern Appalachian rare plant reintroductions on granite outcrops, in Falk, D., Millar, C., and 

Olwell, M., eds., Restoring diversity—Strategies for 

reintroduction of endangered plants: Washington, D.C., 

Island Press, p. 433–444.

Johnson, D., and Smith, W., 2008, Cloud immersion alters 

microclimate, photosynthesis and water relations in Rhododendron catawbiense and Abies fraseri seedlings in the 

southern Appalachian Mountains, USA: Tree physiology, 

v. 28, no. 3, p. 385–392.

Kendeigh, S., and Fawver, B., 1981, Breeding bird populations in the Great Smoky Mountains, Tennessee and North 

Carolina: The Wilson Bulletin, v. 93, no. 2, p. 218–242.

Lee, D., 2004, North Carolina’s mountain region, in Golder, 

W., and Smalling, C., eds., Important Bird Areas of North 

Carolina: Chapel Hill, N.C., Audubon North Carolina 

Chapter, p. 107–146.

Lefohn, A., Jackson, W., Shadwick, D., and Knudsen, H., 

1997, Effect of surface ozone exposures on vegetation 

grown in the southern Appalachian mountains—Identification of possible areas of concern: Atmospheric Environment, v. 31, no. 11, p. 1695–1708.

Lindsay, M., and Bratton, S., 1979a, The vegetation of grassy 

balds and other high elevation disturbed areas in the Great 

Smoky Mountains National Park: Bulletin of the Torrey 

Botanical Club, v. 106, no. 4, p. 264–275.

Lindsay, M., and Bratton, S., 1979b, Grassy balds of the Great 

Smoky Mountains—Their history and flora in relation to 

potential management: Environmental Management, v. 3, 

no. 5, p. 417–430.
Chapter F 83

Lindsay, M., and Bratton, S., 1980, The rate of woody plant 

invasion on two grassy balds: Castanea, v. 45, no. 2, 

p. 75–87.

Loarie, S., Duffy, P., Hamilton, H., Asner, G., Field, C., and 

Ackerly, D., 2009, The velocity of climate change: Nature, 

v. 462, no. 7276, p. 1052–1055.

Mark, A., 1958, The ecology of the southern Appalachian grass balds: Ecological Monographs, v. 28, no. 4, 

p. 293–336.

McConnell, J., and McConnell, O., 1983, Breeding birds of 

the Unicoi Mountains: The Chat, v. 47, p. 33– 40.

McCracken, R., Shanks, R., and Clebsch, E., 1962, Soil 

morphology and genesis at higher elevations of the Great 

Smoky Mountains: Soil Science Society of America 

Journal, v. 26, no. 4, p. 384–388.

NatureServe, 2015, NatureServe Explorer—An online 

encyclopedia of life [web application], ver. 7.1, accessed 

October 7, 2015, at http://www.natureserve.org/explorer.

Nesom, G., 2005, Broadened concept of Liatris helleri

(Asteraceae: Eupatorieae): SIDA, Contributions to Botany, 

v. 21, no. 3, p. 1323–1333.

Peine, J., and Farmer, J., 1990, Wild hog management program 

at Great Smoky Mountains National Park, in Proceedings of 

the Fourteenth Vertebrate Pest Conference: Lincoln, Nebr.

Pittillo, J., Hatcher, R., and Buol, S., 1998, Introduction to the 

environment and vegetation of the Southern Blue Ridge 

Province: Castanea, v. 63, no. 3, p. 202–216.

Pounds, J., Fogden, M., and Campbell, J., 1999, Biological 

response to climate change on a tropical mountain: Nature, 

v. 398, p. 611–615.

Powell, J., 2011, Impacts of goat browsing and disease on 

Lilium grayi, Gray’s Lily, on Roan Mountain: Johnson City, 

Tenn., East Tennessee State University, Masters Thesis.

Ramseur, G., 1960, The vascular flora of high mountain 

communities of the southern Appalachians: Journal of the 

Elisha Mitchell Scientific Society, v. 76, no. 1, p. 82–112.

Rentch, J., and Fortney, R., 1997, The vegetation of West 

Virginia grass bald communities: Castanea, v. 62, no. 3, 

p. 147–160.

Schafale, M., 2012, Guide to the natural communities of 

North Carolina, fourth approximation: Raleigh, N.C., 

North Carolina Natural Heritage Program.

Schafale, M., and Weakley, A., 1990, Classification of the 

natural communities of North Carolina, third approximation: 

Raleigh, N.C., North Carolina Natural Heritage Program.

Shafer, D., 1986, Flat Laurel Gap bog, Pisgah Ridge, North 

Carolina—Late Holocene development of a high-elevation 

heath bald: Castanea, v. 51, no. 1, p. 1–10.

Sullivan, J., and Pittillo, J., 1988, Succession of woody plants 

into a high elevation grassy bald of the Balsam Mountains: 

Castanea, v. 53, no. 4, p. 245–251.

Sullivan, J., Pittillo, J., and Keen, T., 2010, Invasion of Abies 

fraseri and Picea rubens in a southern Appalachian grassy 

bald over 30 years, in Proceedings of the 95th ESA Annual 

Meeting: Pittsburg, Pa.

Thomas, R., and Pittillo, J., 1987, Invasion of Fagus grandifolia Ehrh. into a Rhododendron catawbiense Michx. heath 

bald at Craggy Gardens, North Carolina: Castanea, v. 52, 

no. 3, p. 157–165.

Toti, D., Coyle, F., and Miller, J., 2000, A structured inventory 

of Appalachian grass bald and heath bald spider assemblages and a test of species richness estimator performance: 

Journal of Arachnology, v. 28, no. 3, p. 329–345.

Tucker, G., 1972, The vascular flora of Bluff Mountain, Ashe 

County, North Carolina: Castanea, v. 37, no. 1, p. 2–26.

U.S. Geological Survey, National Gap Analysis Program, 

2011, Land cover data portal, accessed December 2015 at 

http://gapanalysis.usgs.gov/gaplandcover/data/download/.

VanPelt, A., 1963, High altitude ants of the southern 

Blue Ridge: American Midland Naturalist, v. 69, no. 1, 

p. 205–223.

Walther, G., 2010, Community and ecosystem responses to 

recent climate change: Philosophical Transactions of the 

Royal Society, B, v. 365, p. 2019–2024.

Walther, G., Post, E., Convey, P., Menzel, A., Paremesan, C.,

Beebee, T., Fromentin, J., Hoegh-Guldberg, O., and 

Bairlein, F., 2002, Ecological responses to recent climate 

change: Nature, v. 416, p. 389–395.

Weigl, P., and Knowles, T., 1995, Megaherbivores and 

southern Appalachian grass balds: Growth and Change, 

v. 26, no. 3, p. 365–382.

Wells, B., 1936a, Andrews Bald—The problem of its origin: 

The Journal of the Southern Appalachian Botanical Club, 

v. 1, no. 5, p. 59–62.

Wells, B., 1936b, Origin of the southern Appalachian grass 

balds: Science, v. 83, no. 2151, p. 283.

Wells, B., 1961, The southern Appalachian grass bald 

problem: Castanea, v. 26, no. 2, p. 98–100.

White, P., Miller, R., and Ramseur, G., 1984, The speciesarea relationship of the southern Appalachian high 

peaks—Vascular plant richness and rare plant distributions: 

Castanea, v. 49, no. 2, p. 47–61.
84 Insular Ecosystems of the Southeastern United States

White, P., and Sutter, R., 1999, Managing biodiversity in 

historic habitats—Case history of the southern Appalachian 

grassy balds, in Peine, J., ed., Ecosystem management for 

sustainability—Principles and practices illustrated by a 

regional biosphere reserve cooperative: Washington, D.C., 

Lewis Publishers, p. 375–395.

White, P., Wilds, S., and Stratton, D., 2001, The distribution of 

heath balds in the Great Smoky Mountains, North Carolina 

and Tennessee: Journal of Vegetation Science, v. 12, no. 4, 

p. 453–466.

Whittaker, R., 1952, A study of summer foliage insect 

communities in the Great Smoky Mountains: Ecological 

Monographs, v. 22, no. 1, p. 1–44.

Whittaker, R., 1956, Vegetation of the Great Smoky Mountains: Ecological Monographs, v. 26, no. 1, p. 1–80.

Whittaker, R., 1963, Net production of heath balds and forest 

heaths in the Great Smoky Mountains: Ecology, v. 44, no. 1, 

p. 176–182.

Wiser, S., 1993, Vegetation of high-elevation rock outcrops of 

the southern Appalachians—Composition, environmental 

relationships, and biogeography of communities and rare 

species: Chapel Hill, N.C., University of North Carolina, 

Ph.D. dissertation.

Wiser, S., 1994, High-elevation cliffs and outcrops of the 

southern Appalachians: vascular plants and biogeography: 

Castanea, v. 59, no. 2, p. 85–116.

Wiser, S., 1998, Comparison of southern Appalachian highelevation outcrop plant communities with their northern 

Appalachian counterparts: Journal of Biogeography, v. 25, 

no. 3, p. 501–513.

Wiser, S., Peet, R., and White, P., 1996, High elevation rock 

outcrop vegetation of the southern Appalachian Mountains: 

Journal of Vegetation Science, v. 7, no. 5, p. 703–722.

Wiser, S., Peet, R., and White, P., 1998, Prediction of rareplant occurrence—A southern Appalachian example: 

Ecological Applications, v. 8, no. 4, p. 909–920.

Wiser, S., and White, P., 1999, High elevation outcrops 

and barrens of the southern Appalachian mountains, in

Anderson, R., Fralish, J., and Baskin, J., eds., Savannas, 

barrens, and rock outcrop plant communities of North 

America: New York, Cambridge University Press, 

p. 119–132.
Chapter G. Carolina Bays

Introduction

Carolina bays (fig. G–1) are a class of depression wetlands 

in the Coastal Plain that support a diverse suite of community 

types along hydrologic gradients from open water to xeric 

conditions (Kirkman and others, 2012). The name “Carolina 

bay” derives from the geographic concentration of this 

ecosystem in North and South Carolina and from the frequent 

presence of bay tree species (Gordonia lasianthus, Magnolia 

virginiana, Persea palustris, and P. borbonia) (Bennett and 

Nelson, 1991; Lide, 1997). Carolina bays contribute to regional 

biodiversity by providing habitat to a number of plant taxa 

of conservation concern, as well as to diverse assemblages of 

amphibian, reptile, and invertebrate taxa. 

Geographic Range 

Carolina bays are present within the Embayed and 

Sea Island sections of the Coastal Plain Physiographic 

Province (Fenneman and Johnson, 1946), hereafter referred 

to as the Atlantic Coastal Plain, from northern Florida to 

New Jersey (Prouty, 1952; Sharitz, 2003). Within North and 

South Carolina, the Inner Coastal Plain extends from the 

Fall Line (the Piedmont boundary) seaward to the Orangeburg 

Scarp (see fig. 17–2 in Soller and Mills, 1991). Seaward from 

the Orangeburg Scarp, the Middle Coastal Plain is partly 

dissected with elevations from roughly 100 to 300 feet (about 

30 to 90 meters) above sea level and extends to the Outer 

Coastal Plain, which is a gently seaward-sloping set of terraces 

Figure G–1. Aerial view of Dry Bay, Savannah River National Environmental Research Park, 

South Carolina. Photograph by Linda Lee, Savannah River Ecology Laboratory.
86 Insular Ecosystems of the Southeastern United States

generally below 100 feet (about 30 meters) of elevation with 

poorly dissected topography (Soller and Mills, 1991; Tesoriero 

and others, 2004). Carolina bays are particularly common 

within the Middle Coastal Plain and the older terraces of the 

Outer Coastal Plain, with a region of high density extending 

roughly from the Ogeechee River area in eastern Georgia to the 

Cape Fear region of North Carolina (Sharitz and Gibbons, 1982; 

Lide, 1997; Ross, 2003). Within South Carolina, the greatest 

concentration of Carolina bays is in a band approximately 

60 miles (about 100 kilometers) wide, extending northeast 

to southwest and generally coinciding with the Orangeburg 

Scarp and the Surry Scarp (Lide, 1997; Willoughby and others, 

2005). Within Maryland and Virginia, bays are restricted to the 

Delmarva Peninsula, where they are referred to as Delmarva 

bays (Olivero and Zankel, 2001), and to the Outer Coastal Plain 

(Tyndall and others, 1990). Figure G–2 shows the regions of 

highest concentrated occurrence of Carolina bays according to 

Lide (1997) and Ross (2003).

Figure G–2. Regions of concentrated occurrence of Carolina bays. 

85° 80° 75°

35°

30°

25°

Region of concentrated occurrence—

 From Lide (1997)

Region of major occurrence—

 From Ross (2003)

Region of secondary occurrence—

 From Ross (2003)

Coastal Plain Physiographic Province—

 From Fenneman and Johnson (1946)

Fall Line—From Fenneman and 

 Johnson (1946)

Orangeburg Scarp—Modified

 from Soller and Mills (1991)

Figure G-2

EXPLANATION

0 100 200 KILOMETERS

0 100 200 MILES

Base modifed from U.S. Department of Commerce, Census Bureau,

U.S. States, 2015, Mercator Auxiliary Sphere Projection

Horizontal datum is World Geodedic System of 1984

TENNESSEE

ALABAMA

GEORGIA

SOUTH

CAROLINA

NORTH

CAROLINA

VIRGINIA

FLORIDA

KENTUCKY
Chapter G 87

Physical Geography

Surficial Geology and Geomorphology

Geographic differences in soil and vegetation characteristics of Coastal Plain depression wetlands have led some 

ecologists to differentiate between “clay-based” Carolina bays 

of the Inner Coastal Plain and “peat-based” Carolina bays of the 

Middle and Outer Coastal Plain, noting that the latter commonly 

have thicker organic deposits, more ombrotrophic hydrologic 

conditions, and a greater occurrence of pocosin vegetation, pond 

cypress savanna, and cypress pond community types (Newman 

and Schalles, 1990; Bennett and Nelson, 1991; Sharitz, 2003). 

On the poorly dissected sandy surficial units of the Middle and 

Outer Coastal Plain, Carolina bays commonly have identifiable sand rims, an elliptical shape, and a northeast-southwest 

orientation, they are erosional in origin (aeolian or wave action), 

and have hydroperiods and geochemistry favorable to peat 

accumulation (Prouty, 1952; Bennett and Nelson, 1991; Lide, 

1997; Sharitz, 2003; Kirkman and others, 2012). 

Geomorphic classification of depression wetlands of the 

Inner Coastal Plain, including “clay-based” Carolina bays, is 

more problematic. Sediments in this region, extending between 

the Fall Line to the west and the Orangeburg Scarp to the 

east (with the eastern boundary somewhat ambiguous north 

of the Cape Fear River valley), are more highly dissected, 

weathered, and colluviated than those of the Outer Coastal 

Plain (Soller and Mills, 1991). Depression wetlands in this 

region tend to have mineral soils and may lack identifiable 

sand rims or a clear angled-ellipse orientation (Richardson 

and Gibbons, 1993). Some of these wetlands may be analogous 

or equivalent to the Citronelle ponds of the East Gulf Coastal 

Plain described by Folkerts (1997), with their origins linked 

to solution-subsidence of kaolinite and iron-oxide materials 

(Diane De Steven, U.S. Department of Agriculture Forest 

Service, written commun., 2014). Although clear geographic 

differences exist in the hydrologic and ecological characteristics 

of Atlantic Coastal Plain depression wetlands, the relationship 

between these differences and the underlying surficial geology 

has not been conclusively resolved. Indeed, the term “Carolina 

bay” has been widely applied to a diverse array of depression 

wetlands contexts, with no clear scientific consensus concerning 

geomorphic and geologic criteria for inclusion (Lide, 1997).

Uncertainty regarding the number of extant Carolina 

bays stems not only from disagreement as to their geomorphic 

definition but also from the size limits of detection because 

many are smaller than the mapping units of the National 

Wetlands Inventory (Sharitz, 2003). Carolina bays range 

in size from less than 2 acres (less than 1 hectare) to more 

than 1,000 acres (hundreds of hectares),11 with smaller bays 

11Published estimates of the upper limit on Carolina bay size greater than 

8,900 acres (about 3,600 hectares) require the inclusion of Lake Waccamaw, 

which is an outlier in terms of size and which may not actually be a Carolina 

bay (Diane De Steven, U.S. Department of Agriculture Forest Service, 

written commun., 2014).

outnumbering larger bays (Bennett and Nelson, 1991; Sharitz, 

2003). Semlitsch and Bodie (1998), for example, found that 

46 percent of bays at the Savannah River National Environmental Research Park near Aiken, South Carolina were about 

3 acres (1.2 hectares) or smaller and 87 percent were about 

10 acres (4 hectares) or smaller. Throughout their geographic 

range, the total number of Carolina bays in existence is 

probably between 10,000 and 20,000, although estimates have 

been as high as 500,000 (Prouty, 1952; Sharitz and Gibbons, 

1982; Sharitz, 2003).

Soil and Water Characteristics

The mineral content of Carolina bay soils or sediments is 

commonly dominated by quartz, but also includes vermiculite, 

kaolinite, feldspars, and mica (Ewing, 2003). Many Carolina 

bays have some amount of surface peat deposit, ranging from 

less than half an inch (roughly 1 centimeter) to greater than 

13 feet (about 4 meters), underlain by a black sapric muck that 

overlies a mineral base (Ewing, 2003; Sharitz, 2003). Limited 

research on cation availability in Carolina bay soils suggests 

generally low levels of exchangeable calcium, magnesium, 

and potassium (Sharitz, 2003) and low levels of total cations 

(Schalles and others, 1989). Carolina bay soils are typically 

very acidic (pH ranging from 3 to 5), with high organic carbon 

content within the surface horizons (Ewing, 2003).

Carolina bay waters are generally soft (low in calcium), 

acidic (pH ranging from 3.5 to 6), dystrophic, and low in total 

cations, with variability among Carolina bays in concentrations 

of other water-quality constituents including sulfate, dissolved 

oxygen, and manganese (Schalles and others, 1989; Mahoney 

and others, 1990; Newman and Schalles, 1990). Factors influencing water chemistry include landscape position, surficial 

geology, surrounding land use, vegetation type, disturbance 

regime, hydroperiod, and degree of hydrologic connectivity to 

groundwater and surface water (Kirkman and others, 2012). 

Water temperature and dissolved-oxygen concentrations 

may be stratified according to water depth, with stratification 

intensifying during the growing season corresponding to 

coverage by macrophytes (Schalles and others, 1989; Schalles 

and Shure, 1989).

Hydrologic Connectivity

Early assumptions of complete hydrologic isolation for 

Carolina bays—based largely on observed perched water and 

widely fluctuating water levels—have been challenged by 

recent investigations demonstrating connectivity to surfacedrainage networks and (or) to shallow groundwater (Lide, 

1995; Pyzoha and others, 2008). Some Carolina bays have 

small creeks draining into them, and some are the headwaters 

for perennial streams; however, many others are only seasonally connected to surface-drainage networks by natural 

overflow outlets (Lide, 1997). Because of the surrounding low 

relief and well-drained sandy soils, overland flow into and out 
88 Insular Ecosystems of the Southeastern United States

of many Carolina bays is typically limited to only the most 

extreme precipitation events, such as hurricanes (Sharitz, 2003; 

Barton and others, 2008). However, some Carolina bays on 

the Virginia coast and on the Delmarva Peninsula are flooded 

under high tides and so they are connected to the Chesapeake 

Bay and are not strictly freshwater wetlands (Sharitz, 2003).

The hydrologic regimes of individual bays also commonly 

involve complex and sporadic interactions with shallow 

groundwater. Hydrologic modeling indicates that certain 

Carolina bays can be involved in groundwater recharge 

and discharge, and can function as flow-through wetlands, 

depending on precipitation levels (Phillips and Shedlock, 

1993; Lide, 1995; Sun and others, 2006; Vepraskas and others, 

2007; Pyzoha and others, 2008). The degree of connectivity 

to groundwater is determined by the interaction of numerous 

site-specific factors including water-table fluctuations and 

landscape position (Sun and others, 2006; Vepraskas and 

others, 2007). For example, De Steven and Toner (2004) found 

that bays located on river terraces showed less variability in 

hydroperiod than did upland bays, consistent with the likely 

influence of groundwater for bays in close proximity to rivers. 

Subsurface topography and stratigraphy also influence the 

degree of groundwater connectivity because many Carolina 

bays are underlain by layers of humate-infused sand or by 

alternating layers of sand and clay that function as confining 

layers, slowing vertical drainage and resulting in a perched 

water table (Schalles and others, 1989; Bennett and Nelson, 

1991; Sharitz, 2003; Barton and others, 2008; Kirkman and 

others, 2012). The degree of hydrologic connectivity influences 

water chemistry and vegetation structure in Carolina bays, 

such that vegetation community types can sometimes be 

separated along a gradient from minerotrophic, groundwaterinfluenced sites to ombrotrophic, hydrologically isolated sites 

(Newman and Schalles, 1990).

Stress and Disturbance Regimes

Some Carolina bays seasonally range from inundated 

to dry, while others are generally inundated with fluctuating 

water levels, typically from about 6.5 feet (roughly 2 meters) 

above the soil surface to about 3 feet (1 meter) or more below 

it (Sharitz, 2003; Kirkman and others, 2012). Figure G–3 

shows an example of standing water in a Carolina bay. Water 

levels are generally higher in winter and early spring when 

evapotranspiration is low and precipitation is moderate, and 

gradually decline through the summer and fall, with the exception of late-summer storm events (Sharitz, 2003; Kirkman and 

others, 2012). Hydroperiod is highly variable, spatially and 

temporally, because it is influenced by regional climate and 

localized weather patterns as well as by site-specific factors 

such as the degree of hydrologic isolation, basin morphology, 

and characteristics of the surrounding landscape (Battaglia and 

Collins, 2006). Soil physical and chemical properties likely also 

play an important role. Barton and others (2008) examined a 

suite of soil properties and found that exchangeable acidity, 

total nitrogen, and total carbon were the best predictors of 

hydroperiod in unrestored Carolina bays, and noted a strong 

positive correlation (P<0.05) between clay content of the 

argillic horizon of reference bays and their hydroperiod.

Hydroperiod generally covaries with other environmental 

characteristics, including pH, dissolved oxygen, dissolved 

solids, and canopy coverage (and thus shading and water 

temperature), so the interactive effects of hydroperiod on 

biological outcomes may be direct or indirect because they 

mediate a suite of other variables (Sharitz and Gibbons, 

1982; Fairchild and others, 2003). In Carolina bay communities, hydroperiod and water-depth fluctuation are major 

determinants of vegetation composition, interacting with 

other factors such as soil disturbance, fire history, and lifehistory traits of individual species (De Steven and Toner, 

2004; Battaglia and Collins, 2006; Kirkman and others, 

2012). The exceptional species richness of some community 

types, such as pond cypress savannas in clay-based Carolina 

bays, is thought to be maintained largely by a disturbance 

regime characterized by widely fluctuating and unpredictable 

hydrologic conditions—such that plants must endure both 

prolonged inundation and drought—which exclude mesic or 

upland species that might otherwise reduce species richness 

through competitive displacement (Kirkman and Sharitz, 

1994; Sutter and Kral, 1994). 

In addition, fire during dry periods can be an important 

component of the disturbance regime, with effects on 

geochemistry and vegetation structure (Kirkman, 1995). 

Figure G–4 shows an example of burned pocosin vegetation 

at Pocosin Lakes National Wildlife Refuge in North Carolina. 

Under appropriate conditions, peat thickness gradually 

increases over time in the absence of fire but (depending 

on fire intensity) can be dramatically reduced by burning, 

a process that increases pH and releases nutrients for plant 

use (Schalles and others, 1989). Fire frequency and intensity 

can influence canopy composition in forested Carolina bays 

(Sutter and Kral, 1994) and are positively associated with 

plant-species richness in graminoid-dominated communities (Kirkman, 1995). Natural fire regimes, however, may 

be substantially altered by human activity. Even for sites 

managed by controlled burn programs, fire during drought 

periods may be restricted due to safety considerations 

(Stroh and others, 2008).
Chapter G 89

Figure G–3. A Carolina bay at Francis Marion National Forest, South Carolina. Photograph by Linda Lee, 

Savannah River Ecology Laboratory.

Figure G–4. Burned pocosin vegetation, Pocosin Lakes National Wildlife Refuge, North Carolina. Photograph 

by Linda Lee, Savannah River Ecology Laboratory.
90 Insular Ecosystems of the Southeastern United States

Community Types and Vegetation 

Dynamics

Like other nonalluvial wetlands of the Atlantic Coastal 

Plain, Carolina bays contribute to regional biodiversity by 

providing habitat diversity caused by heterogeneity among 

bays of different depression sizes, hydroperiods, and yearly 

patterns of inundation, as well as temporal and spatial 

heterogeneity of hydrologic conditions within any given bay 

(Sharitz, 2003). A variety of vegetation communities exist 

within Carolina bays (figs. G–5 through G–7), with considerable overlap of species and mosaics and (or) gradients existing 

between community types. As a result of these gradients, 

and of vegetation dynamics whereby one community may 

convert to another through directional or cyclical succession 

patterns, various investigations and reviews have suggested 

differing categorization schemes for Carolina bay communities 

(Schafale and Weakley, 1990; Tyndall and others, 1990; Bennett 

and Nelson, 1991). 

Depending on subsurface topography and stratigraphy, 

some Carolina bays can support aquatic communities. Lakes 

and small depression ponds are inundated at their centers, may 

have organic or mineral soils, and often support concentric 

zones of vegetation along hydrologic gradients from obligate 

aquatics (such as Nymphaea odorata) nearest the centers to 

emergent herbaceous vegetation (such as Leersia hexandra) 

and shrubs at the upland margins (Bennett and Nelson, 1991; 

Nifong, 1998; Sharitz, 2003; De Steven and Toner, 2004; 

Schafale, 2012). Large, natural, open-water lakes supporting 

aquatic species are relatively rare across the geographic range 

of Carolina bays (Bennett and Nelson, 1991). 

Carolina bays support a number of distinctive and 

species-rich herbaceous communities with relatively open 

canopies (Kirkman and others, 2012). Vernal pools are seasonally inundated with relatively short hydroperiods, support 

sparse-to-dense herbaceous vegetation that is intolerant of 

longer hydroperiods, and may be subject to fire when dry 

(Sharitz, 2003; Schafale, 2012). Pond-cypress savannas, an 

exceptionally floristically diverse community type associated 

with but not limited to Carolina bays, have 6- to 9-month 

hydroperiods, mineral soils with little to no peat underlain by 

a clay hardpan, and a fire frequency interval of greater than 

20 years (Sutter and Kral, 1994). The canopy of pond cypress 

savannas is dominated by Taxodium ascendens, Nyssa biflora, 

Acer rubrum, Diospyros virginiana, and (or) Persea palustris, 

Figure G–5. Seventeen Frog Pond, Scotland County, North Carolina. Photograph by Lisa Kelly, University of North Carolina 

at Pembroke.
Chapter G 91

Figure G–6. Antioch Bay, Hoke County, North Carolina. Photograph by Lisa Kelly, University of North Carolina 

at Pembroke.

Figure G–7. Hamby’s Bay, Hoke County, North Carolina. Photograph by Lisa Kelly, University of North Carolina 

at Pembroke.
92 Insular Ecosystems of the Southeastern United States

with an herbaceous understory rich in sedges, grasses, Juncus

spp., and wetland herbs (Sutter and Kral, 1994). Pond-cypress 

savannas support many species of conservation concern, 

including Rhexia aristosa, Lobelia boykinii, Litsea aestivalis, 

Stillingia aquatica, Helenium pinnatifidum, Rhynchospora 

inundata, and Oxypolis canbyi (Bennett and Nelson, 1991). 

Depression meadows—which are floristically similar to pondcypress savannas but with a greater forb component and lower 

grass and sedge richness—also support a number of species 

of concern, including Rhexia aristosa, Hypericum adpressum, 

Echinodorus tenellus, Eleocharis robbinsii, Ludwigia 

glandulosa, Coreopsis rosea, Ptilimnium nodosum, Sagittaria 

isoetiformis, and Amphicarpum species (Bennett and Nelson, 

1991; De Steven and Toner, 2004).

Forest and woodland communities are present in 

Carolina bays across a range of hydrologic regimes. Pondcypress ponds are characterized by a canopy of pond cypress 

(T. ascendens) with a mixed shrub and herbaceous understory. 

They are semipermanently flooded and thus distinguishable 

from cypress savannas, which are seasonally dry (Bennett and 

Nelson, 1991; Sharitz, 2003). Non-alluvial swamps are forest 

communities characterized by obligate wetland trees—with a 

mixture of Taxodium species, pond pine (Pinus serotina), and 

broad-leaved trees in the canopy—and are closely related to 

pond-cypress ponds but with typically shorter hydroperiods 

and greater species diversity (Bennett and Nelson, 1991; 

Sharitz, 2003; De Steven and Toner, 2004). Pond-pine 

woodlands, commonly associated with temporarily flooded 

and peat-filled Carolina bays, are characterized by a canopy 

of pond pine with a dense and tall layer of evergreen and 

semi-evergreen shrubs (Bennett and Nelson, 1991; Sharitz, 

2003; Schafale, 2012). Atlantic white-cedar forests may be 

located on hydrologically isolated sites or on stream heads, 

are intermittently or seasonally saturated, and are dominated 

by Chamaecyparis thyoides in the canopy (Schafale, 2012). In 

the absence of fire, succession in pond pine woodlands and 

Atlantic white-cedar forests in peaty Carolina bays can produce 

bay forests that are dominated by bay species with a contingent 

of pond pine (Bennett and Nelson, 1991; Schafale, 2012). 

Oak-hickory forests can form an ecotone between outer sand 

rims of Carolina bays and wet interiors (Bennett and Nelson, 

1991). When these forests are disturbed by fire or timber 

cutting, they commonly succeed to mesic “regrowth” communities dominated by sweetgum (Liquidambar styraciflua) and 

other facultative wetland species (De Steven and Toner, 2004). 

Carolina bays with sand rims commonly support xeric sandhill 

scrub communities on excessively drained and infertile sites, 

which are floristically similar to sandhill communities along 

the Fall Line—dominated by longleaf pine (Pinus palustris) 

with a scrub layer of predominantly turkey oak (Quercus 

laevis)—and are maintained in part by disturbance from fire 

(Bennett and Nelson, 1991; Schafale, 2012). 

Pocosin vegetation (fig. G–8) is characterized by very 

dense evergreen and deciduous shrub communities, often 

including Ilex and ericaceous species, with a relatively 

open or absent canopy and a limited herbaceous component 

(Bennett and Nelson, 1991; Christensen, 2000; Ross, 

2003). Drier zones with shallower peat deposits and higher 

phosphorous availability support higher-stature vegetation 

(“high pocosins”) than do wet, phosphorous-limited, thick peat 

zones that may include peat accumulations elevated above the 

surrounding landscape (“low pocosins”) (Bennett and Nelson, 

1991; Richardson, 1991; Schafale, 2012). Shrub-bog and 

pocosin vegetation typically are present in Carolina bays with 

peat and (or) sandy soils and with relatively long hydroperiods 

and ombrotrophic conditions (Kirkman and others, 2012; 

Richardson, 2012). These conditions are more common in the 

Middle and Outer Coastal Plain than in the Inner Coastal Plain 

(Bennett and Nelson, 1991; Richardson, 1991; Nifong, 1998). 

Historically, the terms “Carolina bay” and “pocosin” 

have been used interchangeably (Thom, 1970; Bennett and 

Nelson, 1991; Richardson and Gibbons, 1993). Carolina bays 

and pocosins are distinct entities, however, despite similarities 

in geographic range, low nutrient levels, and the presence 

of shrub-bog communities (Sharitz and Gibbons, 1982). 

Figure G–9, adapted from similar figures by Richardson and 

Gibbons (1993) and Ross (2003), illustrates the relationship 

between Carolina bays and pocosins. Whereas all pocosins 

are palustrine wetlands dominated by shrub vegetation or 

forest, such that pocosins may be considered a sub-class 

of shrub-bogs or of forested wetlands, Carolina bays as a 

class contain a number of other community types apart from 

pocosin vegetation, including open-water habitats with little 

vegetative cover that could be classified as lacustrine wetlands 

(Sharitz and Gibbons, 1982; Richardson and Gibbons, 1993). 

Some individual Carolina bays are entirely dominated by 

pocosin vegetation and have pocosin soil characteristics—for 

example, Jerome Bog described by Buell (1946)—whereas 

in other Carolina bays, pocosin vegetation may be limited 

or entirely absent. Pocosin communities as defined by their 

vegetation also are present in other contexts lacking the 

distinct geomorphology of Carolina bays (Richardson and 

Gibbons, 1993).

Vegetation dynamics in Carolina bays include directional 

succession as well as cyclical patterns that maintain relatively 

stable community structures (Kirkman and others, 1996; 

De Steven and Toner, 2004; Stroh and others, 2008; Kirkman 

and others, 2012). The disturbance regime components 

regulating vegetation dynamics are complex and interactive. 

Invasion of predominantly herbaceous wetland communities 

by upland woody species is facilitated by dry conditions 

(including multi-year droughts and short hydroperiods) in 

the absence of fire, and is typically impeded by fire and (or) 

flooding (Kirkman and others, 2000; Stroh and others, 2008). 

Some Carolina bays experience cyclical succession patterns, 

with periodic inundation, droughts, and fires maintaining 

sub-climax conditions over long time periods and with 

limited inorganic filling maintaining relatively constant basin 

morphology (Schalles and others, 1989). With such cyclical 

dynamics, aquatic and emergent herbaceous species are 
Chapter G 93

Figure G–8. Pocosin vegetation, Pocosin Lakes National Wildlife Refuge, North Carolina. Photograph by Linda Lee,

Savannah River Ecology Laboratory.

Figure G–9. Relationship between Carolina bays and pocosins, 

adapted from Richardson and Gibbons (1993) and Ross (2003).

favored during wet periods, and flood-intolerant upland and 

woody species may invade during droughts (Kirkman, 1995; 

Stroh and others, 2008). For Carolina bays with concentric 

zones of vegetation along xeric-to-hydric gradients, from 

exterior sand rims to inundated centers (Sharitz and Gibbons, 

1982; Schalles and others, 1989; Tyndall and others, 1990), 

zonation may shift in response to water-level changes and 

disturbance events through local extinction and recolonization 

processes (Edwards and Weakley, 2001; Sharitz, 2003). In 

such cases, a particular assemblage of species at a particular 

time represents a subset of the overall diversity possible 

(based on seed-bank diversity), rather than a seral stage of 

directional succession (Sharitz and Gibbons, 1982). Some 

Carolina bay taxa, including some of conservation concern, 

are adapted to a disturbance regime characterized by widely 

fluctuating and unpredictable hydrologic conditions via the 

maintenance of persistent seed banks from which recruitment 

can occur following disturbance events and when hydrologic 

conditions are favorable (Kirkman and Sharitz, 1994; Sutter 

and Kral, 1994; Sharitz, 2003; Battaglia and Collins, 2006). 

This type of disturbance adaptation can produce substantial 

inter-annual variation in above-ground species composition 

that is associated with hydrologic conditions, as well as seedbank diversity that is higher than above-ground diversity such 

that rare taxa may be present in the seed bank but absent in the 

vegetation (Kirkman and Sharitz, 1994; Sutter and Kral, 1994; 

Poiani and Dixon, 1995).

Palustrine wetlands

(greater than 30 percent

vegetative cover) 

Forested

pocosins 

Carolina bays

Shrub-scrub

pocosins

Lacustrine wetlands

(less than 30 percent

vegetative cover)
94 Insular Ecosystems of the Southeastern United States

In other cases, short hydroperiod or the presence of 

flood-tolerant hardwoods enables directional succession of 

herbaceous communities to forest (Kirkman, 1995; De Steven 

and Toner, 2004; Stroh and others, 2008). Hardwood establishment can accelerate the succession process by increasing 

overall transpiration, resulting in shortened hydroperiod 

(Sun and others, 2001; Stroh and others, 2008). Fire exclusion 

and other changes to historical fire regimes may also promote 

succession from herbaceous communities to mesic hardwood 

forest in relatively dry sites (De Steven and Toner, 2004). 

Pocosin vegetation may be subject to directional succession, 

but competing theories exist as to whether fire regime or 

paludification-associated nutrient limitation is the predominant successional control, with corresponding ambiguity 

as to the prevailing sequence of seral stages (Otte, 1981; 

Richardson, 2012). 

Several interacting influences likely determine the relative 

stability of community type and composition in Carolina bays. 

For example, pocosin successional patterns are likely highly 

dependent on the interactions of site-specific conditions, 

including peat depth, fire regime, hydroperiod, and species 

composition of surrounding community types (Christensen, 

2000). Experimental results from propagule banks of sediment cores suggest that hydrologic control over vegetation 

composition is highest at the wettest end of the hydrologic 

gradient—which exhibited reduced species richness with 

community composition converging on more homogenous 

assemblages of obligate wetland species—and decreases under 

drier conditions more favorable to the inclusion of facultative 

wetland and upland species (Collins and Battaglia, 2001; 

Battaglia and Collins, 2006). In one study of Carolina bay 

vegetation dynamics over the course of a multi-year drought, 

Mulhouse and others (2005) found that vegetation communities dominated by less robust, small-statured species were 

more susceptible to changes in community composition than 

those dominated by more robust clonal species. Additionally, 

land-use change and anthropogenic hydrologic alteration 

can substantially influence succession trajectories, although 

separation of these effects from those of geomorphology 

and topographic position may be impossible in some cases 

(Kirkman and others, 1996).

Contributions to Regional Biodiversity

Along with other non-alluvial wetlands in the Atlantic 

Coastal Plain, Carolina bays are important to regional 

biodiversity. Carolina bays (1) are the predominant source of 

lentic habitat for a large geographic area of the southeastern 

Coastal Plain, (2) are one of the most important natural 

breeding sites for amphibians in the region, (3) support 

extraordinarily high invertebrate richness, and (4) contain 

threatened and taxa-rich wetland communities such as pondcypress savannas (Sharitz and Gibbons, 1982; Pechmann and 

others, 1989; Schalles and others, 1989; Sutter and Kral, 1994; 

Taylor and others, 1999).

Endemic and Biogeographically Disjunct 

Plant Taxa

The exact number of rare, threatened, or endemic species in 

Carolina bays is not known, largely because several of the more 

extensive inventories and studies of rare plants, including Sutter 

and Kral (1994) and Edwards and Weakley (2001), have focused 

on broadly defined “depression wetlands” of the Coastal Plain 

rather than differentiating Carolina bays from other depression 

wetland types such as karst depressions. However, a general 

pattern emerges from floristic surveys that Carolina bays provide 

important habitat for a number of plant taxa of conservation 

concern (table 3–6 in appendix 3). Of these, Oxypolis canbyi

(fig. G–10) can be considered a near-endemic to Carolina 

bays because it is endemic to wet savannas of the Atlantic 

Coastal Plain (Bennett and Nelson, 1991). A number of other 

species, such as Peltandra sagittifolia, Helenium pinnatifidum, 

Scleria baldwinii, and Sagittaria isoetiformis, are found in 

other ecosystems throughout the Coastal Plain but have been 

found in South Carolina only in Carolina bays; thus, at the State 

level, these species might be regarded as Carolina bay endemics 

(Bennett and Nelson, 1991). Within Carolina bays, occurrence 

of rare taxa appears to be concentrated within certain community 

types, such as pond-cypress savannas and depression meadows 

(Sutter and Kral, 1994; Sharitz, 2003) and possibly within certain 

clusters of bays. In an analysis of Carolina bays on the Delmarva 

Peninsula, Olivero and Zankel (2001) found that “disjunct” bay 

complexes nearest the coast—characterized by relatively low 

density of bays and geographic separation from regions of high 

bay density—had rare-species occurrences disproportionate to 

the total number of bays.

Species-level studies, both observational and experimental, 

have yielded insight into the physiology, life histories, and 

population genetics of a number of rare plants found in 

Carolina bays, including their particular adaptations to stress 

and disturbance regimes. For example, some Carolina bay 

plants such as Sagittaria isoetiformis are adapted to widely 

fluctuating hydrology and carbon dioxide levels by having the 

ability to switch from CAM photosynthesis when submerged 

to C3 photosynthesis when emergent (Edwards and Sharitz, 

2000). Most rare species—and most plant species in general—

in Carolina bays and other Coastal Plain depression wetlands 

are perennials, a growth strategy allowing occupation of limited 

space to prevent establishment of competing seedlings and 

the development of specialized organs for storing water and 

nutrients (a drought-tolerance capacity) and for gas exchange 

(a flood-tolerance capacity) (Edwards and Weakley, 2001). For 

clonally reproducing species, a further source of stress tolerance 

may derive from physiological connections between genets 

and ramets, enabling the sharing of water and nutrients and 

mitigating the risk for ramets as they spread into new microhabitats with unpredictable hydrologic conditions (Edwards and 

Weakley, 2001). Persistent seed banks are another adaptation 

to such conditions, offering a buffer against hydrologically 

induced population crashes and the resulting risk of inbreeding 

depression (Edwards and Weakley, 2001; Sharitz, 2003).
Chapter G 95

Figure G–10. Oxypolis canbyi, Lee County, Georgia. Photograph by Alan M. Cressler, U.S. Geological Survey.

Animal Habitat

Although the widely fluctuating water levels and seasonal 

drying of most Carolina bays render them inadequate habitat 

for fish, some of the largest Carolina bays are inundated 

at their centers and support fish populations (Sharitz and 

Gibbons, 1982). In a survey of 63 Carolina bays in South 

Carolina, Snodgrass and others (1996) found that the presence of fish was inversely related to three factors: drying 

frequency, the distance to other intermittent aquatic habitats, 

and the elevation difference with the closest permanent 

aquatic habitat. The presence or absence of predatory fish in a 

Carolina bay has important ecological effects based on trophic 

interactions, especially for aquatic invertebrates (Taylor and 

others, 1999) and amphibians (Snodgrass and others, 2000a). 

In Carolina bays lacking fish, odonates and salamanders are 

typically the dominant predators (Schalles and Shure, 1989). 

Carolina bays have been documented as sites of 

extraordinarily high biomass, species diversity, and migration activity for amphibians (Sharitz and Gibbons, 1982). 

Figure G–11 shows two examples of amphibian taxa 

supported by Carolina bay habitat. The importance of Carolina 

bays to amphibian populations and to their trophic interactions 

extends beyond the high water line because amphibians that 

complete their life cycles in Carolina bays migrate annually to 

home ranges in surrounding terrestrial ecosystems (Pechmann 

and others, 1989). Hydroperiod—which is fundamentally 

linked to vegetation structure, water chemistry, and other 

important habitat characteristics—is an important factor 

regulating the numbers and species diversity of metamorphosing juvenile amphibians in Carolina bays (Pechmann 

and others, 1989). Carolina bays also provide habitat for 

several species of reptiles, such as aquatic turtles, snakes, and 

Alligator mississippiensis (American alligator) (Sharitz and 

Gibbons, 1982). Many bays serve as refugia for migratory 

waterfowl and provide forage and water for wildlife of the 

surrounding upland areas (Schalles and others, 1989).

During periods of drought when many smaller Carolina 

bays become entirely dry, larger and semipermanently inundated Carolina bays can serve as refugia for mobile species 

such as aquatic reptiles (Sharitz and Gibbons, 1982). Even so, 

reproductive output may be substantially lower during drought 

years depending on mobility and habitat requirements of 

individual species (Gibbons and others, 1983). Pechmann and 

others (1991) suggested that drought was largely responsible 

for recruitment failures for several amphibian species they 

studied in one Carolina bay for more than a decade, and they 

hypothesized that adult and larval mortality associated with 

breeding during dry years may have selected for amphibians 

capable of responding to precipitation as a predictive cue and 
96 Insular Ecosystems of the Southeastern United States

Figure G–11. Amphibian taxa supported by Carolina bay habitat: A, Ambystoma maculatum (spotted salamander) 

in a Carolina bay on the eastern shore of Maryland; B, Lithobates sylvaticus (wood frog) in a Carolina bay on the 

eastern shore of Maryland. Photographs by Joel Snodgrass, Virginia Polytechnic Institute and State University.

A

B
Chapter G 97

reproducing only during wet years. Observing the movement 

of Coluber constrictor (subspecies of racers) surrounding a 

Carolina bay in South Carolina, Plummer and Congdon (1994) 

observed unexpectedly large home ranges and hypothesized 

that recent drought may have reduced prey availability and 

prompted the snakes to hunt over wider areas.

Aquatic invertebrates in Carolina bays also must cope 

with widely fluctuating water levels and variable hydroperiods. While some invertebrates must regularly recolonize 

Carolina bays following dry periods, others rely on aestivation 

and other mechanisms of desiccation resistance to survive 

dry conditions (Taylor and Mahoney, 1990; Taylor and 

others, 1999; Dietz-Brantley and others, 2002). Comparison 

of assemblages of aquatic and semiaquatic insects between 

Carolina bays and temporary ponds outside the Coastal Plain 

suggests that Carolina bays may support unusually high levels 

of insect species richness, especially for chironomids (Leeper 

and Taylor, 1998). Individual bays commonly support more 

than 100 species, with some cases of more than 300 species 

collected from a single bay (Taylor and others, 1999). Carolina 

bays also appear to have an extraordinarily high diversity of 

crustaceans compared to other temporary ponds, including 

cladocerans, other branchiopods, and calanoid copepods 

(Mahoney and others, 1990; Taylor and others, 1999). At least 

one copepod species, newly described by DeBiase and Taylor 

(1997), is believed to be endemic to Carolina bays. Most 

crustaceans are small—less than 0.15 inch (about 4 millimeters) in length—and are considered permanent residents, 

surviving dry periods through dormancy or by the production 

of resting eggs (Taylor and others, 1999). Other ubiquitous 

aquatic invertebrates in Carolina bays include annelids, 

nematodes, and rotifers (Taylor and others, 1999).

Invertebrate community structures in Carolina bays are 

regulated by a number of biotic and abiotic factors. Insect 

assemblages are often dominated by species adapted to the 

geochemistry of their habitat, including low pH, low dissolvedoxygen concentrations, and widely fluctuating hydrologic 

conditions (Leeper and Taylor, 1998). Relative scarcity of 

mollusks in Carolina bays is believed to result from calcium 

limitation (Schalles and Shure, 1989). The composition of 

aquatic invertebrate communities in Carolina bays is highly 

dynamic in response to fluctuating hydrology—as water levels 

rise, some species emerge from their resting stages sooner 

than others (Taylor and Mahoney, 1990). Aquatic invertebrate 

populations also show strong association with vegetation 

types, such that particular plant communities tend to support 

particular assemblages of aquatic invertebrates (Schalles and 

Shure, 1989).

Conservation Considerations

As described in this section, Carolina bays have been 

subjected to anthropogenic degradation of various kinds for 

decades. Kirkman and others (2012) asserted that Carolina 

bays and other southeastern depression wetlands have been 

“greatly undervalued, particularly in regard to the important 

habitats they provide.” Increasingly, the importance of these 

ecosystems has been recognized, prompting efforts to restore 

their hydrologic and ecological function (Barton and others, 

2008; De Steven and others, 2010).

Threats to Ecosystem Integrity

Although no consensus exists on current or historical 

numbers of Carolina bays, the general agreement is that 

many, if not most, have been heavily damaged, contributing to 

overall loss of palustrine wetlands in the Coastal Plain (Dahl, 

1999). Large-scale surveys of Carolina bays have concluded 

that previous estimates of the total number of bays were too 

high and that most of the existing bays had been dramatically altered by human activity (South Carolina Department 

of Natural Resources, 1999). Based on a survey of several 

thousand Carolina bays in South Carolina in the 1980s, 

Bennett and Nelson (1991) determined that 97 percent of bays 

showed some evidence of human alteration (including ditches, 

roads, farm ponds, development, logging, tree plantations, and 

row cropping), with half of all bays having at least 20 percent of 

their area damaged. They found this destruction to be especially pronounced for Carolina bays in the Inner Coastal Plain.

Many human activities have produced visible destruction and degradation of Carolina bays. Ditching has been 

a widespread and severe form of alteration, with many 

Carolina bays having been drained to enable conversion to 

agriculture (Sharitz and Gibbons, 1982; Dahl, 1999). Sand 

rims and associated xeric sandhill scrub communities have 

also been damaged by road construction, pine logging, and 

sand quarrying (Bennett and Nelson, 1991). Agricultural and 

forestry practices in upland areas surrounding Carolina bays 

can also degrade bay vegetation composition and ecosystem 

integrity. For example, Poiani and Dixon (1995) demonstrated 

that proximity to clear cuts was positively associated with 

the proportion of weedy species present in Carolina bay 

seed banks. 

Other less readily visible processes threaten the ecological integrity of Carolina bays, such as the bioaccumulation 

of mercury. Despite the lack of any obvious anthropogenic 

or geologic sources of mercury, several characteristics of 

Carolina bays may exacerbate mercury methylation and 

bioaccumulation, including fluctuating water levels, high 

dissolved organic carbon content, and low pH (Brant and 

others, 2002; Unrine and others, 2005). Unrine and others 

(2005) observed bioaccumulated mercury concentrations in 

amphibian larvae within the range at which adverse effects 

(development problems and metamorphic failure) had been 

observed in laboratory studies. Evidence also exists for dietary 

mercury exposure from Carolina bay fish as a potential threat 

to sensitive piscivorous birds, including the wood stork 

(Mycteria americana), a federally listed species (Snodgrass 

and others, 2000b; Brant and others, 2002). 
98 Insular Ecosystems of the Southeastern United States

Conservation Strategies

By the early 1980s, conservation officials had begun to 

recognize the need to inventory, study, monitor, and protect 

Carolina bays based on the paucity of scientific information 

on these ecosystems combined with mounting evidence 

of their ongoing destruction (Sharitz and Gibbons, 1982; 

Bennett and Nelson, 1991). Protected lands now sheltering 

Carolina bays, owned either by state parks or by the Nature 

Conservancy, include Woods Bay State Park and Cathedral 

Bay Heritage Preserve in South Carolina; Bladen Lake State 

Forest, Bushy Lake State Natural Area, Jones Lake State Park, 

and the McIntire Bay complex in North Carolina; Millington 

Wildlife Management Area in Maryland (Sharitz, 2003); and 

the Blackbird State Forest in northern Delaware (Fairchild 

and others, 2003). One of the most highly studied clusters of 

Carolina bays is on the Savannah River Site, a nuclear reservation owned and managed by the U.S. Department of Energy, 

which is located within a region of high Carolina bay density 

and which contains approximately 200 intact bays (De Steven 

and Toner, 2004). 

Although most of the well-studied sites are on protected 

land, it is likely that the majority of Carolina bays throughout 

the range of the ecosystem are not. On the Delmarva Peninsula, for example, less than 14 percent of all Carolina bays and 

less than 30 percent of associated rare biodiversity elements are 

on protected land (Olivero and Zankel, 2001). Recent changes 

in U.S. Federal and State regulatory protections for “isolated” 

depression wetlands have heightened threats of future anthropogenic destruction or degradation of Carolina bays (Tiner, 

2003a, b). Additionally, local and regional decisionmaking for 

Carolina bay conservation may contradict scientific evidence. 

In general, large and semipermanently inundated Carolina 

bays are more likely to be protected than smaller “dry-end” 

Carolina bays, with the assumptions that small wetlands 

support fewer species and that these species are subsets of the 

larger set of species found in larger wetlands, despite empirical 

evidence on amphibian communities contradicting both 

assumptions (Snodgrass and others, 2000c; Sharitz, 2003). 

Indeed, highly variable species diversity and community 

structure between Carolina bays suggests that conservation 

strategies predicated on protecting only a limited number 

of “representative” bays will fail to conserve the full suite of 

biodiversity (Taylor and others, 1999).

Many projects have attempted to restore individual 

Carolina bays or clusters of bays to their previous ecological 

states. Hydrologic restoration efforts often include the 

disruption of artificial drainage mechanisms (such as by 

plugging ditches), allowing examinations of the resulting 

hydroperiod changes and the associated effects on plant and 

animal communities (Dietz and others, 2001; De Steven and 

others, 2006, 2010). In addition to hydrologic restoration, 

conservation efforts may require prescribed burning or other 

forms of soil disturbance, as well as the control of dominant 

competitors (Kirkman and Sharitz, 1994). Because Carolina 

bays naturally support gradients and matrices of community 

types determined by many interactive abiotic and biotic 

factors, restoration projects must take into account sitespecific information such as basin size, soil type, thickness 

of organic soil layers, nutrient (especially phosphorous) 

levels, disturbance history, topographic position on the 

landscape, surrounding land use, hydroperiod, and ponding 

depth (De Steven and Toner, 2004; Vepraskas and others, 

2007; De Steven and others, 2010). “Passive” restoration 

approaches—relying on seed-bank recruitment and natural 

dispersal following hydrologic restoration and the removal of 

upland forest vegetation—may offer low-cost alternatives to 

active seeding or planting, but they may not fully establish the 

characteristic assemblages found in reference bays (De Steven 

and others, 2006), requiring supplemental transplantation of 

selected species (De Steven and Sharitz, 2007).

Evaluation of restoration projects in Carolina bays typically requires long-term monitoring. If hydroperiod change 

or hydroperiod comparison of restored bays to reference 

bays is used as an evaluation standard, then long periods of 

pre-restoration and post-restoration data are required because 

of the high degree of spatial and temporal variability in hydroperiod, which is common to Carolina bays (Barton and others, 

2008). Hydrologic restoration can be hampered by regional 

droughts (De Steven and others, 2006; 2010) and by regional 

hydrologic alteration (Vepraskas and others, 2007). Even 

when hydrologic restoration is apparently successful based on 

hydroperiod similarity between restored and reference bays, 

ecological restoration can be complicated by lingering effects 

of previous agricultural use, such as altered soil physical 

properties as well as high pH and nutrient levels from lime and 

fertilizer applications (Bruland and others, 2003; Vepraskas 

and others, 2007). Other factors such as seed-bank persistence 

and variable germination ability necessitate long-term monitoring of rare plant populations (Edwards and Weakley, 2001). 

Knowledge Gaps

Although Carolina bays are a relatively well-studied 

ecosystem, a number of knowledge gaps persist. For example, 

additional research is warranted on floristic diversity of 

Carolina bays. Persistent seed banks and corresponding high 

levels of inter-annual variation in community composition 

render one-time floristic inventories insufficient to capture the 

true diversity of Carolina bay plant communities, necessitating 

multi-year studies to compare seed-bank to above-ground 

vegetation diversity (Sutter and Kral, 1994). Although 

seed-bank dynamics may be critical to the survival of 

certain rare plant populations (Sharitz, 2003), little is known 

about the seed banks for many rare species. For example, 

only about 20 percent of the depression-wetland rare-plant 

species listed by Edwards and Weakley (2001) had seed-bank 

information available. 

Microbial autotrophs, including algae and purple 

photosynthetic bacteria, constitute an important but littlestudied component of Carolina bay communities. Purple 
Chapter G 99

photosynthetic bacteria may play a role in nitrogen fixation 

and in acidification resulting from sulfuric acid production 

(Schalles and Shure, 1989). Based on the construction of 

16S rRNA gene libraries from sediments of one Carolina 

bay, Wise and others (1997) found a high level of microbial 

diversity with numerous sequences unrepresented in existing 

databases, suggesting a strong prevalence of novel, undescribed microbial species. The microbial ecology of Carolina 

bays deserves further study, including trophic relationships 

and food-web dynamics.

Additionally, the drivers and ecological effects of hydroperiod dynamics need further empirical study. The interaction 

between Carolina bay surface water and groundwater, only 

recently accepted as an important hydrologic feature of these 

ecosystems, is still not fully understood (Schalles and Shure, 

1989; Lide, 1995; Pyzoha and others, 2008). Soil physical 

and chemical properties appear to be related to hydroperiod 

response in Carolina bays (for example, argillic horizon clay 

content is positively associated with hydroperiod); however, 

the exact nature of these relationships has not been quantified 

at large scales (Barton and others, 2008). Although hydroperiod is a known factor determining vegetation expression from 

seed banks, the strength of this relationship is not known, 

making it difficult to predict the magnitude and type of 

community restructuring resulting from hydrologic changes 

(Battaglia and Collins, 2006). Understanding this relationship 

will be critical for future conservation efforts in the face of 

changing hydrologic conditions, including those that might 

result from regional climate change. Other important conservation needs include the quantification of upland land area 

surrounding Carolina bays necessary to maintain water quality, 

hydrologic integrity, and habitat for migrating animal species 

(Sutter and Kral, 1994), and the determination of appropriate 

evaluation metrics for Carolina bay restoration projects 

(De Steven and others, 2010).

Possible Ecological Effects From 

Climate Change

Although several studies have alluded to climate change 

as a possible factor affecting Carolina bays, systematic 

analyses are lacking in the scientific literature regarding the 

potential current and future effects from climate change on 

Carolina bay hydrology and ecological outcomes. In a recent 

review of the literature on southeastern depression wetlands, 

Kirkman and others (2012) mentioned briefly that regional 

changes in precipitation timing and magnitude would likely 

affect hydroperiod and, in turn, ecological outcomes, but did 

not elaborate further. Mulhouse and others (2005) and Stroh 

and others (2008) assessed vegetation change in response 

to short-term climatic variability (multi-year droughts and 

drought/re-flooding cycles) but not long-term directional 

climate change. Similarly, Sun and others (2006) used a 

modeling approach based on short-term climate variability 

(a wet period and multi-year drought) to predict the direction 

of groundwater/surface-water interactions resulting from 

precipitation inputs and noted that climate change would 

likely affect hydroperiod through such interactions, but 

did not offer specific long-term projections based on 

climate-change scenarios.

In the absence of systematic, empirical research on the 

ecological effects of long-term directional climate change, 

hypotheses can be formulated based on current knowledge of 

the factors governing vegetation structure and faunal habitat 

in Carolina bays. Given the well-documented importance of 

hydroperiod to community structure for plants (Sharitz and 

Gibbons, 1982; Sutter and Kral, 1994; Sun and others, 2001; 

De Steven and Toner, 2004; Battaglia and Collins, 2006; Stroh 

and others, 2008; Kirkman and others, 2012) and for animals 

and their trophic interactions (Pechmann and others, 1989; 

Snodgrass and others, 1996; Fairchild and others, 2003), 

hydroperiod alteration is likely to be a key mechanism by 

which climate change induces ecological shifts in Carolina 

bays. Paleoecological investigations may offer some evidence 

of how previous climatic shifts have been associated with 

hydrologic and vegetation changes (Whitehead, 1964). For 

example, based on micropetrographic analysis of peat cores 

from a Carolina bay showing a large number of thin peat 

layers reflecting alternating wet and dry periods, Cohen and 

others (1999) inferred that biotic responses in that wetland had 

been highly sensitive to hydrologic and climatic changes over 

time. Comparing these results to those from other peatlands 

in the southeastern United States with larger areal extents, 

such as the Everglades in Florida and Okefenokee Swamp in 

Georgia, Cohen and others (1999) inferred that the small size 

of Carolina bay wetlands might make them especially sensitive to vegetation shifts in response to climatic fluctuations.

Over the long term, if hydroperiods shorten on average 

or if droughts become more common or severe, ecological 

consequences might resemble the effects of artificial drainage, 

possibly including directional succession of rare and speciesrich herbaceous wetland communities to forests. Drought 

intensification would also be expected to degrade habitat for 

aquatic reptiles (Gibbons and others, 1983) and amphibians 

(Pechmann and others, 1991). Conversely, climate-changeinduced reductions in drought frequency or intensity might 

adversely affect Carolina bay plant taxa whose recruitment 

from the seed bank is dependent on drought-related 

disturbance (Kirkman and Sharitz, 1994). Empirical evidence 

suggests that aquatic habitat in Carolina bays supports reduced 

plant-species richness compared to zones of frequent fluctuation in water level (Collins and Battaglia, 2001). If so, then 

hydroperiod lengthening as a result of climate change could 

conceivably reduce plant-species richness if zones currently 

subjected to frequent drying shift toward semipermanent 

inundation. At this point in time, such hypotheses are largely 

speculative. To test such hypotheses about climate-changedriven ecological outcomes, empirical studies are needed to 

quantify the effects of precipitation and temperature changes 

in terms of hydroperiod alteration and to link this hydroperiod 

alteration to long-term change in community structures.
100 Insular Ecosystems of the Southeastern United States

References

Barton, C., Andrews, D., and Kolka, R., 2008, Evaluating 

hydroperiod response in restored Carolina bay wetlands 

using soil physicochemical properties: Restoration Ecology, 

v. 16, no. 4, p. 668–677.

Battaglia, L., and Collins, B., 2006, Linking hydroperiod 

and vegetation response in Carolina bay wetlands: Plant 

Ecology, v. 184, no. 1, p. 173–185.

Bennett, S., and Nelson, J., 1991, Distribution and status of 

Carolina bays in South Carolina: Columbia, S.C., South 

Carolina Wildlife and Marine Resources Department, 

Nongame and Heritage Trust Publications.

Brant, H., Jagoe, C., Snodgrass, J., Bryan, A., and Gariboldi, J.,

2002, Potential risk to wood storks (Mycteria americana) 

from mercury in Carolina bay fish: Environmental Pollution, v. 120, p. 405–413.

Bruland, G., Hanchey, M., and Richardson, C., 2003, Effects 

of agriculture and wetland restoration on hydrology, soils, 

and water quality of a Carolina bay complex: Wetlands 

Ecology and Management, v. 11, p. 141–156.

Buell, M., 1946, Jerome Bog, a peat-filled “Carolina bay”: 

Bulletin of the Torrey Botanical Club, v. 73, no. 1, p. 24–33.

Christensen, N., 2000, Vegetation of the southeastern Coastal 

Plain, in Barbour, M., and Billings, W., eds., North American terrestrial vegetation: New York, N.Y., Cambridge 

University Press, p. 397– 448.

Cohen, A., Gage, C., Moore, W., and VanPelt, R., 1999, 

Combining organic petrography and palynology to assess 

anthropogenic impacts on peatlands, part 2—An example 

from a Carolina bay wetland at the Savannah River Site 

in South Carolina: International Journal of Coal Geology, 

v. 39, p. 47–95.

Collins, B., and Battaglia, L., 2001, Hydrology effects on 

propagule bank expression and vegetation in six Carolina 

bays: Community Ecology, v. 2, no. 1, p. 21–33.

Dahl, T., 1999, South Carolina’s wetlands status and trends, 

1982–1989: Washington, D.C., U.S. Department of the 

Interior, Fish and Wildlife Service. 

DeBiase, A., and Taylor, B., 1997, Aglaodiaptomus atomicus, 

a new species (Crustacea: Copepoda: Calanoida: Diaptomidae) from freshwater wetland ponds in South Carolina, 

USA, and a redescription of A. saskatchewanensis (Wilson 

1958): Proceedings of the Biological Society of Washington, v. 110, no. 4, p. 569–580.

De Steven, D., and Sharitz, R., 2007, Transplanting native 

dominant plants to facilitate community development in 

restored Coastal Plain wetlands: Wetlands, v. 27, no. 4, 

p. 972–978.

De Steven, D., Sharitz, R., and Barton, C., 2010, Ecological 

outcomes and evaluation of success in passively restored 

southeastern depressional wetlands: Wetlands, v. 30, no. 6, 

p. 1129–1140.

De Steven, D., Sharitz, R., Singer, J., and Barton, C., 2006, 

Testing a passive revegetation approach for restoring coastal 

plain depression wetlands: Restoration Ecology, v. 14, no. 3, 

p. 452–460.

De Steven, D., and Toner, M., 2004, Vegetation of upper 

coastal plain depression wetlands—Environmental 

templates and wetland dynamics within a landscape 

framework: Wetlands, v. 24, no. 1, p. 23–42.

Dietz, S., Batzer, D., Taylor, B., and DeBiase, A., 2001, 

Invertebrate communities of twenty ditched Carolina bay 

wetlands scheduled for restoration, in Proceedings of the 

2001 Georgia Water Resources Conference: Athens, Ga., 

p. 321–324.

Dietz-Brantley, S., Taylor, B., Batzer, D., and DeBiase, A., 

2002, Invertebrates that aestivate in dry basins of Carolina 

bay wetlands: Wetlands, v. 22, no. 4, p. 767–775.

Edwards, A., and Sharitz, R., 2000, Population genetics of two 

rare perennials in isolated wetlands: Sagittaria isoetiformis

and S. teres (Alismataceae): American Journal of Botany, 

v. 87, no. 8, p. 1147–1158.

Edwards, A., and Weakley, A., 2001, Population biology and 

management of rare plants in depression wetlands of the 

southeastern coastal plain, USA: Natural Areas Journal, 

v. 21, no. 1, p. 12–35.

Ewing, J., 2003, Characterization of soils in a drained 

Carolina Bay wetland prior to restoration: Raleigh, N.C., 

North Carolina State University, Masters Thesis. 

Fairchild, G., Cruz, J., Faulds, A., Short, A., and Matta, J., 

2003, Microhabitat and landscape influences on aquatic 

beetle assemblages in a cluster of temporary and permanent ponds: Journal of the North American Benthological 

Society, v. 22, no. 2, p. 224–240.

Fenneman, N., and Johnson, D., 1946, Physical divisions of 

the United States: U.S. Geological Survey map prepared 

in cooperation with the Physiographic Commission, 

scale 1:7,000,000. 

Folkerts, G., 1997, Citronelle ponds—Little-known wetlands 

of the central Gulf Coastal Plain, USA: Natural Areas 

Journal, v. 17, p. 6–16.
Chapter G 101

Gibbons, J., Greene, J., and Congdon, J., 1983, Droughtrelated responses of aquatic turtle populations: Journal of 

Herpetology, v. 17, no. 3, p. 242–246.

Kirkman, L., 1995, Impacts of fire and hydrological regimes 

on vegetation in depression wetlands of southeastern USA, 

in Cerulean, S., and Engstrom, R., eds., Fire in wetlands—

A management perspective, Proceedings of the Tall Timbers 

Fire Ecology Conference, No. 19: Tall Timbers Research 

Station, Tallahassee, Fla., p. 10–20.

Kirkman, L.K., Goebel, P.C., West, L., Drew, M.B., and 

Palik, B.J., 2000, Depressional wetland vegetation types—

A question of plant community development: Wetlands, 

v. 20, no. 2, p. 373–385.

Kirkman, L., Lide, R., Wein, G., and Sharitz, R., 1996, 

Vegetation changes and land-use legacies of depression 

wetlands of the western Coastal Plain of South Carolina—

1951–1992: Wetlands, v. 16, no. 4, p. 564–576.

Kirkman, L., and Sharitz, R., 1994, Vegetation disturbance and 

maintenance of diversity in intermittently flooded Carolina 

bays in South Carolina: Ecological Applications, v. 4, no. 1, 

p. 177–188.

Kirkman, L., Smith, L., and Golladay, S., 2012, Southeastern 

depressional wetlands, in Batzer, D., and Baldwinl, A., eds., 

Wetland habitats of North America—Ecology and conservation concerns: Berkeley, Calif., University of California 

Press, p. 203–215.

Leeper, D., and Taylor, B., 1998, Insect emergence from 

a South Carolina (USA) temporary wetland pond, with 

emphasis on the Chironomidae (Diptera): Journal of 

the North American Benthological Society, v. 17, no. 1, 

p. 54–72.

Lide, R., 1995, Hydrology of a Carolina bay located on the 

upper Coastal Plain of western South Carolina: Wetlands, 

v. 15, no. 1, p. 47–57.

Lide, R., 1997, When is a depression wetland a Carolina bay?: 

Southeastern Geographer, v. 37, p. 90–98.

Mahoney, D., Mort, M., and Taylor, B., 1990, Species richness 

of calanoid copepods, cladocerans and other branchiopods 

in Carolina bay temporary ponds: American Midland Naturalist, v. 123, no. 2, p. 244–258.

Mulhouse, J., De Steven, D., Lide, R., and Sharitz, R., 

2005, Effects of dominant species on vegetation change 

in Carolina bay wetlands following a multi-year drought: 

The Journal of the Torrey Botanical Society, v. 132, no. 3, 

p. 411–420.

Newman, M., and Schalles, J., 1990, The water chemistry of 

Carolina bays—A regional survey: Stuttgart, Archiv für 

Hydrobiologie, v. 118, no. 2, p. 147–168.

Nifong, T., 1998, An ecosystematic analysis of Carolina bays 

in the coastal plain of the Carolinas: Chapel Hill, N.C., 

University of North Carolina, Ph.D. dissertation.

Olivero, A., and Zankel, M., 2001, Delmarva bay density 

analysis: The Nature Conservancy, accessed April 8, 2015, 

at http://gis.tnc.org/data/MapbookWebsite/map_page.

php?map_id=11.

Otte, L., 1981, Origin, development, and maintenance of 

pocosin wetlands of North Carolina. Report to the North 

Carolina Natural Heritage Program: Raleigh, N.C., North 

Carolina Department of Natural Resources and Community 

Development.

Pechmann, J., Scott, D., Gibbons, J., and Semlitsch, R., 1989, 

Influence of wetland hydroperiod on diversity and abundance of metamorphosing juvenile amphibians: Wetlands 

Ecology and Management, v. 1, no. 1, p. 3 –11.

Pechmann, J., Scott, D., Semlitsch, R., Caldwell, J., Vitt, L., 

and Gibbons, J., 1991, Declining amphibian populations—

The problem of separating human impacts from natural 

fluctuations: Science, v. 253, p. 892–895.

Phillips, P., and Shedlock, R., 1993, Hydrology and chemistry 

of groundwater and seasonal ponds in the Atlantic Coastal 

Plain in Delaware, USA: Journal of Hydrology, v. 141, 

p. 157–178.

Plummer, M., and Congdon, J., 1994, Radiotelemetric study 

of activity and movements of racers (Coluber constrictor) 

associated with a Carolina bay in South Carolina: Copeia, 

v. 1994, no. 1, p. 20–26.

Poiani, K., and Dixon, P., 1995, Seed banks of Carolina 

bays—Potential contributions from surrounding landscape 

vegetation: American Midland Naturalist, v. 134, no. 1, 

p. 140–154.

Prouty, W., 1952, Carolina bays and their origin: Geological 

Society of America Bulletin, v. 63, p. 167–224.

Pyzoha, J., Callahan, T., Sun, G., Trettin, C., and Miwa, M., 

2008, A conceptual hydrologic model for a forested 

Carolina bay depressional wetland on the Coastal Plain 

of South Carolina, USA: Hydrological Processes, v. 22, 

p. 2689–2698.

Richardson, C., 1991, Pocosins—An ecological perspective: 

Wetlands, v. 11, no. 1, p. 335–354.

Richardson, C., 2012, Pocosins—Evergreen shrub bogs of 

the southeast, in Batzer, D., and Baldwin, A., eds., Wetland 

habitats of North America—Ecology and conservation 

concerns: Berkeley, Calif., University of California Press, 

p. 189–202.
102 Insular Ecosystems of the Southeastern United States

Richardson, C., and Gibbons, J., 1993, Pocosins, Carolina 

bays, and mountain bogs, in Martin, W., Boyce, S., and 

Echternacht, A., eds., Biodiversity of the southeastern 

United States—Lowland terrestrial communities: New 

York, John Wiley and Sons, Inc., p. 257–310.

Ross, T.E., 2003, Pocosins and Carolina bays compared: 

The North Carolina Geographer, v. 11, p. 22–32.

Schafale, M., 2012, Guide to the natural communities of 

North Carolina, fourth approximation: Raleigh, N.C., 

North Carolina Natural Heritage Program. 

Schafale, M., and Weakley, A., 1990, Classification of the 

natural communities of North Carolina, third approximation: Raleigh, N.C., North Carolina Natural Heritage 

Program.

Schalles, J., Sharitz, R., Gibbons, J., Leversee, G., and 

Knox, J., 1989, Carolina bays of the Savannah River plant: 

Aiken, S.C., Savannah River Plant National Environmental 

Research Park Program. 

Schalles, J., and Shure, D., 1989, Hydrology, community 

structure, and productivity patterns of a dystrophic Carolina bay wetland: Ecological Monographs, v. 59, no. 4, 

p. 365–385.

Semlitsch, R., and Bodie, J., 1998, Are small, isolated 

wetlands expendable?: Conservation Biology, v. 12, 

no. 5, p. 1129–1133.

Sharitz, R., 2003, Carolina bay wetlands—Unique habitats 

of the southeastern United States: Wetlands, v. 23, no. 3, 

p. 550–562.

Sharitz, R., and Gibbons, J., 1982, The ecology of southeastern shrub bogs (pocosins) and Carolina bays—A 

community profile, FWS/OBS-82/04: Washington, D.C., 

Department of Interior, U.S. Fish and Wildlife Service, 

Division of Biological Services.

Snodgrass, J., Bryan, A., and Burger, J., 2000a, Development 

of expectations of larval amphibian assemblage structure in 

southeastern depression wetlands: Ecological Applications, 

v. 10, no. 4, p. 1219–1229.

Snodgrass, J., Bryan, A., Lide, R., and Smith, G., 1996, 

Factors affecting the occurrence and structure of fish assemblages in isolated wetlands of the upper Coastal Plain, USA: 

Canadian Journal of Fisheries and Aquatic Sciences, v. 454, 

p. 443–454.

Snodgrass, J., Jagoe, C., Bryan, A., Brant, H., and Burger, J., 

2000b, Effects of trophic status and wetland morphology, 

hydroperiod, and water chemistry on mercury concentrations in fish: Canadian Journal of Fisheries and Aquatic 

Sciences, v. 57, no. 1, p. 171–180.

Snodgrass, J., Komoroski, M., Bryan, A., and Burger, J., 

2000c, Relationships among isolated wetland size, hydroperiod, and amphibian species richness—Implications for 

wetland regulations: Conservation Biology, v. 14, no. 2, 

p. 414–419.

Soller, D., and Mills, H., 1991, Surficial geology and geomorphology, in Horton, J., and Zullo, V., eds., The geology of 

the Carolinas: Knoxville, Tenn., University of Tennessee 

Press, p. 290–308.

South Carolina Department of Natural Resources, 1999, 

Advance identification of Carolina bays for South Carolina wetlands protection: Columbia, S.C., South Carolina 

Department of Natural Resources, Division of Land, Water 

and Conservation.

Stroh, C., De Steven, D., and Guntenspergen, G., 2008, Effect 

of climate fluctuations on long-term vegetation dynamics in 

Carolina bay wetlands: Wetlands, v. 28, no. 1, p. 17–27.

Sun, G., Callahan, T., Pyzoha, J., and Trettin, C., 2006, 

Modeling the climatic and subsurface stratigraphy controls 

on the hydrology of a Carolina bay wetland in South Carolina, USA: Wetlands, v. 26, no. 2, p. 567–580.

Sun, G., McNulty, S., Shepard, J., Amatya, D., Riekerk, H., 

Comerford, N., Skaggs, W., and Swift, L., 2001, Effects of 

timber management on the hydrology of wetland forests in 

the southern United States: Forest Ecology and Management, v. 143, no. 1–3, p. 227–236.

Sutter, R., and Kral, R., 1994, The ecology, status, and conservation of two non-alluvial wetland communities in the south 

Atlantic and eastern Gulf coastal plain, USA: Biological 

Conservation, v. 68, p. 235–243.

Taylor, B., Leeper, D., McClure, M., and DeBiase, A., 1999, 

Carolina bays—Ecology of aquatic invertebrates and 

perspectives on conservation, in Batzer, D., Rader, R., and 

Wissinger, S., eds., Invertebrates in freshwater wetlands 

of North America: New York, John Wiley and Sons, Inc., 

p. 167–196.

Taylor, B., and Mahoney, D., 1990, Zooplankton in Rainbow 

Bay, a Carolina bay pond—Population dynamics in a 

temporary habitat: Freshwater Biology, v. 24, p. 597–612.

Tesoriero, A., Spruill, T., and Eimers, J., 2004, Geochemistry 

of shallow ground water in coastal plain environments in 

the southeastern United States—Implications for aquifer 

susceptibility: Applied Geochemistry, v. 19, p. 1471–1482.

Thom, B., 1970, Carolina bays in Horry and Marion Counties, 

South Carolina: Geological Society of America Bulletin, 

v. 81, p. 783–814.
Chapter G 103

Tiner, R., 2003a, Estimated extent of geographically isolated 

wetlands in selected areas of the United States: Wetlands, 

v. 23, no. 3, p. 532–540.

Tiner, R.W., 2003b, Geographically isolated wetlands of the 

United States: Wetlands, v. 23, no. 3, p. 494–516.

Tyndall, R., McCarthy, K., Ludwig, J., and Rome, A., 1990, 

Vegetation of six Carolina bays in Maryland: Castanea, 

v. 55, no. 1, p. 1–21.

Unrine, J., Jagoe, C., Brinton, A., Brant, H., and Garvin, N., 

2005, Dietary mercury exposure and bioaccumulation in 

amphibian larvae inhabiting Carolina bay wetlands: Environmental Pollution, v. 135, no. 2, p. 245–253.

Vepraskas, M., White, J., Amoozegar, A., Hesterberg, D., 

Huffman, R., Skaggs, R., Gregory, J., Broome, S., 

Caldwell, P., Davis, K., Abit, S., Smith, H., Zelasko, A., and 

Brownfield, C., 2007, Methodology to assess soil, hydrologic, and site parameters that affect wetland restoration, 

phase 2: Raleigh, N.C., North Carolina State University. 

Whitehead, D., 1964, Palynology and Pleistocene phytogeography of unglaciated eastern North America, in Wright, 

H., and Frey, D., eds., The Quaternary of the United States: 

Princeton, N.J., Princeton University Press, p. 417–432.

Willoughby, R., Howard, C.S., and Nystrom, P., 2005, Generalized geologic map of South Carolina: Columbia, S.C., 

South Carolina Department of Natural Resources.

Wise, M., McArthur, J., and Shimkets, L., 1997, Bacterial 

diversity of a Carolina bay as determined by 16S rRNA 

gene analysis—Confirmation of novel taxa: Applied and 

Environmental Microbiology, v. 63, no. 4, p. 1505–1514.

Chapter H. Karst-Depression Wetlands

Introduction

Karst depressions (sinkholes) are a landscape feature in 

many areas underlain by carbonate rocks (White, 1988). Sinkhole drainage patterns integrate effects of geologic structure, 

lithology, soil and regolith properties, and climate to produce 

a continuum ranging from freely and permanently drained 

to perennially ponded (White, 1988; Ford and Williams, 

1989). Within that continuum are sinkholes whose interiors 

meet the scientific and jurisdictional definitions of wetlands: 

prolonged inundation or saturation during the growing season, 

obligate or facultative wetland vegetation, and hydric soils 

(fig. H–1; Cowardin and others, 1979). In comparison with 

bottomland hardwoods, coastal marshes, and similar extensive 

systems, karst-depression wetlands constitute a small fraction 

of wetland area within the southeastern United States as a 

whole, but they can account for a much larger percentage of 

total wetland area in regions underlain by carbonate bedrock 

(Wolfe, 1996b). Karst-depression wetlands are typically 

isolated from each other geographically, though not necessarily hydrologically (Hendricks and Goodwin, 1952; Wolfe, 

1996a, b; Hill, 2007).

Numerous studies in the southeastern United States have 

shown that karst-depression wetlands contribute substantially 

to regional biodiversity, notably by supporting high levels 

of biogeographic endemism and disjunction for vascular 

plants and by providing habitat to reptiles, amphibians, and 

invertebrates (Ellis and Chester, 1989; Sutter and Kral, 1994; 

Buhlmann and others, 1999; Fleming and Van Alstine, 1999; 

Edwards and Weakley, 2001; Grant and others, 2003; Miller 

and others, 2005; Kirkman and others, 2012). However, many 

of these studies have been site-specific or focused on multiple 

sites in relatively close geographic proximity of one another. 

A literature synthesis concerning southeastern depressional 

wetlands (Kirkman and others, 2012) was restricted to the 

Coastal Plain and did not emphasize distinctions between 

karst and nonkarst systems.

Two classes of southeastern depression wetlands not 

treated in detail in this report warrant mention: Citronelle ponds 

and depressional mountain bogs. Folkerts (1997) described 

geographically isolated depression wetlands on the Citronelle 

Formation (Citronelle ponds) in the Eastern Gulf Coastal 

Plain of Mississippi, Alabama, and Florida and attributed their 

origins to subsidence caused by dissolution of kaolinite or iron 

oxide cementing materials rather than by dissolution of the 

deeply mantled subsurface limestone. The literature on these 

systems is sparse; their relationship to other geographically 

isolated depression wetlands of the Coastal Plain has yet to be 

resolved (D. De Steven, U.S. Forest Service, written commun., 

2014), and their mechanisms of formation are not known 

in sufficient detail to classify them as noncarbonate karst, 

pseudokarst, or nonkarst. Stronger arguments have been made 

for a noncarbonate karstic origin of some depressional mountain bogs of the Blue Ridge in Virginia and North Carolina 

where peat-derived organic acids have evidently accelerated 

dissolution of underlying quartzite and progressively enlarged 

depressions at the bedrock-regolith interface (Reed and others, 

1963; Whittecar and Lawrence, 1999). 

Geographic Range

Wetlands in topographic depressions are present in 

several karst regions of the southeastern United States. 

Three regions in which karst-depression wetlands have been 

relatively well studied include (1) the Highland Rim section of 

the Interior Low Plateaus Physiographic Province, especially 

the Eastern Highland Rim, Pennyroyal District, and Mitchell 

Plain subsections, (2) the Dougherty Plain ecoregion within 

the Coastal Plain Physiographic Province, and (3) the Valley 

Figure H–1. Harriston Ponds, Augusta County, Virginia. 

Photograph by Gary Fleming, Virginia Department of 

Conservation and Recreation.
106 Insular Ecosystems of the Southeastern United States

and Ridge Physiographic Province, especially the Shenandoah 

Valley (fig. H–2). Karst-depression wetlands are distributed 

widely across the Highland Rim in Tennessee, Kentucky, 

and southern Indiana, with notable concentrations in the 

Mitchell Plains, Eastern Highland Rim, and Pennyroyal 

District (Homoya and Hedge, 1985; Ellis and Chester, 1989; 

Wolfe, 1996b). The most extensively studied karst-depression 

wetlands in the Valley and Ridge Physiographic Province are 

in the Shenandoah Valley of Virginia (Buhlmann and others, 

1999; Fleming and Van Alstine, 1999), but examples have also 

been reported from northwestern Georgia (Greear, 1967) to 

central Pennsylvania (O’Driscoll and Parizek, 2003). Within 

Figure H–2. General geographic distribution of selected areas containing concentrations of karst-depression wetlands 

in the southeastern United States.

7

9

1

6

8

5

3

10

2

4

90° 85° 80° 75° 70°

40°

35°

30°

25°

Dougherty Plain ecoregion—From

 U.S. Environmental Protection

 Agency (2013)

Prominent research locations for 

 karst-depression wetlands

1 Wetland sites in Harrison County, Ind. (Homoya and Hedge, 1985)

2 Algood Wetland, Tenn. (Hill and others, 2006)

3 Cedar Hill Swamp, Tenn. (Ellis and Chester, 1989)

4 Anderson Pond, Tenn. (Ellis and Chester, 1989)

5 Mingo Swamp, Ala. (Ellis and Chester, 1989)

6 Arnold Air Force Base, Tenn. (for example, Wolfe and others, 2004)

7 Sagponds in Bartow County, Ga. (Greear, 1967)

8 Ichauway Reserve, Ga. (for example, Kirkman and others, 2000),

9 Shenandoah Valley sinkhole ponds, Va. (for example,

 Buhlmann and others, 1999)

10 Wetland sites in Centre County, Pa. (O'Driscoll and Parizek, 2008)

Figure H-2

0 100 200 300 400 KILOMETERS

0 100 200 300 400 MILES

Base modifed from U.S. Department of Commerce, 

Census Bureau, 2015, Mercator Auxiliary Sphere Projection

Horizontal datum is World Geodedic System of 1984

EXPLANATION

Physiographic province—From 

 Fenneman and Johnson (1946)

Cumberland Plateau section of 

 the Appalachian Plateaus

East Gulf Coastal Plain section

 of the Coastal Plain

Highland Rim section of the

 Interior Low Plateaus

Valley and Ridge

Blue Ridge

TENNESSEE

ALABAMA

GEORGIA

SOUTH

CAROLINA

NORTH CAROLINA

WEST

VIRGINIA

OHIO

PENNSYLVANIA

ILLINOIS INDIANA

MISSOURI

VIRGINIA

FLORIDA

KENTUCKY

NEW JERSEY

DELAWARE

MARYLAND

CONNECTICUT

RHODE 

ISLAND

MISSISSIPPI

LOUISIANA
Chapter H 107

the East Gulf Coastal Plain of the Coastal Plain Physiographic Province, karst-depression wetlands—locally called 

limesinks—have been described extensively throughout the 

Dougherty Plain, which is a Level IV ecoregion (U.S. Environmental Protection Agency, 2013) that extends from southeastern Alabama and northwestern Florida into southwestern 

Georgia (Hendricks and Goodwin, 1952; Beck and Arden, 

1984; Sutter and Kral, 1994; Golladay and others, 1999). 

Physical Geography

In general, the development of a karst-depression wetland 

first requires the formation of a karst depression (sinkhole) 

and then requires drainage characteristics that support wetland 

plant communities and soils. Allowing for climatic variability, 

the most important determinants of sinkhole formation 

and drainage characteristics are geologic. In particular, the 

lithology, stratigraphy, and structure of underlying bedrock 

and the physicochemical properties and thickness of the 

unconsolidated cover (regolith) interact to determine the 

spatial character and drainage of karst features (White, 1988). 

Lithology largely determines each bedrock unit’s susceptibility 

to karst formation, and stratigraphy and structure combine to 

determine the spatial arrangement of units that are susceptible 

or resistant to karst formation. Together, lithology, stratigraphy, and structure exert a major influence on groundwater 

gradients and on the hydraulic and storage characteristics of 

underlying karst aquifers (Beck and Arden, 1984; Brahana 

and Bradley, 1986; White, 1988; Hollyday and Hileman, 

1996; Wolfe and others, 1997). The regolith mediates movement of water from the land surface to the top of rock and 

exerts a strong influence on local relief, notably on sinkhole 

depth. Impermeable regolith restricts flow from a sinkhole to 

underlying aquifers, whereas permeable regolith facilitates 

such flow (Hendricks and Goodwin, 1952; Greear, 1967; 

Wolfe, 1996a, b; O’Driscoll and Parizek, 2008). Thick regolith 

generally obscures the surface expression of some karst 

features and may reduce the depth of sinkholes or fill them 

entirely (Greear, 1967; O’Driscoll and Parizek, 2008). The 

“perching” of water above the water table by impermeable 

regolith has been reported in all areas where the hydrology of 

southeastern karst-depression wetlands has been quantified 

(Hendricks and Goodwin, 1952; Greear, 1967; Wolfe, 1996a, 

b; O’Driscoll and Parizek, 2008; Hill and Neary, 2012). 

Commonly, a wetland’s relative moisture content is 

expressed in terms of “hydroperiod,” a word with numerous 

and, in some cases, inconsistent definitions (Nuttle, 1997). 

Despite efforts to broaden use of the term hydroperiod (for 

example Nuttle, 1997), the word is used here in its simplest 

sense: number of days in a given year that water is ponded 

in a given depression. The term “hydrologic regime” is used 

to describe broader patterns of hydrologic conditions defined 

by long-term variability of the timing, duration, depth, and 

frequency of water-level fluctuation, year to year and among 

different sites. The karst-depression wetlands considered here 

illustrate how bedrock geology and regolith influence the 

geomorphology and hydrology of karst-depression wetlands 

in three contrasting physiographic settings: the Highland Rim 

section of the Interior Low Plateaus Physiographic Province, 

the Valley and Ridge Physiographic Province, and the 

Dougherty Plain ecoregion of the Coastal Plain Physiographic 

Province (Fenneman and Johnson, 1946).

The Highland Rim

On the Highland Rim, variations in bedrock lithology are 

a major determinant of topography and drainage, including 

the distribution of karst-depression wetlands. Across most of 

its extent, the Highland Rim is underlain by gently dipping 

(typically less than 1 percent and rarely more than 2 percent 

gradient) Mississippian limestones overlying resistant cherts, 

shales, and siltstones of lower Mississippian age (Brahana 

and Bradley, 1986; Hess and others, 1989). Dips are generally oriented away from either of two broad regional uplift 

structures, the Nashville and the Jessamine (Lexington) 

geologic domes, whose exposed cores of karstic carbonate 

rocks underlie the Nashville Basin and Bluegrass Region 

(Lexington Plain), respectively (Fenneman, 1938; Ryder, 

1987). Sinkhole development is generally strongest in areas 

where the limestones are thickest and most soluble, notably 

the sinkhole plains of the Pennyroyal Plateau (Hess and others, 

1989), Eastern Highland Rim (Wolfe, 1996b), and Mitchell 

Plain (Unterreiner, 2006). Sinkhole development is weak or 

absent where the limestones are thinner and resistant cherts, 

shales, and siltstones approach the surface, such as along the 

fluvially dissected escarpment separating the Highland Rim 

from the Nashville Basin (Wolfe, 1996b). Areas such as The 

Barrens of the Eastern Highland Rim where karst development 

and surface drainage are somewhat weakly expressed, tend 

to be where karst-depression wetlands are most likely to be 

concentrated (Wolfe, 1996b). However, these systems, like 

sinkholes in general, are widely distributed across the Highland Rim (Homoya and Hedge, 1985; Ellis and Chester, 1989). 

Highland Rim soils are developed from insoluble 

material (residuum) left from the dissolution of limestone, 

extensively overlain by approximately 3–6 feet (about 

1–2 meters) of loess or by localized river or hillslope 

deposits (Love and others, 1956; Patterson, 1989). Wolfe 

(1996b) reported overall regolith thicknesses ranging from 

23 to 108 feet (about 7 to 33 meters) in an area with numerous 

karst-depression wetlands. Soils in and around karstdepression wetlands in Tennessee commonly feature fragipans 

(layers of low hydraulic conductivity and impeded drainage), 

and an association between these wetlands and one soil 

series, the Guthrie Silt Loam, has been widely reported (Love 

and others, 1956; Ellis and Chester, 1989; Patterson, 1989; 

Wolfe, 1996b). In contrast, although Homoya and Hedge 

(1985) found the air-photo base of county soil maps useful 

in locating karst-depression wetlands in southern Indiana, 
108 Insular Ecosystems of the Southeastern United States

they found only limited association between karst-depression 

wetlands and particular soil series and emphasized regolith 

thickness rather than fragipan presence as a soil correlate of 

karst-depression wetlands. 

Where the hydrology of Highland Rim karst-depression 

wetlands has been examined in detail, perching of surface 

water by impermeable unconsolidated layers above 

the seasonal water table has been commonly observed 

(Wolfe, 1996a, b; Hill and Neary, 2007, 2012). A second 

mode of sinkhole inundation and drainage observed by 

Wolfe (1996a, b) is direct intersection of the seasonal water 

table with the sinkhole bottom. Wolfe (1996a, b) noted such 

intersection only in the deepest depressions of the study area, 

which he called “compound sinks” (fig. H–3), emphasizing 

their coalescent geomorphic character and multiple internal 

drains in contrast to shallower perched “karst pans” with no 

visible internal drains. Compound sinks with close interactions 

between surface water and groundwater can experience rapid 

fluctuations in water levels—as much as 6 feet (about 2 meters) 

within 1 to 3 days—whereas water-level fluctuation in karst 

pans typically occurs more gradually and within a narrow 

range (Wolfe, 1996a, b). For compound sinks and karst pans, 

hydroperiods exceeding 200 days per year are common (Wolfe, 

1996a, b; Wolfe and others, 2004; Hill and Neary, 2012).

In a shallow karst pan on the Eastern Highland Rim 

of Tennessee, Hill (2007) and Hill and Neary (2007) found 

surface runoff to be the dominant input to the annual water 

budget. Deep recharge accounted for more than half of the 

total outflow from this wetland despite relatively high rates 

of evapotranspiration during the growing season, suggesting 

that groundwater recharge is an important function of some 

karst-depression wetlands (Hill and Neary, 2012).

The Valley and Ridge

Investigations of karst-depression wetlands have been 

conducted in several regions of the Valley and Ridge Physiographic Province. Greear (1967) described the hydrology and 

ecology of karst-depression wetlands (“sagponds”) along the 

eastern margin of the Great Valley of northwestern Georgia 

near the valley’s boundary with the Piedmont Physiographic 

Province. His model of their origins and evolution emphasized 

progressive filling with relatively fine-grained sediment to the 

point of “extinction” or complete filling. For these wetlands, 

hydroperiod was positively related to regolith thickness 

within the depression and inversely related to sinkhole depth. 

Hydroperiod is thus presented as dynamic, growing progressively longer over time until abruptly shortened by complete 

filling (Greear, 1967). 

O’Driscoll and Parizek (2003, 2008) examined drainage 

characteristics, hydroperiods, and groundwater/surface-water 

gradients in karst-depression wetlands in Centre County, 

Pennsylvania. Primary water inputs to these wetlands are 

direct precipitation and groundwater less than 20 feet (about 

Figure H–3. Nyssa aquatica stand in Westall Swamp, Arnold Air Force Base, Coffee County, Tennessee. 

Photograph by William Wolfe, U.S. Geological Survey.
Chapter H 109

6 meters) below land surface. Counterintuitively, hydroperiods 

were found to be longer in sinkholes with sandy unconsolidated 

fill than in those filled with clay. Perched water in these ponds, 

as well as mounding of groundwater directly below them, 

indicates that low-permeability zones in the regolith impede 

but do not prevent infiltration. In contrast to the Shenandoah 

Valley sinkhole ponds (Buhlmann and others, 1999) and 

Georgia sagponds (Greear, 1967), the unconsolidated fill in the 

Centre County depressions is residual rather than depositional 

in origin (O’Driscoll and Parizek, 2003, 2008). 

One of the most distinctive groups of karst-depression 

wetlands within the Valley and Ridge is the Shenandoah 

Valley “sinkhole ponds” (figs. H–4 and H–5). Steep northeast/

southwest-trending folds and widespread faults have produced 

a landscape of generally parallel valleys and ridges from 

which the province takes its name (Fenneman, 1938). The 

folds and faults also juxtapose rocks of contrasting lithology 

in spatial arrangements far more complex than the top-tobottom stratigraphic order of the Highland Rim. Limestones, 

dolomite, and argillaceous carbonates underlie about one-third 

of the area of the Valley and Ridge (Johnson and others, 

2011), but their occurrences are discontinuous and can be 

widely separated. Carbonate rocks, overlain with varying 

thicknesses of alluvium, typically form broad valley bottoms 

separated by ridges composed of chert-capped residuum 

(Hollyday and Hileman, 1996; Wolfe and others, 1997). 

Within the Shenandoah Valley, karst-depression wetlands 

are concentrated along an approximately 55-mile (roughly 

89-kilometer) segment of the valley’s southeastern margin, 

from Page County to Augusta County, Virginia, corresponding 

to the western toe of the Blue Ridge Physiographic Province 

(Buhlmann and others, 1999). 

Karst-depression wetlands of the western toe of the 

Blue Ridge occupy a highly distinctive hydrogeologic setting 

that has been studied as a resource for industrial groundwater 

supply (Hinkle and Sterrett, 1976, 1978; Hollyday and 

Hileman, 1996). Folded karstic dolomites of Cambrian to 

Ordovician age are covered by a mantle that is generally at 

least 100 feet (about 30 meters) thick. This mantle generally 

consists of fine-grained residuum overlain by coarse, permeable 

stream deposits and rubble derived from the steep slopes of the 

Blue Ridge that developed from siliciclastic rocks, commonly 

quartzite. The steep, impermeable upper slopes produce 

large volumes of runoff, which infiltrates readily through the 

coarse stream deposits and rubble to recharge the underlying 

karst aquifers (Hinkle and Sterrett, 1976, 1978; Hollyday and 

Hileman, 1996). This “western toe” setting is not confined 

to the Shenandoah Valley but can be found in discontinuous 

occurrences along the eastern margin of the Valley and Ridge 

from Alabama to Pennsylvania (Hollyday and Hileman, 1996).

In contrast to their floristics, the hydrology of the 

Shenandoah Valley sinkhole ponds has received relatively 

limited scientific attention. Buhlmann and others (1999) 

assigned each of 36 wetlands to 1 of 6 hydrologic classes, 

Figure H–4. Deep Pond, Maple Flats Pond Complex, Augusta County, Virginia. Photograph by Gary Fleming, 

Virginia Department of Conservation and Recreation.
110 Insular Ecosystems of the Southeastern United States

ranging from permanently ponded to highly ephemeral, 

based on qualitative observations of ponding conditions during 

repeated site visits over a decade, supplemented by air photos 

as early as the 1930s. They noted large hydrologic variability 

between ponds and, for some ponds, variability from one year 

to another. Although qualitative in nature, these observations 

support the argument by Buhlmann and others (1999) that 

karst-depression wetlands of the Shenandoah Valley constitute 

a coherent hydroecological system whose preservation 

depends upon the resilience offered by high degrees of spatial 

and temporal hydrologic variability.

The Dougherty Plain

The Dougherty Plain is one of several Coastal Plain 

“belts” that wrap, crescent-like, around the southern end of 

the Appalachians (Fenneman and Johnson, 1946; U.S. Environmental Protection Agency, 2013). In contrast to the 

lithologic variability of the Highland Rim and the structural 

and geomorphic complexity of the Shenandoah Valley, the 

Dougherty Plain presents a relatively simple geologic setting. 

The area is dominated by a single geologic unit, the upper 

Eocene Ocala Limestone, which is a relatively pure karstic 

limestone with high primary and secondary porosity that 

forms the Upper Floridian aquifer (Torak and others, 1993). 

Locally known as the Ocala aquifer (Hicks and others, 1981), 

the Upper Floridan aquifer is a major water-bearing unit of 

the Floridian aquifer system (Miller, 1986). In the Dougherty Plain, the Ocala Limestone is typically about 165 to 230 

feet (roughly 50 to 70 meters) thick and is confined at its base 

by the middle Eocene Lisbon Formation and discontinuously 

at its top by a layer of residuum with a typical thickness range 

of approximately 45 to 110 feet (roughly 13 to 33 meters) 

(Hicks and others, 1981). The residual mantle is generally 

reddish sandy clay (Hicks and others, 1981), but is not 

uniform; geophysical logs obtained near Albany, Georgia, 

show an extensive clay-rich layer that restricts infiltration and 

perches shallow groundwater above the regional water table 

of the Upper Floridian aquifer (Torak and others, 1993).

Sinkholes (limesinks) are widespread throughout much 

of the Dougherty Plain, but the relation of karst-depression 

wetlands to the overall sinkhole population is difficult to 

discern from the published literature. Torak and others (1993) 

reported a general range of sinkhole depths in the Dougherty 

Plain from less than 3 to about 25 feet (less than 1 to about 

8 meters) but provided little information on their drainage 

characteristics beyond noting that many are filled with water 

year round. Hicks and others (1981) identified sinkhole 

collapse, along with stream-channel erosion, as a mechanism 

for removing the residual confining unit, producing locally 

unconfined groundwater conditions. A reasonable inference is 

that deep sinkholes are more likely to penetrate the residual 

confining unit than shallow ones. Where depths of limesink 

wetlands have been reported, they have not exceeded 6.5 feet 

Figure H–5. Deep Run Ponds, Rockingham County, Virginia. Photograph by Gary Fleming, Virginia Department of 

Conservation and Recreation.
Chapter H 111

(about 2 meters) (Hendricks and Goodwin, 1952; Golladay 

and others, 1999; Opsahl, 2005). Palik and others (2000) 

distinguished “shallow depression” upland ecosystems from 

“depression” wetlands but did not specify a depth threshold 

separating the two; their finding that shallow depressions are 

drier than deep depressions is consistent with results reported 

by Hendricks and Goodwin (1952).

Information on groundwater/surface-water interactions in 

and around limesink wetlands is largely limited to the analysis 

of Hendricks and Goodwin (1952). All but 2 of 13 sites had 

at least 15 observations of reversed gradients between the 

pond stages and measured groundwater levels, and 9 sites had 

more than 30 such observations, suggesting that flow from the 

groundwater system to these ponds is common when watertable elevations are high. However, these wells had relatively 

short screened intervals (approximately 3 feet, or 0.9 meter) 

and apparently shallow depth (not reported but probably less 

than 25 feet, or about 8 meters, based on maximum estimated 

groundwater levels). These well characteristics suggest that 

their water levels may be best interpreted as indications 

of hydraulic head within the residuum and not necessarily 

as measurements of the water-table elevation. Hydrologic 

interactions between the limesink wetlands and the underlying 

Upper Floridan aquifer do not appear to have been examined. 

Stress and Disturbance Regimes

The most obvious, and in many ways dominant, abiotic 

stress factor in karst-depression wetlands is water. Fire, 

disturbance, and edaphic stresses may be present and even 

important, but they operate within ecological constraints 

determined by hydrologic stress, chiefly in the form of 

inundation or soil saturation. 

Hydrologic Stress Regimes of Karst-Depression 

Wetlands

Numerous investigators have characterized karst-depression 

wetland hydroperiods for time scales ranging from 1 year 

to a decade (Hendricks and Goodwin, 1952; Wolfe, 1996a, 

b; Golladay and others, 1997; Battle and Golladay, 2001; 

O’Driscoll and Parizek, 2003; Hill and Neary, 2012). One 

general observation that can be drawn from these studies is that 

annual hydroperiod is highly variable, interannually and site to 

site, even within short distances. O’Driscoll and Parizek (2003) 

noted annual hydroperiods ranging from 33 to 192 days among 

17 karst-depression wetlands monitored in 1999 within a 54-acre 

(22-hectare) area of the Valley and Ridge in Pennsylvania. 

Knox (1997) reported hydroperiods of 4 to 16 months in one 

Shenandoah Valley sinkhole pond monitored for 9 years. 

Short-term ecological responses to variation in annual 

hydroperiod have been observed, including measureable 

differences in the composition and density of aquatic 

macroinvertebrate communities of Dougherty Plain limesinks 

(Golladay and others, 1997, 1999) and in tree recruitment 

in a Highland Rim karst swamp (Wolfe and others, 2004). 

However, the high level of variability in annual hydrologic 

conditions of karst-depression wetlands suggests that caution 

is in order when trying to relate wetland ecosystems to 

hydrology based on a few years of hydrologic observations 

at a limited number of sites. For example, Knox (1997) 

observed apparently universal mortality among Helenium 

virginicum, a federally listed threatened species, during 

prolonged (16 months) inundation in a 2.5-acre (1-hectare) 

Shenandoah Valley sinkhole pond, but concluded that blocking 

of light by a mat of floating live and dead vegetation was 

a more likely cause of mortality than duration of ponding 

because H. virginicum populations were undiminished by 

equal inundation depths and durations in nearby sites where 

floating vegetation mats were absent.

The most detailed analysis found in this review of the 

ecological role of hydroperiod in a karst-depression wetland 

is a series of studies examining Sinking Pond, an 87-acre 

(35-hectare) compound sink on Arnold Air Force Base, 

Tennessee, on the eastern Highland Rim (fig. H–6). This 

wetland supports an assemblage of Quercus lyrata, Betula 

nigra, and Pleopeltis polypodioides with a preliminary ranking 

of G1 (Wolfe and others, 2004; NatureServe, 2015). Wolfe 

(1996a, b) established the pond’s hydrologic framework and 

related tree distribution patterns to elevation and ponding 

duration. Wolfe (1996b) explicitly discussed the balance of 

abiotic and competitive stress factors and noted suppression 

of Q. lyrata recruitment in the deepest parts of the pond—

attributed to unspecified effects of winter ponding—despite 

widespread germination of seedlings produced by apparently 

healthy adult trees. 

McCarthy and Evans (2000) found general failure of 

Q. lyrata recruitment in Sinking Pond beyond the sapling stage, 

except in areas of shallow (less than 1.5 feet or about 0.5 meter) 

maximum ponding depth. They also examined age distribution 

and spatial patterns of adult Q. lyrata and concluded that an 

increase in hydroperiod had occurred in Sinking Pond sometime 

after 1950, possibly related to land-use change or geologic 

factors (McCarthy and Evans, 2000). Wolfe and others (2004) 

confirmed that an increase in hydroperiod had indeed caused 

recruitment failure in Sinking Pond, based on seedling plots, 

spatial distributions of Q. lyrata life-history stages, and tree-ring 

analysis. Having eliminated possible anthropogenic causes, 

Wolfe and others (2004) attributed increased hydroperiod 

and recruitment failure in Sinking Pond to a step increase in 

regional average precipitation known to have occurred around 

1970 (Karl and Knight, 1998; McCabe and Wolock, 2002), 

based on ages and elevations of Q. lyrata and approximately 

148 years of modeled daily pond stage. A tripling of the likelihood of annual inundation exceeding 200 days (17 percent 

before 1970, 54 percent after) represents the hydrologic 

threshold separating successful and repressed recruitment of 

Q. lyrata in Sinking Pond (Wolfe and others, 2004, p. 33–36).
112 Insular Ecosystems of the Southeastern United States

Fire, Disturbance, Water Quality, and 

Edaphic Stress

The roles of fire and mechanical disturbance in the abiotic 

stress regimes of karst-depression wetlands have received 

limited scientific attention, with most of this research having 

been conducted in the Dougherty Plain. Limesinks of the 

Dougherty Plain are generally embedded within a longleaf 

pine ecosystem (fig. H–7) in which savanna-like physiognomy 

and herbaceous species richness are maintained by frequent 

(1–10-year interval), low-intensity fires (Drew and others, 

1998). Drew and others (1998) presented a conceptual model 

of interactions between fire and hydroperiod as a determinant 

of ecosystem composition, structure, and trajectory. Martin 

(2006) examined the effects of mowing and fire in 10 limesink 

wetlands, showing that these treatments produced a shift from 

woody to herbaceous vegetation facilitated by soil-based seed 

banks that included herbaceous species; such shifts might be 

stable if the fire and disturbance regimes were maintained. 

Conversely, shifts in the fuel structure, possibly driven by 

climate change or invasive plants such as Lygodium japonicum, 

might alter either the maintenance and composition of seed 

banks or the general behavior of fire in these wetlands, potentially producing new plant assemblages rather than restoring 

native herbaceous communities (Martin, 2006). 

Historically, fire was important in maintaining the 

barrens matrix in which Highland Rim karst-depression 

wetlands are locally concentrated (Wolfe, 1996b), but 

little if any detailed examination of fire’s role within these 

wetlands has been attempted. One Highland Rim karst 

pan, Goose Pond in Coffee County, Tennessee, supports a 

natural marsh community dominated by Panicum hemitomon

(DeSelm, 1973) with little apparent need of fire or mowing 

to keep woody vegetation confined to the pond margins, 

though one prescribed burn reportedly was conducted in 

the 1990s (Geoff Call, U.S. Fish and Wildlife Service, 

oral commun., 2015).

Water quality has received limited attention as an 

abiotic stressor in karst-depression wetlands, but it appears 

to play a discernible ecological role in some contexts. 

Low pH (less than 6) and low acid neutralizing capacity 

(less than 20 acid-base microequivalents per liter) of some 

Shenandoah Valley sinkhole ponds create an apparently 

Figure H–6. Two views of Sinking Pond, Arnold Air Force Base, Coffee County, Tennessee: A, fully drained with leaves on in 

the summer, and B, fully drained with snow in the winter. Visible in both photos but most striking in winter is the moss line at 

the base of the trees, marking the typical high-water elevation. Photographs by Michael Hodge, retired, Arnold Engineering 

Development Center.

A

B
Chapter H 113

stressful but tolerable environment for the eggs and larvae of 

Ambystoma tigrinum (tiger salamander), which is state-listed 

as endangered in Virginia (Downey and others, 1999). Among 

Dougherty Plain limesinks, cypress-gum swamps, cypress 

savannahs, and grass-sedge marshes each have distinctive 

water-quality profiles, with swamps having greater concentrations of nutrients and dissolved organic carbon than the 

other community types (Battle and Golladay, 2001; Liner and 

others, 2008). Water-quality patterns and vegetation structure 

correlate to at least some measures of animal diversity. Macroinvertebrate diversity was highest in marshes and lowest in 

swamps (Battle and Golladay, 2001). Swamps had the lowest 

larval species richness for frogs and toads but the highest for 

salamanders (Liner and others, 2008). 

Edaphic stress not related to soil saturation is another 

component of the karst-depression wetland stress regimes that 

has received little attention. The fragipans that characterize 

soils in areas of the Eastern Highland Rim are found in karst 

depressions and in the surrounding uplands (Love and others, 

1956). Wolfe (1996b) noted that rooting depths of Quercus 

lyrata and other trees in Highland Rim karst swamps were 

limited to about 3 feet (approximately 1 meter) or less by this 

fragipan, independent of soil saturation, producing a characteristic mode of failure through tipping over, root mat intact, 

rather than through broken trunks.

Community Types and Vegetation 

Dynamics

At the broadest level, karst-depression wetland 

community types can be divided into major classes based on 

vegetation physiognomy. Swamps are wetlands with a closed 

forest canopy and moderate-to-sparse shrub cover, savannas 

have a relatively open canopy and herbaceous understory, 

and marshes are predominantly herbaceous, often without a 

distinct overstory (Kirkman and others, 2000). Individual karst 

depressions may contain one or more of these broad vegetation 

types, with or without inclusions of open-water habitat.

Swamps in karst-depression wetlands are dominated 

by inundation-tolerant woody taxa. Many karst-depression 

wetlands on the Highland Rim are swamps; however, marsh 

and open-water habitats are also present (Ellis and Chester, 

1989; McKinney, 1989; Wolfe, 1996b). Typically, forest 

species composition differs substantially between depression wetlands and the surrounding uplands—which may be 

mixed mesophytic forest or oak-hickory forest—based on 

differential tolerance of prolonged soil saturation, which 

causes anoxic conditions (Roberts and others, 2004). Upland 

species intolerant of anoxic soil are effectively excluded from 

karst-depression wetland communities, providing competitive 

Figure H–7. Neyami Pond, Lee County, Georgia. Photograph by Alan M. Cressler, U.S. Geological Survey. 
114 Insular Ecosystems of the Southeastern United States

advantage to obligate and facultative wetland species. On 

the Highland Rim, common tree-canopy constituents in 

karst-depression wetlands include Acer rubrum, several 

Quercus species (Q. bicolor, Q. lyrata, Q. phellos, and 

Q. palustris), Fraxinus pennsylvanica, Ilex decidua, Liquidambar styraciflua, Platanus occidentalis, Nyssa sylvatica, 

and Ulmus americana (Hannan and Lassetter, 1982; Ellis 

and Chester, 1989; Patterson, 1989; Cranfill, 1991; Roberts 

and others, 2004). Common components of the shrub layer 

include Alnus serrulata, Sparganium americanum, Photinia 

melanocarpa, Cephalanthus occidentalis, Decodon verticillatus, Hibiscus moscheutos, Ilex verticillata, Itea virginica, 

Parthenocissus quinquefolia, Toxicodendron radicans, 

Rosa palustris, Sambucus nigra ssp. canadensis, and 

Vaccinium species (Delcourt, 1979; Ellis and Chester, 1989; 

Cranfill, 1991; Roberts and others, 2004). Swamp habitat in 

depression wetlands on the Dougherty Plain is commonly 

dominated by Taxodium ascendens and Nyssa biflora in the 

tree canopy, and thus is referred to as cypress-gum swamp, 

with Liquidambar styraciflua and Quercus and Salix species 

at swamp margins (Hendricks and Goodwin, 1952; Drew and 

others, 1998; Kirkman and others, 2000). Shrub components 

in these cypress-gum swamps can include Cephalanthus 

occidentalis, Ilex myrtifolia, Styrax americanus, Leucothoe 

racemosa, Clethra alnifolia, and Lyonia lucida (Hendricks 

and Goodwin, 1952; Kirkman and others, 2000). 

Grass-sedge marshes are primarily composed of 

submerged and emergent aquatic and semiaquatic plants. 

Such marshes are an important community type in limesink 

wetlands of the Dougherty Plain, with community constituents 

from the genera Rhexia, Rhynchospora, Lespedeza, Juncus, 

Panicum, Typha, Scirpus, Xyris, and Andropogon (Hendricks 

and Goodwin, 1952; Drew and others, 1998; Kirkman 

and others, 2000). In these marshes, important aquatic 

constituents include species of the genera Ludwigia, Myriophyllum, Nymphaea, Limnobium, Potamogeton, and Lemna

(Hendricks and Goodwin, 1952; Kirkman and others, 2000). A 

few examples of marsh habitat in karst depressions have also 

been described on the Highland Rim. Goose Pond (McKinney, 

1989), Anderson Pond, and Mingo Pond (Delcourt, 1979; 

Ellis and Chester, 1989), all on the Eastern Highland Rim of 

Tennessee, support zones of aquatic macrophytic communities 

and grass-sedge communities along gradients of inundation 

depth surrounded by swamp forest. At these sites, common 

marsh taxa include Dulichium arundinaceum, Pontederia 

cordata, and species of the genera Sagittaria, Eleocharis, 

Scirpus, Rhynchospora, Panicum, Juncus, and Polygonum, 

and aquatic macrophytes inhabiting semipermanent pools 

include Myriophyllum tenellum, Ceratophyllum echinatum, 

C. demersum, and species of the genera Lemna, Polygonum

and Spirodela (Delcourt, 1979; Ellis and Chester, 1989; 

McKinney, 1989).

In Shenandoah Valley sinkhole ponds, physiognomy can 

vary from savanna-woodland types to entirely herbaceous 

with sparse-to-absent shrub cover (fig. H–8). Woodland 

areas commonly include scattered individuals and groves 

of Quercus palustris (Fleming and Van Alstine, 1999). 

Important herbaceous species include Eleocharis acicularis, 

Agrostis perennans, Hypericum boreale, Bidens frondosa, 

Viola lanceolata, and species of the genera Panicum and 

Rhexia (Buhlmann and others, 1999; Fleming and Van Alstine, 

1999; Fleming and Patterson, 2012). A community subtype 

found in these sinkhole ponds and in some sinkhole ponds in 

Coffee County, Tennessee, tends to develop in comparatively 

drier, seasonally flooded areas and is strongly dominated by 

Carex barrattii, which is disjunct to the Coastal Plain (Sorrie 

and Weakley, 2001; Fleming and others, 2012). An additional 

subtype strongly dominated by Orontium aquaticum is 

limited to a single, permanently flooded pond with relatively 

stable water levels because of continual groundwater inputs 

(Buhlmann and others, 1999).

The dynamics of plant community composition in karstdepression wetlands are regulated by a number of interacting 

factors, probably the most important of which is hydroperiod 

(Roberts and others, 2004). Within semipermanently inundated 

karst-depression wetlands of the Highland Rim, vegetation 

is typically zoned. The central, most permanently inundated 

zone is generally dominated by aquatic herbaceous species 

because of the exclusion of woody species. The surrounding 

zone of seasonally variable inundation generally supports tree 

and understory communities similar to those of seasonally 

inundated depressions but with the addition of two notable 

species: Quercus lyrata and Cephalanthus occidentalis. 

Common species within the most permanently inundated 

zone include Panicum hemitomon, Juncus repens, Dulichium 

arundinaceum, Rhynchospora corniculata, Hibiscus 

moscheutos, Saccharum baldwinii, Scirpus cyperinus, and 

species of the genera Potamogeton, Lemna, and Polygonum, 

with a few observed instances of semipermanently inundated 

depressions dominated by relatively homogenous stands 

of Nyssa aquatica (Roberts and others, 2004). Hannan and 

Lassetter (1982) reported stratification of the herbaceous 

community at Brodhead Swamp Forest on the Eastern 

Highland Rim in Kentucky, with Proserpinaca palustris being 

the only herbaceous species in the deepest zones of longest 

hydroperiod and with a more diverse assemblage of Carex 

lupuliformis, C. normalis, C. typhina, Fraxinus pennsylvanica, 

and Acer rubrum dominating the herbaceous-seedling strata 

in zones of shallower water and shorter hydroperiod. Species 

diversity in deep water areas was generally low except where 

increased microrelief was provided by stumps, fallen logs, and 

the bases of trees.

In a study of limesinks on the Dougherty Plain, 

Kirkman and others (2000) found cypress-gum swamps to 

have generally higher organic accumulations of soil and 

longer hydroperiods than cypress savannas or grass-sedge 

marshes and to develop at locally lower elevations than the 

other community types. Inundation depth and timing also 
Chapter H 115

appear to be of central importance in regulating vegetation 

zones in Shenandoah Valley sinkhole ponds, where episodic 

establishment of Quercus palustris is believed to be enabled 

by prolonged droughts, and several community types are 

restricted to zones of semipermanent inundation (Fleming 

and Van Alstine, 1999).

In addition to hydroperiod and inundation depth, 

other important controls on community composition in 

karst-depression wetlands include disturbance regimes, 

spatial patterns of shading, and ecological processes such 

as herbivory, disease, and competition (Roberts and others, 

2004). Small-scale, frequent disturbance events can include 

the mortality of individual trees and localized damage from 

storms. Disturbance from fire can also be an important factor 

during dry periods and can affect the surrounding upland 

forest community even if wet conditions exclude fire from 

depressions themselves. For limesink wetlands of the Dougherty Plain, Kirkman and others (2000) proposed a model 

whereby interactions of fire frequency and inundation depth 

and duration regulate community type, ranging from grasssedge marsh (maintained by inundation-based mortality of 

Pinus species and fire-based mortality of cypress and hardwoods) to cypress-gum swamp (maintained by inundationbased mortality of hardwoods and pines). In karst-depression 

wetlands of this region, species-rich herbaceous communities 

may be largely sustained by fire regimes, such that anthropogenic fire suppression may result in hardwood encroachment 

and establishment (Martin, 2006; Martin and Kirkman, 2009).

Contributions to Regional Biodiversity

Like other small, geographically isolated wetlands in 

the southeastern United States, karst-depression wetlands 

contribute to regional floral and faunal biodiversity at levels 

disproportionate to their collective geographic size (Kirkman 

and others, 2012). In addition to supporting regionally rare 

plant taxa, these wetlands also provide important habitat for 

animals including birds, amphibians, and invertebrates.

Endemic and Biogeographically Disjunct 

Plant Taxa

While many of the plant taxa composing vegetation 

communities in karst-depression wetlands are geographically 

widespread, karst-depression wetlands in several regions 

are known to support certain plant taxa that are rare locally, 

regionally, and in some cases globally, and thus make 

important contributions to biodiversity (DeSelm, 1990; 

Sutter and Kral, 1994). For example, Helenium virginicum

(fig. H–9) appears to be narrowly endemic to Shenandoah 

Valley sinkhole ponds and is federally listed as threatened 

(Knox, 1997; Fleming and Van Alstine, 1999). Hypericum 

lissophloeus is apparently restricted only to karst-depression 

wetlands in two counties of the Florida panhandle on the 

Dougherty Plain (Sorrie, 2012). Ranked G1, Isoetes virginica

is present in karst-depression wetlands of the Shenandoah 

Figure H–8. Davidson Run Pond, Rockbridge County, Virginia. Photograph by Gary Fleming, Virginia 

Department of Conservation and Recreation.
116 Insular Ecosystems of the Southeastern United States

Valley and also in related wetland contexts, such as woodland 

ponds and streams (Fleming and Van Alstine, 1999; Franklin 

and Finnegan, 2006). At one well-studied site containing 

a cluster of karst-depression wetlands, Arnold Air Force 

Base on the Highland Rim of Tennessee, dozens of rare and 

endangered plants have been documented, primarily within or 

near karst-depression wetlands (Wolfe, 1996b). In a number 

of cases, botanical surveys of karst-depression wetlands have 

produced new additions to state flora (Homoya and Hedge, 

1985; Ellis and Chester, 1989).

A primary reason for the concentration of locally rare 

plant taxa in and around karst-depression wetlands is support 

for taxa that are biogeographically disjunct from home 

ranges in other physiographic provinces. For karst-depression 

wetlands of the Interior Low Plateaus, the primary pattern of 

disjunction appears to be to the Coastal Plain (DeSelm, 1990), 

a phenomenon that has been attributed to these wetlands functioning as refugia for relict species from the Tertiary Period 

when the Interior Low Plateaus included vast swamps with 

a subtropical climate (Hannan and Lassetter, 1982; Homoya 

and Hedge, 1985). Examples of Coastal Plain taxa present in 

karst-depression wetlands of the Interior Low Plateaus include 

Ludwigia sphaerocarpa in Mingo Swamp in Tennessee 

(Ellis and Chester, 1989); Panicum hemitomon, Lachnanthes 

caroliana, Lobelia canbyi, Xyris laxifolia var. iridifolia, 

Sagittaria graminea var. graminea, and Woodwardia virginica

in Goose Pond in Tennessee (McKinney, 1989; Grant and 

others, 2003); Platanthera flava var. flava, Triadenum walteri, 

and Ranunculus pusillus at Brodhead Swamp Forest on the 

Eastern Highland Rim in Kentucky (Hannan and Lassetter, 

1982); and Carex gigantea, C. decomposita, Triadenum 

walteri, Woodwardia areolata, Rhynchospora corniculata, 

Itea virginica, and Ranunculus pusillus in karst-depression 

wetlands of southern Indiana (Homoya and Hedge, 1985). 

Coastal Plain affinities have also been noted for Shenandoah 

Valley sinkhole ponds, which support populations of Panicum 

hemitomon and Lachnanthes caroliana, in addition to Coastal 

Plain taxa such as Utricularia striata, Carex barrattii, 

Eleocharis melanocarpa, and E. robbinsii, all of which are 

ranked S1 or S2 in Virginia (Buhlmann and others, 1999; 

Fleming and Van Alstine, 1999; Townsend, 2015).

Certain other plant taxa in karst-depression wetlands of 

the Interior Low Plateaus and the Valley and Ridge are disjunct 

to ranges farther north or west. Carex aquatilis, for example, 

has a widespread distribution in Canada and the Rocky 

Mountains of the United States, but is ranked S1 in Virginia 

where it is known only from the Big Levels–Maple Flats region 

(Fleming and Van Alstine, 1999). Cedar Hill Swamp on the 

Western Highland Rim of Tennessee supports several State-rare 

emergent and floating aquatic species, including Ranunculus 

flabellaris (disjunct to New England and the northern Midwest 

States) and Carex rostrata (disjunct from the most northern 

regions of New England and the northern Midwest) (Chester 

and Ellis, 1989; Ellis and Chester, 1989).

Figure H–9. Helenium virginicum at Twin Ponds, Maple Flats Pond Complex, Augusta County, Virginia. 

Photograph by Gary Fleming, Virginia Department of Conservation and Recreation.
Chapter H 117

In addition to the presence of regionally rare individual 

species, karst-depression wetlands also support rare assemblages and communities. The deeper zones of Sinking Pond 

on the Highland Rim of Tennessee support an assemblage 

of overcup oak, river birch, and resurrection fern (Quercus 

lyrata, Betula nigra, and Pleopeltis polypodioides, respectively) that is ranked G1 (Patterson, 1989; Wolfe and others, 

2004) and has thus far not been documented anywhere else in 

the world (NatureServe, 2015). Similarly, Shenandoah Valley 

sinkhole ponds support an herbaceous assemblage dominated 

by Carex barrattii that is known from only two karst ponds in 

Augusta County, Virginia (Fleming and Patterson, 2012).

Animal Habitat

A number of birds and mammals rely on karst-depression 

wetlands to provide some of their habitat needs and in turn 

shape these ecosystems through trophic interactions. Important 

mammalian predators in karst-depression wetlands include 

Neovison vison (American mink), Procyon lotor (common 

raccoon), and Lontra canadensis (North American river otter) 

(Roberts and others, 2004). Karst-depression wetlands provide 

important breeding grounds and foraging opportunities for a 

number of bird species. On the Highland Rim of Tennessee, 

Aix sponsa (wood duck) uses karst-depression wetlands and 

other forested wetlands year round (Roberts and others, 2004). 

Karst-depression wetlands on the Western Highland Rim 

support populations of Ardea herodias (great blue heron) and 

appear to have served as a refugia for this species during the 

1960s when a number of large colonies were lost in Tennessee 

due to habitat destruction (Fleming and others, 1984; 

Pullin, 1990). 

One of the most biologically important features of 

karst-depression wetlands is their role as habitat for amphibians and reptiles. For example, Shenandoah Valley sinkhole 

ponds provide habitat to locally rare Ambystoma tigrinum

(eastern tiger salamander) and Clemmys guttata (spotted 

turtles) (Buhlmann and others, 1999). Seasonally inundated, 

isolated wetlands of the southeastern Coastal Plain, such as 

the limesink wetlands on the Dougherty Plain, also provide 

important amphibian habitat (Smith and others, 2006), helping 

support as many as 36 amphibian species (Moler and Franz, 

1987). A notable example is Ambystoma cingulatum (frosted 

flatwoods salamander), which is endemic to a small part of 

the Coastal Plain and is thus ranked G2 and is federally listed 

as threatened (Sutter and Kral, 1994). Certain areas of the 

Highland Rim, such as the upper Duck River watershed in 

Coffee County, Tennessee, have been documented as sites 

of exceptional biodiversity for amphibians and reptiles, due 

in part to habitat heterogeneity afforded by karst-depression 

wetland ecosystems (Niemiller, 2005). Miller and others 

(2005) documented several species of State or Federal concern 

associated with wetland habitat in karst-depression wetlands 

on the Highland Rim, including Ambystoma talpoideum

(mole salamander), Hemidactylium scutatum (four-toed 

salamander), and Hyla gratiosa (barking treefrog). The 

G3-ranked Lithobates capito (gopher frog) occupies some 

karst-depression wetlands of the Highland Rim in Tennessee, 

where it is ranked S1, and in the Dougherty Plain, where it 

is ranked S2 in Alabama and S3 in Georgia (Sutter and Kral, 

1994; Miller and others, 2005; NatureServe, 2015).

In some cases, certain reptiles and amphibians tend to 

remain within the depression-wetland boundary—including 

Nerodia sipedon (northern water snake), N. erythrogaster

(plain-bellied water snake), Lithobates catesbeianus 

(American bullfrog), L. clamitans (green frog), and Chelydra 

serpentina (common snapping turtle)—whereas others use 

both wetland and adjacent upland habitats (Miller, 1995). 

Those amphibians and semiaquatic reptiles that require a 

combination of wetland and terrestrial habitat for feeding, 

nesting, or hibernating may use areas that are hundreds of 

meters from the water’s edge (Buhlmann and others, 1999; 

Semlitsch and Jensen, 2001; Semlitsch and Bodie, 2003). 

Seasonal wetlands such as karst-depression wetlands are 

especially important to amphibians because their reproduction 

necessitates access to water and because many species display 

a high degree of philopatry to particular breeding sites; thus, 

the decline of amphibian species across the southeastern 

United States can be attributed in part to wetland habitat 

loss (Roberts and others, 2004). Some evidence suggests 

that differences in hydroperiod and vegetation physiognomy 

between different karst-depression wetlands drive differences 

in amphibian community structure. Liner and others (2008), 

for example, observed relatively higher species richness of 

salamanders in cypress-gum swamps with relatively long 

hydroperiods and high canopy cover, and relatively low 

species richness of amphibians with long larval periods, such 

as aquatic salamanders, in cypress savannas with shorter 

hydroperiods and less canopy cover. Thus, conservation of 

the full suite of amphibian biodiversity supported by karstdepression wetlands likely requires conservation of diverse 

habitat types afforded by these wetlands.

Some evidence exists for high invertebrate diversity in 

habitats provided by karst-depression wetlands. For example, 

Shenandoah Valley sinkhole ponds support an exceptional 

diversity of odonates, making this ecosystem type a locally 

important habitat resource for this insect group (Buhlmann 

and others, 1999). In a study of a number of diverse habitats 

at The Barrens at Arnold Air Force Base, two karst wetland 

sites—Sinking Pond and Goose Pond—had the highest species 

richness of ambrosia and bark beetles (Coleoptera: Curculionidae) and of darkling beetles (Tenebrionidae) (Grant and 

others, 2003; Wiggins and others, 2007). In the same study, 

Sinking Pond yielded a specimen of Wormaldia shawnee (a 

fingernet caddisfly), the first record of an adult in Tennessee. 

A new State record was also reported for Pycnopsyche rossi 

(a northern casemaker caddisfly) found in a grove of Nyssa 

aquatica (Wiggins and Etnier, 2001). Overall insect diversity 
118 Insular Ecosystems of the Southeastern United States

at Sinking Pond as measured by the Shannon Index was high, 

with moderate evenness (Vlach, 1999; Vlach and others, 

2010). Four species collected at Sinking Pond—Cicindela 

unipunctata (one-spotted tiger beetle), Lethe anthedon 

(northern pearly-eye), Glaucopsyche lygdamus (silvery 

blue), and Speyeria cybele (great spangled fritillary)—are 

state-listed as endangered, rare, or threatened in neighboring 

States (Alabama or North Carolina) (Vlach and others, 

2010). In a study of macroinvertebrates in karst-depression 

wetlands of the Dougherty Plain, Battle and Golladay (2001) 

found higher taxa richness in marsh habitat than in cypress 

savannas or cypress-gum swamps and suggested that a higher 

variety of food sources and habitat structures in marshes may 

have explained this, along with stress associated with low 

dissolved-oxygen conditions in swamp habitat.

One of the most prominent sets of environmental factors 

affecting animal habitat is the interaction of hydroperiod, 

timing of inundation, and inundation depth. Amphibian 

density and diversity are largely regulated by hydroperiod. If 

hydroperiod is too short, habitat becomes unavailable for part 

of the time required for egg development and larval maturation. Conversely, if hydroperiod is long enough to allow 

the establishment of predatory fish, then predation on eggs, 

juveniles, and adults may dramatically reduce amphibian 

populations (Sutter and Kral, 1994; Roberts and others, 2004). 

Based on a study of invertebrates in limesink wetlands on the 

Dougherty Plain, Golladay and others (1997) suggested that 

abnormally long inundation may have suppressed populations 

of amphipods and isopods by depriving them of summer 

refugia and may have interrupted life-cycle completion for 

cladocerans and copepods. 

For mammals and birds, forest composition and structure 

are important habitat characteristics. Species composition 

of the overstory commonly affects food availability. For 

example, oaks provide a valuable food source in the form of 

acorns for a variety of mammals and birds. Habitat structural 

complexity, determined by the vertical distribution of foliage 

layers within the canopy, is an especially important habitat 

characteristic for birds. The age and size of the largest trees 

influence availability of tree cavities for nesting. In addition, tree density directly relates to rates of fruit production 

and to shrub cover density of the understory (Roberts and 

others, 2004).

Conservation Considerations

Conservation of karst-depression wetlands is an important component of regional biodiversity conservation because 

these ecosystems provide key habitat to plant and animal 

species, including rich assemblages of aquatic vertebrates and 

invertebrates as well as plant species with disjunct distributions. Conservation of karst-depression wetlands is also 

important from an ecosystem services perspective because 

these wetlands provide a number of geologic and hydrologic 

functions, including groundwater recharge, flood peak 

attenuation, sediment retention, water-quality improvement, 

and the sequestration of carbon, metals, and environmental 

contaminants (Roberts and others, 2004; Hill and others, 

2006). Conservation of these wetlands involves addressing 

a range of interrelated threats, consideration of landscape 

and hydrologic contexts surrounding wetlands, and scientific 

attention to key knowledge gaps.

Threats to Ecosystem Integrity

In many locations, karst-depression wetlands and their 

associated upland areas have a history of anthropogenic 

damage. Commonly, karst-depression wetlands were ditched, 

drained, and (or) filled to enable conversion to agricultural 

production while surrounding upland forests were logged and 

cleared for agriculture (Hannan and Lassetter, 1982; Baskin 

and others, 1997; Buhlmann and others, 1999; Kirkman and 

others, 1999; Roberts and others, 2004). The three karstdepression wetlands described by Ellis and Chester (1989) had 

all been altered by human activities, including logging, trash 

dumping, agricultural use, and road and railroad construction. 

The vast majority of karst-depression wetlands examined by 

Homoya and Hedge (1985) in southern Indiana also were 

highly disturbed. Residential, industrial, and commercial 

development is also a primary threat to karst-depression 

wetlands (Hannan and Lassetter, 1982; Sutter and Kral, 1994; 

Buhlmann and others, 1999; Roberts and others, 2004).

In some cases, the basin geomorphology of karst 

depressions has been intentionally altered (such as by 

infilling), but in other cases the alteration has been unintentional and indirect. Erosion and sediment transport accelerated 

by construction and agriculture can increase the rate of 

sedimentation in karst-depression wetlands, which often serve 

as sediment sinks for the surrounding landscape due to their 

low topographic position. Logging activities can alter the 

evapotranspiration patterns of karst-depression wetlands and 

their adjacent uplands and also can make nearby upland soils 

more prone to erosion, leading to increased sedimentation of 

karst-depression basins (Wolfe, 1996b; Roberts and others, 

2004). If this sedimentation increase is sufficiently high it 

can, in turn, affect basin geomorphology and consequently 

hydrology (Wolfe, 1996b; Roberts and others, 2004).

Watershed-scale land-use changes can have important 

indirect effects on karst-depression wetlands through localized 

hydrologic alteration. Because these wetlands commonly 

receive surface-water inputs as runoff from surrounding 

upland areas, land-use changes that alter the imperviousness of 

the watershed or divert the flow of runoff can have profound 

effects on the hydrologic regime of karst-depression wetlands 

(Roberts and others, 2004). In some cases, these changes can 

be modeled and predicted. Using a soil-plant-atmospherewater model to simulate biological responses to altered 
Chapter H 119

hydrology under different land-management scenarios, Hill 

(2007) predicted an overall decrease in species diversity 

in the Algood wetland on the Highland Rim in Tennessee 

under scenarios of increased conversion of watershed area to 

impervious surfaces. 

In addition, threats from anthropogenic hydrologic 

alteration can be associated with groundwater changes. 

Because some types of karst-depression wetlands, such as 

compound sinks, are closely connected to groundwater, 

localized groundwater extraction that results in changes 

to the local water table has the potential to seriously alter 

wetland hydrologic regimes (Wolfe, 1996b). Because 

patterns of inundation and saturation largely determine the 

soil biogeochemistry of karst-depression wetlands based on 

seasonal patterns of oxic and anoxic conditions, anthropogenic hydrologic alteration (whether from land-use change or 

groundwater extraction) can produce a cascade of secondary 

consequences in terms of nutrient cycling, the sequestration 

of metals and carbon, and plant and animal community 

dynamics (Roberts and others, 2004).

Certain vegetation community types—such as cypress 

savannas and grass-sedge marshes within karst-depression 

wetlands of the Dougherty Plain—appear to be heavily 

dependent on disturbance from fire (Kirkman and others, 

2000). For these communities, anthropogenic fire suppression 

commonly results in hardwood establishment, and thus the 

diminution of herbaceous habitat and the degradation of 

herbaceous biodiversity (Kirkman and others, 2012). Once a 

threshold of woody dominance is surpassed, the accumulation of fire-resistant litter can allow clusters of hardwoods to 

persist even after the reintroduction of historical fire regimes 

(Martin and Kirkman, 2009).

A range of other emerging threats may affect karstdepression wetland ecosystems. Although some invasive 

exotic plant species of the surrounding uplands are excluded 

from karst-depression wetlands due to soil saturation, certain 

stress-tolerant invasives have the potential to destabilize 

communities in karst-depression wetlands, including 

Ligustrum vulgare, Lonicera japonica, and Microstegium 

vimineum (Roberts and others, 2004). Exotic insect species 

have also been noted in karst-depression wetlands (Vlach 

and others, 2010). Contamination may also threaten some 

biota. For example, Fleming and others (1984) reported 

unusually high levels of polychlorinated biphenyls in a great 

blue heron egg from a colony at Sinking Pond on the Eastern 

Highland Rim. The effects on water chemistry from acid 

precipitation have also been mentioned as a possible threat to 

amphibians and invertebrates breeding in Shenandoah Valley 

sinkhole ponds (Buhlmann and others, 1999; Downey and 

others, 1999). Additionally, land-use changes of the upland 

areas surrounding karst-depression wetlands can also affect 

wildlife through habitat fragmentation, a process known to 

affect species diversity and population genetic diversity with 

disproportionate effects on species with highly specialized 

habitat requirements (Roberts and others, 2004). 

Conservation Strategies

A number of karst-depression wetlands are being 

monitored by conservation agencies and are on publicly 

protected land. Protected or publicly owned sites in the 

Interior Low Plateaus include Arnold Air Force Base in 

the Eastern Highland Rim of Tennessee (the location of 

two National Natural Landmarks: Goose Pond and Sinking 

Pond), Chaney Lake in Warren County, Kentucky, and 

Cedar Hill Swamp in Robertson County, Tennessee (Wolfe, 

1996b; Baskin and others, 1997; Haugh, 2006; Wiggins 

and others, 2007). In the Shenandoah Valley, the Loves 

Run Pond Complex and the Maple Flats Pond Complex 

are managed by the U.S. Forest Service (Buhlmann and 

others, 1999). On the Dougherty Plain, a dense concentration of karst wetlands and associated longleaf pine and 

wiregrass are protected within Ichauway, an approximately 

29,200-acre (11,800-hectare) private reserve in Baker 

County, Georgia (Drew and others, 1998; Kirkman and 

others, 2000; Smith and others, 2006). These important sites 

notwithstanding, researchers of karst-depression wetlands 

have argued that current national wetland policy is inadequate to protect these ecosystems because of their small 

geographic size and relative isolation from surface-water 

networks (Sutter and Kral, 1994; Kirkman and others, 1999; 

Kirkman and others, 2012).

Effective conservation of karst-depression wetlands 

and the biological communities they support requires 

consideration of landscape-level processes and interactions 

with the surrounding terrestrial environment. Inadequate 

protection of terrestrial habitat for wetland-breeding amphibians is predicted to reduce amphibian biodiversity and to 

result in localized extinction for vulnerable species (Harper 

and others, 2008). As is the case with other geographically 

isolated wetlands, conservation of amphibian species that 

use karst-depression wetlands to complete their life cycle 

requires conservation of the entire species habitat, which 

includes terrestrial zones necessary for feeding, nesting, and 

overwintering. Herpetologists have argued that conservation 

of upland terrestrial habitat should be driven by an understanding of individual species’ biological requirements and 

the maximization of connectivity between wetlands, rather 

than simply based on buffer strips of an arbitrary distance 

(Semlitsch, 1998; Buhlmann and others, 1999; Semlitsch 

and Jensen, 2001; Semlitsch and Bodie, 2003). 

Additionally, biodiversity conservation in karstdepression wetlands requires preservation of the diversity of 

habitat types provided by these wetlands because wetlands 

of varying hydroperiod and vegetation physiognomy have 

been demonstrated to support different assemblages of plant 

and animal life (Buhlmann and others, 1999;Battle and 

Golladay, 2001; Liner and others, 2008). For disturbancemaintained communities, prescribed burns (and in some 

cases canopy removal) may be required to preserve herbaceous habitat (Martin, 2006; Martin and Kirkman, 2009).
120 Insular Ecosystems of the Southeastern United States

Knowledge Gaps

Effective management and conservation require additional research to fill knowledge gaps about karst-depression 

wetlands as landforms and as ecosystems. In contrast to 

larger and more geographically continuous coastal and 

alluvial wetlands, the small size and scattered distribution 

of karst-depression wetlands have limited research on the 

hydrology and biology of these ecosystems (Sutter and Kral, 

1994). Regional inventories of karst-depression wetlands 

are still incomplete and require hydrogeomorphic and 

ecological assessments in order to fully characterize the 

diversity of existing karst-depression wetlands (Kirkman 

and others, 1999).

Hydrologic regimes vary widely among karst-depression 

wetlands, and the interactions between geomorphology, hydrologic regime, and community structures remain poorly understood (Kirkman and others, 1999). Although the ecological 

importance of these hydrologic regimes is generally accepted, 

current understanding of ecohydrology in karst-depression 

wetlands is based on a small number of studies clustered 

in a few areas, with little consistency regarding objectives, 

methods, or approaches and with a tendency toward description and speculation rather than quantitative evaluation of 

hypotheses. For example, Hendricks and Goodwin (1952) 

reported weekly surface-water and groundwater levels along 

with relative abundance of plant species for 13 ponds within 

the Dougherty Plain of southwest Georgia, but did not connect 

spatial patterns in vegetation to hydrologic observations.

In particular, the dynamics of hydrologic connectivity to 

surface water and groundwater require further study in karstdepression wetlands (Kirkman and others, 2012). Baseline 

characterization of these dynamics is required in order to assess 

possible future alterations due to climate change or land-use 

change. The interaction of these dynamics with evapotranspiration and their combined effects on key components of the 

hydrologic regime (such as hydroperiod, inundation depth, and 

depth to groundwater during dry periods) require further study 

in order to predict and evaluate possible hydrologic changes. 

Conservation strategies require an understanding of these 

interacting factors. For example, while traditional conservation 

approaches that are focused on the wetland and a surrounding 

buffer area may be acceptable for wetlands with primarily 

surface-water connections, conservation of wetlands with 

extensive groundwater connection likely requires strategies 

encompassing entire aquifers (Sutter and Kral, 1994). Because 

hydrologic regime strongly regulates vegetation communities 

in karst-depression wetlands, long-term directional changes to 

key components of the hydrologic regime have the potential to 

destabilize community dynamics, emphasizing the importance 

of quantifying hydrologic indicators and their ecological 

consequences (Wolfe and others, 2004).

Additionally, gaps in knowledge about the community 

ecology of these ecosystems need to be addressed. For 

example, the dynamics of disjunct species in karst-depression 

wetlands, and particularly the direction and ecological controls 

on succession, have not been well studied (Wolfe, 1996b). 

Historical fire regimes and the role of seed banks in rare plant 

communities are not fully understood in these ecosystems 

(Sutter and Kral, 1994). Quantification is needed for invasive 

exotic species, as well as improved understanding of the 

conditions facilitating invasions (Kirkman and others, 1999). 

Restoration of degraded karst-depression wetlands requires 

the identification of reference sites and the determination 

of restoration endpoints, both of which can be difficult in 

regions where anthropogenic alteration of wetlands has been 

widespread (Kirkman and others, 1999; Kirkman and others, 

2000; Martin, 2006).

Conservation of rare and vulnerable species within 

karst-depression wetlands requires improved understanding 

of the particular life cycles and habitat requirements of each 

species (Semlitsch and Bodie, 2003). For example, the stresstolerance thresholds of narrowly endemic plant species require 

further study to enable effective conservation (Knox, 1997). 

Identification of source and sink populations of key amphibians within karst wetland complexes, as well as quantification 

of upland buffer habitat requirements, is required to inform 

conservation approaches for karst wetland fauna (Sutter and 

Kral, 1994; Buhlmann and others, 1999). Additional taxa 

that are dependent upon karst-depression wetland habitat—

especially invertebrates, algae, and fungi—are likely waiting 

to be discovered (Kirkman and others, 1999).

Possible Ecological Effects from Climate Change

Global climate change has the potential to substantially 

alter karst-depression wetlands, especially through its 

effects on precipitation and evapotranspiration patterns and 

through consequent effects on groundwater levels. Wolfe 

(1996b) argued that internal relief and interaction with the 

groundwater system are likely to affect the sensitivity to 

environmental change of karst-depression wetlands. Broad, 

shallow, perched systems may be relatively resilient because 

water depths and internal drainage are respectively constrained 

by topography and impermeable substrate. Conversely, narrow, 

deep depressions that are closely coupled to the groundwater 

system may be vulnerable to prolonged, deep ponding should 

groundwater levels increase or to wetland degradation through 

hydroperiod reduction should groundwater levels decrease 

(Wolfe, 1996b).

Such ecohydrologic vulnerability is exemplified by the 

effects of climate change on tree regeneration and mortality in 

Sinking Pond, an 87-acre (approximately 35-hectare) Highland 

Rim compound sink with a depth of approximately 11.5 feet 

(about 3.5 meters). In a study of the population dynamics of 

Quercus lyrata in Sinking Pond, McCarthy and Evans (2000) 

and Wolfe and others (2004) provided evidence that these 

changes may already be underway. This research demonstrated 

a suppression of recruitment possibly constituting regeneration 

failure: Q. lyrata saplings and young adults were found to be 

generally absent in areas of ponding depth greater than about 
Chapter H 121

1.5 feet (roughly 0.5 meter), despite concentrations of seedlings 

and mature trees in these areas, and were instead confined to a 

small area of relatively shallow ponding depth. Dendrochronological analysis showed a marked suppression of recruitment in 

moderately to deeply flooded areas after 1970, corresponding 

to an increase in ponding durations across ponding-depth 

classes and to increased regional and local precipitation after 

1970. Wolfe and others (2004) evaluated the hypothesis that 

these hydrologic and ecologic changes were linked to localized 

human activities within the Sinking Pond basin, such as direct 

modifications to drainage, land-use changes within the basin, 

or impoundment of surrounding catchments (see appendix B 

in Wolfe and others, 2004); however, evidence was lacking to 

support any of these explanations. Thus, the observed recruitment failure was attributed to an increase in ponding duration 

over time resulting from increased regional precipitation. This 

adverse effect from climate change will likely affect not only 

the population of Q. lyrata in Sinking Pond but also other 

members of the pond’s globally rare community.

References

Baskin, J., Chester, E., and Baskin, C., 1997, Special paper—

Forest vegetation of the Kentucky karst plain (Kentucky 

and Tennessee)—Review and synthesis: Journal of the 

Torrey Botanical Society, v. 124, no. 4, p. 322–335.

Battle, J. and Golladay, S., 2001, Water quality and macroinvertebrate assemblages in three types of seasonally 

inundated limesink wetlands in southwest Georgia: 

Journal of Freshwater Ecology, v. 16, no. 2, p. 189–207.

Beck, B., and Arden, D., 1984, Karst hydrogeology and 

geomorphology of the Dougherty Plain, southwest Georgia, 

guidebook number 26: Tallahassee, Fla., Southeastern 

Geological Society, 63 p.

Brahana, J., and Bradley, M., 1986, Preliminary delineation 

and description of the regional aquifers of Tennessee—

Highland Rim aquifer system: U.S. Geological Survey 

Water-Resources Investigations Report 82–4054, 38 p.

Buhlmann, K., Mitchell, J., and Smith, L., 1999, Descriptive 

ecology of the Shenandoah Valley sinkhole pond system in 

Virginia: Banisteria, v. 13, p. 23–51.

Chester, E., and Ellis, W., 1989, Plant communities of northwestern Middle Tennessee: Journal of the Tennessee 

Academy of Science, v. 64, no. 3, p. 75–78.

Cowardin, L., Carter, V., Golet, F., and LaRoe, E., 1979, 

Classification of wetlands and deepwater habitats of the 

United States: U.S. Department of the Interior, Fish and 

Wildlife Service, FWS/OBS-79/31.

Cranfill, R., 1991, Flora of Hardin County, Kentucky: 

Castanea, v. 56, no. 4, p. 228–267.

Delcourt, H., 1979, Late quaternary vegetation history of the 

eastern Highland Rim and adjacent Cumberland Plateau 

of Tennessee: Ecological Monographs, v. 49, no. 3, 

p. 255–280.

DeSelm, H., 1973, Six potential National Natural Landmarks 

in Tennessee—Report prepared for the U.S. National Park 

Service, Southeast Region, Atlanta, Ga.: Knoxville, Tenn., 

University of Tennessee, NPS--CS5000 31582.

DeSelm, H., 1990, Flora and vegetation of some barrens of the 

eastern Highland Rim of Tennessee: Castanea, v. 55, no. 3, 

p. 187–206.

Downey, D.M., Wirtz, S., Kruer, K.R., and Douglas, S.P., 

1999, Water chemistry assessment of the Shenandoah Valley 

sinkhole ponds in Virginia: Banisteria, no. 13, p. 53–65.

Drew, M.B., Kirkman, L.K., and Gholson, A.K., 1998, 

The vascular flora of Ichauway, Baker County, Georgia—

A remnant longleaf pine/wiregrass ecosystem: Castanea, 

v. 63, no. 1, p. 1–24.

Edwards, A., and Weakley, A., 2001, Population biology and 

management of rare plants in depression wetlands of the 

southeastern Coastal Plain, USA: Natural Areas Journal, 

v. 21, no. 1, p. 12–35.

Ellis, W., and Chester, E., 1989, Upland swamps of the Highland Rim of Tennessee: Journal of the Tennessee Academy 

of Science, v. 64, no. 3, p. 97–101.

Fenneman, N., 1938, Physiography of the eastern United States:

New York, McGraw-Hill, 714 p.

Fenneman, N., and Johnson, D., 1946, Physical divisions of 

the United States: U.S. Geological Survey map prepared 

in cooperation with the Physiographic Commission, 

scale 1:7,000,000. 

Fleming, G., and Patterson, K., 2012, Natural communities 

of Virginia—Ecological groups and community types: 

Richmond, Va., Virginia Department of Conservation and 

Recreation, Division of Natural Heritage, 36 p.

Fleming, G., Patterson, K., Taverna, K., and Couling, P., 

2012, The natural communities of Virginia—Classification 

of ecological community groups, second approximation, 

version 2.5: Richmond, Va., Virginia Department of Conservation and Recreation, Division of Natural Heritage.

Fleming, W., Pullin, B., and Swineford, D., 1984, Population trends and environmental contaminants in herons in 

the Tennessee Valley, 1980–81: Colonial Waterbirds, v. 7, 

p. 63–73.

Fleming, G., and Van Alstine, N., 1999, Plant communities 

and floristic features of sinkhole ponds and seepage 

wetlands in southeastern Augusta County, Virginia: 

Banisteria, v. 13, p. 67–94.
122 Insular Ecosystems of the Southeastern United States

Folkerts, G., 1997, Citronelle ponds—Little-known wetlands 

of the central Gulf Coastal Plain, USA: Natural Areas 

Journal, v. 17, p. 6–16.

Ford, D., and Williams, P., 1989, Karst geomorphology and 

hydrology: London, Unwin Hyman, Ltd., 601 p.

Franklin, M.A., and Finnegan, J.T., 2006, Natural Heritage 

Program list of rare plant species of North Carolina 2006: 

Raleigh, N.C., North Carolina Natural Heritage Program, 

Office of Conservation and Community.

Golladay, S., Entrekin, S., and Taylor, B., 1999, Forested 

limesink wetlands of southwest Georgia—Invertebrate 

habitat and hydrologic variation, in Batzer, D., Rader, R., 

and Wissinger, S., eds., Invertebrates in freshwater wetlands 

of North America: John Wiley & Sons, p. 197–216.

Golladay, S.W., Taylor, B.W., and Palik, B.J., 1997, Invertebrate communities of forested limesink wetlands in southwest Georgia, USA—Habitat use and influence of extended 

inundation: Wetlands, v. 17, no. 3, p. 383–393.

Grant, J., Mayor, A., Lambdin, P., and Wiggins, G., 2003, New 

species records and incidence of bark beetles and ambrosia 

beetles (Coleoptera: Scolytidae) from the barrens of middle 

Tennessee, USA: Natural Areas Journal, v. 22, p. 278–283.

Greear, P., 1967, Composition, diversity, and structure of the 

vegetation of some natural ponds in northwest Georgia: 

University of Georgia, Ph.D. dissertation.

Hannan, R., and Lassetter, J., 1982, The vascular flora of the 

Brodhead Swamp Forest, Rockcastle County, Kentucky: 

Transactions of the Kentucky Academy of Sciences, v. 43, 

no. 1–2, p. 43–49.

Harper, E.B., Rittenhouse, T.A.G., and Semlitsch, R.D., 2008, 

Demographic consequences of terrestrial habitat loss for 

pool-breeding amphibians—Predicting extinction risks 

associated with inadequate size of buffer zones: Conservation Biology, v. 22, no. 5, p. 1205–1215.

Haugh, C.J., 2006, Hydrogeology and simulation of groundwater flow at Arnold Air Force Base, Coffee and Franklin 

Counties, Tennessee—2002 update: U.S. Geological Survey 

Scientific Investigations Report 2006–5157, 16 p.

Hendricks, E., and Goodwin, M., 1952, Water-level fluctuations in limestone sinks in southwestern Georgia: Washington, D.C., U.S. Department of the Interior, Geological 

Survey Water-Supply Paper 1110-E.

Hess, J., Wells, S., Quinlan, J., and White, W., 1989, Hydrogeology of the south-central Kentucky karst: Karst 

Hydrology, p. 15–63.

Hicks, D., Krause, R., and Clarke, J., 1981, Geohydrology of 

the Albany area, Georgia: Atlanta, Ga., Georgia Department 

of Natural Resources, Environmental Protection Division, 

Georgia Geologic Survey.

Hill, A., 2007, Monitoring and modeling the hydrology of a 

forested sinkhole wetland on the Tennessee Highland Rim: 

Tennessee Technological University, Ph.D. dissertation, 

212 p.

Hill, A., and Neary, V., 2007, Estimating evapotranspiration 

and seepage for a sinkhole wetland from diurnal surface 

water cycles: Journal of the American Water Resources 

Association, v. 43, no. 6, p. 1373–1382.

Hill, A., and Neary, V., 2012, Annual water budgets for a 

seasonally inundated sinkhole wetland: Wetlands, v. 32, 

no. 5, p. 963–974.

Hill, A., Neary, V., and Morgan, K., 2006, Hydrologic 

modeling as a development tool for HGM functional 

assessment models: Wetlands, v. 26, no. 1, p. 161–180.

Hinkle, K.R., and Sterrett, R.M., 1976, Rockingham County 

groundwater—Present conditions and prospects: Richmond, 

Va., Commonwealth of Virginia State Water Control Board, 

Bureau of Water Control Management.

Hinkle, K.R., and Sterrett, R.M., 1978, Groundwater resources 

of Augusta County, Virginia: Richmond, Va., Commonwealth of Virginia State Water Control Board.

Hollyday, E.F., and Hileman, G.E., 1996, Hydrogeologic 

terranes and potential yield of water to wells in the Valley 

and Ridge Physiographic Province in the eastern and southeastern United States: U.S. Geological Survey Professional 

Paper 1422–C, 38 p.

Homoya, M., and Hedge, C., 1985, The upland sinkhole 

swamps and ponds of Harrison County, Indiana: Proceedings of the Indiana Academy of Science, v. 92, p. 383–388.

Johnson, G., Zimmerman, T., Lindsey, B., and Gross, E., 

2011, Factors affecting groundwater quality in the Valley 

and Ridge aquifers, eastern United States, 1993–2002: 

U.S. Geological Survey Scientific Investigations 

Report 2011–5115, 70 p.

Karl, T.R., and Knight, R.W., 1998, Secular trends of precipitation amount, frequency, and intensity in the United States: 

Bulletin of the American Meteorological Society, v. 79, 

no. 2, p. 231–241.

Kirkman, L.K., Goebel, P.C., West, L., Drew, M.B., and 

Palik, B.J., 2000, Depressional wetland vegetation types—

A question of plant community development: Wetlands, 

v. 20, no. 2, p. 373–385.
Chapter H 123

Kirkman, L., Golladay, S., Laclaire, L., and Sutter, R., 1999, 

Biodiversity in southeastern, seasonally ponded, isolated 

wetlands—Management and policy perspectives for 

research and conservation: Journal of the North American 

Benthological Society, v. 18, no. 4, p. 553–562.

Kirkman, L., Smith, L., and Golladay, S., 2012, Southeastern 

depressional wetlands, in Batzer, D., and Baldwin, A., eds., 

Wetland habitats of North America—Ecology and conservation concerns: Berkeley, Calif., University of California 

Press, p. 203–215.

Knox, J., 1997, A nine year demographic study of Helenium virginicum (Asteraceae), a narrow endemic seasonal 

wetland plant: Journal of the Torrey Botanical Society, 

v. 124, no. 3, p. 236–245.

Liner, A.E., Smith, L.L., Golladay, S.W., Castleberry, S.B., 

and Gibbons, J.W., 2008, Amphibian distributions within 

three types of isolated wetlands in southwest Georgia: The 

American Midland Naturalist, v. 160, no. 1, p. 69–81.

Love, T., Williams, L., Proffitt, W., Epley, I., and Elder, 

J., 1956, Soil survey of Coffee County, Tennessee: 

U.S. Department of Agriculture, Soil Conservation Service, 

Series 1956, no. 5.

Martin, K., 2006, Restoration of fire-dependent depression 

wetlands: University of Georgia, Masters thesis.

Martin, K., and Kirkman, L., 2009, Management of ecological thresholds to re-establish disturbance-maintained 

herbaceous wetlands of the south-eastern USA: Journal 

of Applied Ecology, v. 46, no. 4, p. 906–914.

McCabe, G.J., and Wolock, D., 2002, A step increase in 

streamflow in the conterminous United States: Geophysical 

Research Letters, v. 29, no. 24, p. 2185.

McCarthy, S., and Evans, J., 2000, Population dynamics of 

overcup oak (Quercus lyrata) in a seasonally flooded karst 

depression: Journal of the Torrey Botanical Society, v. 127, 

no. 1, p. 9–18.

McKinney, L., 1989, Vegetation of the Eastern Highland 

Rim of Tennessee: Journal of the Tennessee Academy of 

Science, v. 64, no. 3, p. 145–147.

Miller, B., 1995, An investigation and assessment of rare, 

threatened, and endangered fauna and their habitats on 

Arnold Air Force Base, Coffee and Franklin Counties, 

Tennessee—Amphibian and reptile survey: Murfreesboro, 

Tenn., Middle Tennessee State University.

Miller, B.T., Lamb, J.W., and Miller, J.L., 2005, The herpetofauna of Arnold Air Force Base in the Barrens of southcentral Tennessee: Southeastern Naturalist, v. 4, no. 1, 

p. 51–62.

Miller, J., 1986, Hydrogeologic framework of the Floridan 

aquifer system in Florida and in parts of Georgia, Alabama, 

and South Carolina: U.S. Geological Survey Professional 

Paper 1403–B, 91 p.

Moler, P., and Franz, R., 1987, Wildlife values of small, 

isolated wetlands in the southeastern Coastal Plain, in

Proceedings of the Third Southeastern Nongame and 

Endangered Wildlife Symposium: Atlanta, Ga., Georgia 

Department of Natural Resources, p. 234–241.

NatureServe, 2015, NatureServe Explorer—An online 

encyclopedia of life [web application], ver. 7.1, accessed 

October 7, 2015, at http://www.natureserve.org/explorer.

Niemiller, M., 2005, The herpetofauna of the upper Duck 

River watershed in Coffee County, Tennessee: Journal of 

the Tennessee Academy of Science, v. 80, no. 1, p. 6–12.

Nuttle, W., 1997, Measurement of wetland hydroperiod using 

harmonic analysis: Wetlands, v. 17, no. 1, p. 82–89.

O’Driscoll, M.A., and Parizek, R.R., 2003, The hydrologic 

catchment area of a chain of karst wetlands in central 

Pennsylvania, USA: Wetlands, v. 23, no. 1, p. 171–179.

O’Driscoll, M.A., and Parizek, R.R., 2008, Geological 

controls on seasonal-pool hydroperiod in a karst setting: 

Wetlands, v. 28, no. 4, p. 1004–1017.

Opsahl, S.P., 2005, Organic carbon composition and oxygen 

metabolism across a gradient of seasonally inundated 

limesink and riparian wetlands in the southeast Coastal 

Plain, USA: Biogeochemistry, v. 76, no. 1, p. 47–68.

Palik, B.J., Goebel, P.C., Kirkman, L.K., and West, L., 

2000, Using landscape hierarchies to guide restoration of 

disturbed ecosystems: Ecological Applications, v. 10, no. 1, 

p. 189–202.

Patterson, W., 1989, Vegetation and soils of the Sinking Pond 

area, Coffee County, Tennessee. University of Tennessee, 

Thesis.

Pullin, P., 1990, Size and trends of wading bird populations 

in Tennessee during 1977–1988: The Migrant, v. 61, no. 4, 

p. 95–104.

Reed, J., Bryant, B., and Hack, J., 1963, Origin of some intermittent ponds on quartzite ridges in western North Carolina: 

The Geological Society of America Bulletin, v. 74, no. 9, 

p. 1183–1188.

Roberts, T., Morgan, K., Hill, A., Cripps, R., and Neary, V., 

2004, Hydrogeomorphic classification and assessment of 

depression and flat wetlands on the Tennessee Highland 

Rim: Nashville, Tenn., State of Tennessee Department 

of Transportation.
124 Insular Ecosystems of the Southeastern United States

Ryder, R.T., 1987, Oil and gas resources of the Cincinnati Arch, Ohio, Indiana, Kentucky, and Tennessee: 

U.S. Geological Survey Open-File Report 87–450Y, 30 p.

Semlitsch, R., 1998, Biological delineation of terrestrial 

buffer zones for pond breeding salamanders: Conservation 

Biology, v. 12, no. 5, p. 1113–1119.

Semlitsch, R.D., and Bodie, J.R., 2003, Biological criteria 

for buffer zones around wetlands and riparian habitats for 

amphibians and reptiles: Conservation Biology, v. 17, 

no. 5, p. 1219–1228.

Semlitsch, R.D., and Jensen, J.B., 2001, Core habitat, not 

buffer zone: National Wetlands Newsletter, v. 23, no. 4, 

p. 4–6.

Smith, L.L., Steen, D.A., Stober, J.M., Freeman, M.C., 

Steve, W., Conner, L.M., and Cochrane, J., 2006, The 

vertebrate fauna of Ichauway, Baker County, GA: 

Southeastern Naturalist, v. 5, no. 4, p. 599–620.

Sorrie, B., 2012, Identification, distribution, and habitat of 

needle-leaved Hypericum (Hypericaceae) in the southeastern United States: Phytoneuron, v. 76, p. 1–14.

Sorrie, B., and Weakley, A., 2001, Coastal Plain vascular 

plant endemics—Phytogeographic patterns: Castanea, 

v. 66, no. 1, p. 50–82.

Sutter, R., and Kral, R., 1994, The ecology, status, and conservation of two non-alluvial wetland communities in the south 

Atlantic and eastern Gulf Coastal Plain, USA: Biological 

Conservation, v. 68, p. 235–243.

Torak, L., Davis, G., Strain, G., and Herndon, J., 1993, 

Geohydrology and evaluation of water-resource potential 

of the Upper Floridan Aquifer in the Albany area, southwestern Georgia: U.S. Geological Survey Water-Supply 

Paper 2391, 59 p. 

Townsend, J., 2015, Natural heritage resources of Virginia—

Rare plants: Richmond, Va., Virginia Department of 

Conservation and Recreation, Division of Natural Heritage, 

Natural Heritage Technical Report 15-10, 60 p.

Unterreiner, G., 2006, Hydrogeology of Harrison County, 

Indiana: Indianapolis, Ind., State of Indiana Department 

of Natural Resources, Division of Water, Bulletin 40.

U.S. Environmental Protection Agency, 2013, Level III and IV 

ecoregions of the continental United States: U.S. Environmental Protection Agency, National Health and Environmental Effects Research Laboratory, scale 1:3,000,000, 

accessed October 28, 2015, at http://www.epa.gov/wed/

pages/ecoregions/level_iii_iv.htm.

Vlach, J., 1999, An assessment of arthropod diversity using 

nine collection methods at Sinking Pond—A registered 

natural landmark in Coffee County, Tennessee: The University of Tennessee, Masters Thesis, 120 p.

Vlach, J., Lambdin, P., Dilling, C., and Grant, J., 2010, 

Diversity of the insect fauna within the unique sinking 

pond habitat in Middle Tennessee: Journal of the Tennessee 

Academy of Science, v. 85, no. 3, p. 62–86.

White, W.B., 1988, Geomorphology and hydrology of karst 

terrains: New York, Oxford University Press, 464 p.

Whittecar, G.R., and Lawrence, J.R., 1999, Hydrology and 

geomorphology of Green Pond—A high-elevation depressional wetland in the Blue Ridge of Virginia: Banisteria, 

v. 13, p. 149–159.

Wiggins, G., and Etnier, D., 2001, New Tennessee records for 

Wormaldia shawnee, Oligostomis ocelligera, Oligostomis 

pardalis, and Pycnopsyche rossi (Trichoptera): Entomological News, v. 112, no. 3, p. 187–190.

Wiggins, G.J., Grant, J.F., and Lambdin, P.L., 2007, Diversity 

of darkling beetles (Coleoptera: Tenebrionidae) from Arnold 

Air Force Base in the Barrens of the Eastern Highland Rim, 

Tennessee: Natural Areas Journal, v. 27, no. 1, p. 66 –71.

Wolfe, W.J., 1996a, Hydrology and tree-distribution patterns 

of karst wetlands at Arnold Engineering Development 

Center, Tennessee: U.S. Geological Survey Water-Resouces 

Investigations Report 96 – 4277, 46 p.

Wolfe, W.J., 1996b, Karst wetlands of the Barrens—

Geomorphic control of wetland hydrology and plant 

distribution on the Highland Rim of Tennessee: Clark 

University, Ph.D. Dissertation, 147 p.

Wolfe, W.J., Evans, J., McCarthy, S., Gain, W., and Bryan, B., 

2004, Tree-regeneration and mortality patterns and hydrologic change in a forested karst wetland—Sinking Pond, 

Arnold Air Force Base, Tennessee: U.S. Geological Survey 

Water-Resources Investigations Report 03– 4217, 62 p. 

Wolfe, W.J., Haugh, C.J., Webbers, A., and Diehl, T.H., 1997, 

Preliminary conceptual models of the occurrence, fate, 

and transport of chlorinated solvents in karst regions of 

Tennessee: U.S. Geological Survey Water-Resources 

Investigations Report 97– 4097, 80 p.
Chapter I. Riverscour Ecosystems

Introduction

Riverscour ecosystems are a subset of a larger category 

of riparian corridors, an ecosystem category known to support 

unusually high biodiversity, especially for vascular plants, 

in many different environmental contexts and on several 

continents (Naiman and others, 1993). The high biodiversity 

of riparian corridors is maintained in part through disturbance 

regimes—which can include floods, fire, landslides, channel 

migration and torrents of debris—as well as by micro-scale 

heterogeneity of habitat characteristics such as topographic 

position, microclimate, and soil moisture levels (Naiman 

and others, 1993; Bailey and Coe, 2001; Rood and others, 

2007). Riverscour ecosystems generally form within or near 

the channel shelf along high-gradient sections of medium- to 

large-order streams (fig. I–1). Riverscour ecosystems contain a 

number of diverse community types. Community composition 

is generally determined by the dynamic integration of factors 

such as biogeographic location, stream channel morphology, 

stream hydrology, gradient, sediment transport patterns, 

substrate chemistry, and habitat requirements of individual 

species (Hupp, 1983). The primary trait uniting the diverse 

communities in riverscour contexts is that ecological integrity 

is maintained largely by disturbance regimes based on 

periodic flood scouring. Following storm events, high-energy, 

sediment-laden flow redistributes stream sediments and 

physically damages riparian vegetation through mechanical 

abrasion (Ogle, 1992; Bailey and Coe, 2001; Murdock and 

others, 2007; Vanderhorst and others, 2007; Wolfe and others, 

2007). This disturbance regime effectively excludes scourintolerant upland vegetation and thus maintains habitat for 

scour-adapted plants. 

Figure I–1. Rough Shoals, Big South Fork Cumberland River, 

Big South Fork National River and Recreation Area, Tennessee. 

Photograph by Nora Murdock, National Park Service.
126 Insular Ecosystems of the Southeastern United States

Geographic Range

Riverscour ecosystems are present in many geographic 

areas of the United States. Within the southeastern 

United States, various types of riverscour ecosystems have 

been described in several physiographic provinces (Fenneman 

and Johnson, 1946), including the Appalachian Plateaus, the 

Blue Ridge, the Interior Low Plateaus, the Valley and Ridge, 

and the Piedmont (Ogle, 1992; Vanderhorst and others, 2007, 

2010). Riverscour habitats have been characterized and 

endemic species have been documented along several dozen 

streams (table I–1). Prominent research sites in riverscour 

ecosystems and their locations along streams are shown in 

figure I–2. Outside the southeastern United States, riverscour 

ecosystems have been described in a number of other locations 

in North America, including New England, the northern midAtlantic States, Alaska, and the Rocky Mountains (Westervelt 

and others, 2006; Rood and others, 2007; Mouw, 2009).

Physical Geography

Riverscour ecosystems may occupy discrete sites and 

(or) continuous linear zones along high-gradient sections of 

second- and higher-order streams (Ogle, 1992; Vanderhorst 

and others, 2007; Wolfe and others, 2007). Ecosystem 

development can occur on a variety of substrates, including 

bedrock (often partially exposed, with areas of thin soil or 

sediment) and aggregations of alluvial substrates ranging in 

size from boulders and cobble to gravel and sand. These rock 

particles may enter stream channels via mass wasting from 

cliffs that form the walls of river gorges or via debris flows 

down steep tributaries (Wolfe and others, 2007). Accretion of 

alluvial substrates into shoals or bars typically takes place in 

relatively stable positions where sediment-laden flow loses 

power, such as channel bends, major confluences, and zones 

of deflected flow from obstructions such as large non-alluvial 

boulders (Wolfe and others, 2007). Once alluvial bars begin 

Table I–1. Selected streams supporting riverscour ecosystems within the southeastern United States. 

[Stream names indicate rivers unless specified as creeks]

State Streams References

West Virginia Bluestone Buckhannon Cheat Ogle (1991), Suiter and Evans (1999), 

Mitchem (2004), Byers and others 

(2007), Vanderhorst and others (2007, 

2010), Wells (2012)

Gauley Greenbrier Meadow

Monongahela New Potomac

Shenandoah Tygart Valley

Pennsylvania Delaware Youghiogheny Ogle (1991), Grund and Zimmerman 

(2009), Zimmerman (2011)

Virginia Guest James Maury Hupp (1983), Ogle (1991), Belden and 

others (2003), Steury and others 

(2008), Fleming and Patterson (2012)

New Passage Creek Potomac

Pound Rappahannock Roanoke

Russell Fork Shenandoah

Maryland Monocacy Potomac Lea (2000), Vanderhorst and others 

(2007), Wells (2012)

North Carolina Cane Cheoah Hominy Creek Ogle (1991), Hutchens and others (2009), 

Little Tennessee New Nolichucky Schafale (2012)

Roanoke South Toe Yadkin

Tennessee Abrams Creek Cane Creek Clear Creek Schmalzer (1989), Ogle (1991), Pyne 

and Withers (1996), Major and others 

(1999), Bailey and Coe (2001), Wolfe 

and others (2007), U.S. Fish and Wildlife Service (2011), Floden (2012)

Clear Fork Clifty Creek Cumberland

Daddys Creek Duck Hiwassee

Little Little Tennessee Nolichucky

Obed Ocoee White Oak Creek

Kentucky Clear Fork Cumberland Green Meijer (1976), Medley and Wofford 

(1980), Ogle (1991, 1992) Pyne 

and Withers (1996), Bailey and Coe 

(2001), Reed (2003), Taylor (2003)

Kentucky Kinniconick Creek New

Red Rockcastle Sinking Creek

Georgia Chattahoochee Bear Creek Rock Creek Ogle (1991)

Alabama Little Black Warrior 

(Locust Fork)

Black Warrior

(Mulberry Fork)

Ogle (1991), Rinehart (2008)

Cahaba Creek Cypress Creek
Figure I–2. Selected locations of riverscour ecosystems in the southeastern United States. 

3

2

1

6

5

7

8

4

11

9

10

90° 85° 80° 75°

40°

35°

30°

25°

Selected streams in the

 southeastern United States

Tennessee R

Alabama

River

Prominent research location for

 riverscour ecosystems and identifier

Figure I-2

0 100 200 300 400 KILOMETERS

0 100 200 300 400 MILES

Base modifed from U.S. Department of Commerce,

Census Bureau, 2015, Mercator Auxiliary Sphere Projection

Horizontal datum is World Geodedic System of 1984

EXPLANATION

Physiographic province—From 

 Fenneman and Johnson (1946)

Appalachian Plateaus

Piedmont

Interior Low Plateaus

Valley and Ridge

Blue Ridge

Cumberland River

Ohio River

Potomac

River

1 Shenandoah River (Vanderhorst, 2000)

2 Potomac River (Steury and others, 2008)

3 Gauley River (Vanderhorst and others, 2010)

4 New River (for example, Suiter and Evans, 1999; 

 Vanderhorst and others, 2007)

5 and 6 Red and Rockcastle Rivers (Reed, 2003; Taylor, 2003)

7 Big South Fork Cumberland River (for example,

 Shaw and Wofford, 2003; Murdock and others, 2007)

8 Obed River (for example, Schmalzer, 1989;

 Wolfe and others, 2007)

9 and 10 Ocoee and Hiwassee Rivers (Major and others, 1999)

11 Little River (Rinehart, 2008)

TENNESSEE

ALABAMA

GEORGIA

SOUTH

CAROLINA

NORTH CAROLINA

WEST

VIRGINIA

OHIO

PENNSYLVANIA

ILLINOIS INDIANA

MISSOURI

ARKANSAS

VIRGINIA

FLORIDA

KENTUCKY

NEW

JERSEY

DELAWARE

MARYLAND

MISSISSIPPI

LOUISIANA

to form, their structural stability may be enhanced by the 

imbricated (diagonally stacked) arrangement of larger boulders and by the protection from scouring provided by these 

boulders to interstitial deposits of cobbles, gravel, sand, and 

silt (Vanderhorst and others, 2007; Wolfe and others, 2007). 

Once vegetation begins to colonize sites of alluvial deposition, a positive feedback loop can begin whereby plant roots 

stabilize the substrate and plant shoots filter and trap sediment, 

promoting further deposition of fine-grained suspended 

particles (Mouw, 2009).

Chapter I 127
128 Insular Ecosystems of the Southeastern United States

Stress and Disturbance Regimes

The primary component of the abiotic stress regime for 

riverscour ecosystems is disturbance, namely the mechanical 

scouring from periodic high-energy floods (Vanderhorst 

and others, 2007; Wolfe and others, 2007). Depending on 

factors such as topography and basin substrate permeability, 

streamflow can respond rapidly to storm events in some cases. 

For example, in the Obed River and in the Big South Fork 

Cumberland River, flow has been documented to increase from 

100 cubic feet per second (about 3 cubic meters per second) 

to 6,000 cubic feet per second (about 170 cubic meters per 

second) in a matter of hours following storm events, creating 

powerful flood-scouring effects on the entire riparian corridor 

(Murdock and others, 2007). High-energy flow of this type can 

be powerful enough to redistribute boulders and finer materials 

within the channel. For example, along some sections of the 

Obed River in Tennessee, hydraulic forces sufficient to transport 3-foot (approximately 1-meter) boulders along channel 

beds probably occur every few years (Wolfe and others, 2007). 

Different magnitudes and intensities of floods produce 

different types of disturbances that affect riverscour communities. Medium-intensity floods may uproot individual trees 

but leave most other vegetation relatively intact, whereas 

high-intensity floods are capable of removing entire patches 

of vegetation and substantially restructuring alluvial 

substrates, resulting in the periodic redistribution of sediment 

and reconfigurations of channel morphology (Naiman and 

others, 1993; Bailey and Coe, 2001; Vanderhorst and others, 

2007). The magnitude of the effects of disturbance regime on 

riverscour communities (frequency and depth of inundation, 

frequency and intensity of scouring) is generally proportional 

to proximity to the active channel, such that disturbance forms 

a gradient from the channel interior to the riparian forest edge 

(Hupp, 1983). Topographic variation within channels also 

affects scouring intensity. For example, island heads are generally more exposed to scouring, whereas areas immediately 

downstream of large boulders and other obstacles are more 

sheltered from the action of high-energy flows (Wolfe and 

others, 2007). At higher latitudes and elevations, winter ice 

scouring can be an additional disturbance factor (Westervelt 

and others, 2006; Byers and others, 2007; Rood and others, 

2007; Mouw, 2009).

Disturbance from periodic scouring creates stress on 

vegetation in several ways. Direct effects include uprooting 

and damage to aboveground parts of plants. Scouring may 

also remove accumulated soil or sediment, resulting in thin 

soils. Soil shallowness in turn can limit soil water and nutrient 

availability, a set of factors that can be considered secondary 

components of the stress regime (Mitchem, 2004). Other stress 

factors include prolonged saturation and inundation for some 

microhabitats and, for others, xeric soil conditions caused by 

excessive drainage of coarse-textured substrates combined 

with moderate to full insolation from lack of canopy coverage 

(Vanderhorst and others, 2007; Wolfe and others, 2007). 

Soil water availability can be highly variable spatially and 

temporally, such as in zones that are seasonally inundated in 

winter and spring and xeric in summer except for occasional 

flash floods (Bailey and Coe, 2001). Figure I–3 shows an 

example of riverscour habitat containing zones of exposed 

bedrock and standing water within topographic depressions. 

The combination of thin to absent soil, widely fluctuating 

hydrologic conditions, and high insolation is a set of stress 

regime components that some riverscour ecosystems share 

with other glade and outcrop ecosystems (Mitchem, 2004).

To maintain viable populations under the stress and 

disturbance regimes of the most heavily scoured zones of 

river channels, plant species commonly display specialized 

adaptations. These adaptations include the ability to recolonize rapidly following disturbance events, as well as dispersal 

methods that allow periodic scouring to spread seeds and 

vegetative propagules downstream (Bartgis, 1997; Bailey and 

Coe, 2001; Reed, 2003; Taylor, 2003; Wells, 2012). Tolerance 

of drought and (or) inundation are common adaptations, as are 

deep and strong anchor roots to minimize dislocation during 

floods (Ogle, 1991; Bailey and Coe, 2001). One interesting 

example is Podostemum ceratophyllum, a member of a family 

of tropical waterfall plants, which occupies riverscour habitats 

on the Red River, Little Tennessee River, and Big South 

Fork Cumberland River and other streams. This species 

uses specially adapted thalli to attach to rock surfaces with 

suction discs and glue (Meijer, 1976; Hutchens and others, 

2009). A number of other riverscour-adapted plants die back 

in the fall and persist through the seasons with greatest flood 

likelihood (winter and early spring) as perennating rootstocks 

(Hupp, 1983).

This disturbance regime from flood scouring and the 

associated abiotic stress regime effectively excludes most 

mesic upland species, especially large trees (Vanderhorst and 

others, 2007; Wolfe and others, 2007). For riverscour-adapted 

plant taxa, tolerance of this stress and disturbance regime 

provides a competitive advantage over upland species as 

long as the disturbance regime remains active; without such 

a regime, these species, including many rare and endemic 

species, would likely be displaced by mesic competitors as 

part of the succession process (Schmalzer, 1989; Bailey and 

Coe, 2001; Wolfe and others, 2007). Because some riverscour 

species are shade-intolerant (Vanderhorst, 2000; Vanderhorst 

and others, 2007), overshading may be an important 

mechanism by which competitive displacement would occur 

in the absence of the flood-scour regime. Thus, a disturbance 

regime predicated on periodic inundation and flood scouring, 

which maintains a state of disequilibrium in riverscour 

ecosystems such that climatic or edaphic climax is rarely or 

never achieved, is critical for the maintenance of ecosystem 

integrity (Hupp, 1983). 

The disturbance regime of riverscour ecosystems 

illustrates a broader concept linking high biodiversity with 

nonequilibrium situations where stochastic events periodically reduce populations and slow the process of competitive 
Chapter I 129

exclusion (Huston, 1979; Naiman and others, 1993). Floristic 

studies have been conducted in areas containing riverscour 

habitat—as well as other riparian and upland ecosystems that 

are not subject to the scour-based disturbance regime—and 

have demonstrated the exceptional floristic richness and 

density of rare taxa supported by riverscour habitat. Examples 

include streams and adjacent land in the Big South Fork 

National River and Recreation Area (Bailey and Coe, 2001), 

the Obed Wild and Scenic River (Schmalzer and others, 

1985), and the New River Gorge National River (Grafton 

and Grafton, 1980).

Community Types and Vegetation 

Dynamics

Within the broad category of riverscour ecosystems, a 

number of community types can be differentiated, varying 

in their topographic positions, soil moisture conditions (with 

micro-scale heterogeneity from xeric to saturated), substrate 

types, levels of canopy coverage (from a nearly closed canopy 

to the near absence of trees), herbaceous vegetation, levels of 

species diversity, and in the frequency and intensity of flood 

disturbance, among other factors (Schmalzer, 1989). The 

general gradient of disturbance along the channel cross section 

is hypothesized to account for observed gradients in community composition, from herbaceous vegetation nearest the 

active channel to woody shrubs and trees along the forest edge 

(Hupp, 1983). Often, community types form linear zones—

either continuous or, more often, patchily distributed—

parallel to the active channel (Hupp, 1983). 

Considerable variability exists across riverscour 

community types in terms of scale and levels of geographic 

restriction. Some community types are relatively widespread, 

with distributions along multiple streams within a region, 

while others are known from only one stream or even from 

only one site. Levels of species diversity are also highly 

variable across riverscour community types. Some riverscour 

communities support high levels of plant-species richness, but 

others are largely dominated by a single species (for example 

Carex torta and Justicia americana), which may approximate 

monocultures (Vanderhorst and others, 2007). The community classification below uses colloquial names from the 

U.S. National Vegetation Classification (Jennings and others, 

2009; Franklin and others, 2012) unless otherwise noted.

Figure I–3. Appalachian Riverside Flatrock Community, Sandstone Falls, New River, New River Gorge National River, 

Raleigh County, West Virginia. Photograph by Alan M. Cressler, U.S. Geological Survey.
130 Insular Ecosystems of the Southeastern United States

Appalachian/Northern Piedmont Riverside 

Outcrop Woodland

This woodland community, termed the riverside flat rock 

plant community by Mitchem (2004) and riverside bedrock 

terrace woodland by Fleming and Patterson (2012), is ranked 

G1 and is ranked S1 in both Virginia and West Virginia 

(Vanderhorst and others, 2007; Fleming and Patterson, 2012). 

This community has been documented at two sandstone 

bedrock outcrop exposures along the New River (fig. I– 4) 

near two major waterfalls, with soils that are shallow (average 

of 0.7 inch, or about 1.8 centimeters at one site), acidic (pH 

approximately 3.9), highly organic, and nutrient-limited 

sands and sandy loams over bedrock (Suiter and Evans, 1999; 

Mitchem, 2004; Vanderhorst and others, 2007). Similar thinsoil communities on dolomite, calcareous shale, charnockite, 

and acidic schists and gneisses have been documented along 

high-gradient rivers in Virginia, including the Potomac, 

Shenandoah, and James Rivers (Belden and others, 2003; 

Fleming and Patterson, 2012). Soils are well drained, creating 

a relatively dry moisture regime. The mostly open canopy is 

dominated by Juniperus virginiana and Pinus virginiana with 

a diverse understory shrub and herb layer and a substantial 

component of mosses and lichens (Mitchem, 2004; Vanderhorst and others, 2007). Herbaceous species of this flatrock 

woodland community include Asplenium platyneuron, Juncus 

tenuis, and Hypericum gentianoides (Suiter and Evans, 1999). 

Within the New River Gorge National River, formation of this 

community is believed to have occurred following catastrophic flooding in the late 1800s by primary succession on 

scoured bedrock; however, flooding is now rare, and the flood 

regime has been altered by construction of the Bluestone Dam 

in 1949 (Vanderhorst and others, 2007). The consequences of 

this flooding regime alteration, in terms of the successional 

dynamics and ultimate fate of this rare community, are unclear.

Appalachian/Cumberland Sycamore–Birch 

Riverscour Woodland

This woodland community is dominated by Platanus 

occidentalis, Betula nigra and Salix spp., is classified as a 

temporarily flooded woodland, and is ranked G3 and S3 in 

West Virginia (Belden and others, 2003; Vanderhorst and 

others, 2007). Similar, related community types include the 

Piedmont/Central Appalachian Sycamore–River Birch Scour 

Woodland and the Central Appalachian/Piedmont Bedrock 

Floodplain Woodland (Fleming and Patterson, 2012; Schafale, 

2012). Sites of these related communities have been documented along the New River in West Virginia and the James, 

Potomac, and Maury Rivers in Virginia (Suiter and Evans, 

1999; Belden and others, 2003; Vanderhorst and others, 2007; 

Steury and others, 2008). This community occupies zones that 

are subjected to frequent high-energy flooding sufficient to 

damage and remove most trees, resulting in the maintenance 

Figure I–4. Bedrock outcrop exposures along the New River at 

Camp Brookside, New River Gorge National River, Summers County, 

West Virginia. Photograph by Alan M. Cressler, U.S. Geological Survey.

of an open canopy (Belden and others, 2003; Vanderhorst and 

others, 2007). Trees are typically short and flood-battered, 

with the tallest trees often being the youngest, which have 

not yet been exposed to extreme flooding. The substrate is a 

mixture of cobbles and boulders with deposition of gravel and 

sand but little true soil development. The moisture regime is 

heterogeneous, with contrasting microsites of well-drained, 

coarse-textured alluvium and poorly drained areas of standing 

water caused by proximity to the water table and fluvial 

topography. Along sycamore-birch riverscour woodland areas 

are narrow, linear zones dominated by herbaceous annuals 

such as Eragrostis hypnoides, Lindernia dubia, Cyperus 

squarrosus, Acalypha rhomboidea, Chamaesyce humistrata, 

Fimbristylis autumnalis, and Panicum dichotomiflorum

but which may also include herbaceous perennials such as 

Ludwigia palustris (Vanderhorst and others, 2007). Ecologically and spatially along the New River in West Virginia, 

the sycamore-birch riverscour woodland is intermediate 

between the riverscour prairie community, which is subjected 

to more frequent and intense scouring, and the sycamore-ash 
Chapter I 131

floodplain forest community, which is less affected by 

scouring (Vanderhorst and others, 2007). The shrub and herb 

layers are composed of flood-tolerant warm-season grasses 

and forbs with a large component of woody vines.

Riverscour Prairie

Riverscour prairie communities, such as those of the 

Appalachian and Interior Riverscour Barrens and Prairie 

Group, occupy the active channels of major rivers in zones 

intermediate between the low-water level and the bankfull 

level and are subjected to frequent, high-energy flooding and 

scouring (Vanderhorst and others, 2007). Unlike conditions 

in related riverscour woodland communities, the frequency 

and intensity of scouring in riverscour prairies are so high that 

most trees and other woody vegetation are either excluded 

entirely or they never reach mature stature (Vanderhorst, 

2000; Vanderhorst and others, 2007). Soils are thin psamments 

composed of cobble and sand and are commonly restricted 

to small fissures in exposed bedrock or to narrow pockets 

between closely spaced boulders, with vegetation generally 

restricted to these limited areas of soil accumulation (Belden 

and others, 2003; Vanderhorst and others, 2007). Some 

microsites are excessively drained and xeric, whereas others 

are poorly drained due to their locations relative to the water 

table or to small depressions in bedrock (potholes) created by 

scouring. Because riverscour prairies are exposed to frequent, 

high-intensity flooding, dramatic restructuring of substrates 

(in the case of boulders and cobble) or complete scouring 

(in the case of bedrock) is possible, such that some instances 

of these communities are considered ephemeral. Riverscour 

prairie vegetation is primarily tall, warm-season grasses 

(Andropogon gerardii, Panicum virgatum, and Sorghastrum 

nutans) and prairie-associated forbs with a limited shrub 

component and low-growing vines (Vanderhorst, 2000; Belden 

and others, 2003; Vanderhorst and others, 2007). Riverscour 

prairie communities along the New River are ranked G3 

and S1 in West Virginia (Vanderhorst and others, 2007). A 

riverscour prairie community type along the Cheat River 

in the Allegheny Mountains of West Virginia, dominated 

by Marshallia grandiflora, Rhododendron arborescens, 

Triantha glutinosa, and Platanthera flava var. herbiola, was 

classified as the Monongahela Barbara’s-buttons Riverscour 

Prairie and ranked G1 and S1 by Byers and others (2007). In 

Virginia, riverscour prairies, termed “riverside prairies” by 

Belden and others (2003) and Fleming and Patterson (2012), 

are State-rare, and the few known instances are confined to 

the Potomac River Gorge west of Washington, D.C., and 

to sites along the James, Shenandoah, and Maury Rivers. 

Globally and State-rare subtypes of riverside prairies in 

Virginia include the Piedmont/Central Appalachian Riverside 

Outcrop Prairie (G1, S1), the Piedmont/Central Appalachian 

Bedrock Floodplain Prairie (G3, S2) and the Ridge and Valley 

Gravel-Wash Riverside Prairie (G2, S2). These associations 

were delineated based on substrate type, relative elevation, 

and moisture regime (Fleming and Patterson, 2012). Although 

the exact number of riverscour prairie association occurrences 

is not known, depending upon the specifics of community 

classification and occurrence delineation, riverscour prairies 

are probably limited to fewer than 100 occurrences in the 

southeastern United States (Vanderhorst and others, 2007). 

Twisted-Sedge Rocky Creekbed

The twisted-sedge rocky creekbed community, also 

known as the Rocky Bar and Shore (Twisted Sedge Type), 

typically occupies boulder, cobble, and gravel bars within 

the channel shelves of high-gradient rivers and streams. This 

community consists of small patches of herbaceous vegetation 

strongly dominated by Carex torta, typically partially shaded 

by an open canopy of variable composition (Belden and 

others, 2003; Vanderhorst and others, 2007). The dominant 

community members, like Carex torta, are tough-rooted 

herbaceous perennials that are tolerant of high insolation and 

frequent inundation and scouring (Vanderhorst and others, 

2007; Schafale, 2012). Along rivers in the Delaware Estuary, 

Westervelt and others (2006) describe a “willow river-bar 

shrubland” community type dominated by Carex torta along 

with Salix nigra and other willow saplings.

American Water-Willow Cobble Bar (Justicia 

americana Herbaceous Vegetation)

The American water-willow cobble bar community 

typically occupies the heads and tails of islands or deposition 

bars and in other positions within the stream channel with 

substrates including bedrock, boulders, cobble, gravel, and 

sand (Belden and others, 2003; Vanderhorst and others, 2007; 

Schafale, 2012). Except during periods of low flow, these 

locations are saturated or partially inundated; during high-flow 

periods, they are completely submerged (Steury and others, 

2008). The vegetation is dominated by Justicia americana, 

an emergent aquatic herbaceous species. In more scour-prone 

sites, J. americana may form nearly homogenous stands. In 

more sheltered sites, other herbs such as Saururus cernuus and 

Schoenoplectus pungens may be present, along with scattered, 

flood-suppressed trees such as Acer saccharinum, Betula 

nigra, Fraxinus pennsylvanica, and Platanus occidentalis

(Vanderhorst and others, 2007).

Riverbank Annuals

The riverbank annuals community occupies positions 

within the active channel shelf that are submerged for long 

periods and highly exposed to scouring, such as the lower 

parts of alluvial bars and eroded banks, with substrates of sand 

or silty sand (Vanderhorst and others, 2007). Because of the 

intensity of the disturbance regime, this community can be 

highly ephemeral and is dominated by fast-growing annuals 
132 Insular Ecosystems of the Southeastern United States

and short-lived perennials such as Eragrostis hypnoides, 

Ludwigia palustris, Lindernia dubia, Eupatorium serotinum, 

and Cyperus squarrosus (Vanderhorst and others, 2007; 

Steury and others, 2008). The vegetation is almost exclusively herbaceous plants less than about 1.5 feet (roughly 

0.5 meter) in height, and vegetation development is greatest 

in areas that are fully insolated and may only become 

exposed during the low-flow period from early summer to 

early autumn (Vanderhorst and others, 2007; Steury and 

others, 2008).

Piedmont Riverscour Shrubland 

As described by Steury and others (2008), the Piedmont riverscour shrubland community is geographically 

restricted to the Piedmont region of Maryland and Virginia 

and specifically to sites along the high-gradient sections of 

the Potomac and Rappahannock Rivers that are relatively 

protected from scouring floods. This is primarily a woody 

shrub community, dominated by Carpinus caroliniana, 

Quercus rubra, Toxicodendron radicans, Nyssa sylvatica, 

Ulmus americana, Physocarpus opulifolius, Rhododendron periclymenoides, and Ilex decidua, with a sparse 

component of herbaceous graminoids and forbs (Steury 

and others, 2008).

Potomac Gorge Riverside Outcrop Barren 

The Potomac Gorge riverside outcrop barren community type is believed to be endemic only to the Potomac 

River Gorge adjacent to the Potomac River in Virginia and 

Maryland. The community is found on massive, exposed 

outcrops of schist, migmatitic metagraywacke, and granodiorite bedrock (Steury and others, 2008). The sites where 

this community is present, typically at around 10 to 50 feet 

(about 3 to 15 meters) above mean river level, are only 

rarely scoured by catastrophic floods. Vascular vegetation 

is sparse and limited to microhabitats in crevices and 

depressions that accumulate fine gravel and organic matter. 

This community type supports some State-rare riverscouradapted species also found in other related riverscour 

habitats such as Baptisia australis and Solidago simplex

var. racemosa (Steury and others, 2008).

Scoured Riverine Bluff Prairie/Duck River 

Scour Prairie

The scoured riverine bluff prairie community type, 

described by Pyne and Withers (1996), is apparently 

restricted to upstream-facing scoured limestone bluffs and 

bedrock shelves on the Duck River in the Central Basin 

of Tennessee and to related contexts on the Green and 

Kentucky Rivers in the Interior Low Plateaus of Kentucky. 

The dominant vegetation consists of robust grasses such as 

Andropogon gerardii, Sorghastrum nutans, and Chasmanthium latifolium, with woody vines and shrubs such as Cornus 

amomum ssp. obliqua, Hypericum prolificum, and Toxicodendron radicans (Pyne and Withers, 1996).

Hiwassee/Ocoee River Boulder Scour Vegetation 

The Hiwassee/Ocoee River boulder scour community, 

described by Major and others (1999), is apparently restricted 

to the Hiwassee and Ocoee Rivers within the Blue Ridge 

Physiographic Province of extreme southeastern Tennessee. 

Its herbaceous community is floristically similar to other 

riverscour prairie community types because it is dominated 

by perennial grasses and forbs such as Schizachyrium 

scoparium, Andropogon ternarius, Liatris microcephala, 

Coreopsis tripteris, Panicum species, Solidago species, and 

Agalinis species. This community is distinct, however, in that 

it supports the only known occurrences of Pityopsis ruthii, 

a G1-ranked and federally listed endangered species (Major 

and others, 1999; Thomson and Schwartz, 2006). A related 

community type, the Hiwassee/Ocoee River bedrock scour 

community, is composed of vegetation confined to crevices in 

scoured bedrock outcrops and is dominated by Schizachyrium 

scoparium, Schoenoplectus americanus, Juncus marginatus, 

Eupatorium serotinum, and a sparse assemblage of woody 

plants (Major and others, 1999). Both communities are 

ranked G2.

Yadkin Falls Bedrock Scour Community 

The Yadkin Falls bedrock scour community is extremely 

limited geographically and is ranked G1 because it is apparently restricted to a limited section of the Yadkin River in the 

eastern Piedmont of North Carolina, where it transects mafic 

meta-volcanic bedrock (Schafale, 2012). This community 

type occupies exposed bedrock outcrops and is dominated by 

herbaceous species such as Schizachyrium scoparium. This 

community is distinct from all other riverscour communities 

in its particular geology, extreme geographic restriction (to 

one river), and by the occurrence of Solidago plumosa, a 

G1-ranked species that had been a candidate for endangered 

species status until its removal in 2013 (Schafale, 2012).

Contributions to Regional Biodiversity 

Riparian corridors are highly dynamic ecosystems that 

can provide an exceptional diversity of habitats and ecological 

processes (Naiman and others, 1993). In the southeastern 

United States, riverscour communities contribute to regional 

biodiversity by supporting rare plant species, including several 

that are federally listed as endangered or threatened (fig. I–5).
Chapter I 133

Endemic and Biogeographically Disjunct Taxa

Among riparian ecosystems, riverscour communities 

commonly support unusually high numbers of rare species. 

For example, in floristic studies within the Big South Fork 

National River and Recreation Area and in the Obed Wild and 

Scenic River, both on the Cumberland Plateau of Tennessee, 

Bailey and Coe (2001) found that 75 percent and 56 percent, 

respectively, of all rare plant species in riparian zones were 

restricted to scour-prone boulder and cobble bars. Many of 

these rare species constituted new additions to the flora of 

the counties in which they were found. In another floristic 

survey of the Big South Fork National River and Recreation 

Area, which included upland areas as well as riparian zones, 

Shaw and Wofford (2003) determined that nearly all of 

the species of conservation concern were found in “unique 

habitats,” including riverscour areas. In addition to the many 

rare plant taxa supported by riverscour ecosystems, evidence 

supports their importance as habitat for rare animal species. 

For example, some American water-willow cobble bars 

provide habitat for globally rare dragonflies (Vanderhorst 

and others, 2007).

Figure I–5. Rare taxa associated with riverscour habitat in the southeastern United States: A, Conradina verticillata, Big South 

Fork National River and Recreation Area, Tennessee; B, Spiraea virginiana, Walker County, Georgia; C, Pityopsis ruthii, Cherokee 

National Forest, Polk County, Tennessee; D, Marshallia grandiflora, Daddys Creek, Cumberland County, Tennessee. Photograph A

by Nora Murdock, National Park Service; photographs B and C by Alan M. Cressler, U.S. Geological Survey; photograph D by 

Devin Rodgers, Austin Peay State University.

A

C

D

B
134 Insular Ecosystems of the Southeastern United States

Some of the rare vascular plant species supported 

by riverscour ecosystems are apparently endemic to these 

systems, meaning that their habitat is restricted to floodscoured riparian contexts. Conradina verticillata (fig. I–5A) is 

a federally listed threatened species and is considered to be a 

narrow endemic to riverscour ecosystems on the Cumberland 

Plateau, with populations consisting of small, geographically 

isolated, clonally reproducing colonies (Albrecht and Penagos, 

2012). C. verticillata belongs to a genus that contains only 

three other species, all of which are endemic to the Gulf Coast 

region of Alabama and Florida (Nicholson, 1986). Populations 

of C. verticillata have so far been documented only from 

nine major streams of the Cumberland Plateau. Characteristic 

habitat has little canopy coverage, includes crevices with 

moderately deep, sandy, well-drained soils with low organic 

matter, and is subjected to periodic flood scouring (U.S. Fish 

and Wildlife Service, 2011). Scouring not only maintains 

open habitat for C. verticillata but also helps propagate it by 

spreading dislodged plant sections downstream where they can 

root and establish new populations (Reed, 2003; Taylor, 2003).

Spiraea virginiana (fig. I–5B) is a federally listed 

threatened species and is a narrow endemic to riverscour 

ecosystems of the Appalachian Plateaus and Blue Ridge 

Physiographic Provinces (Ogle, 1991). S. virginiana is a shrub 

with a modular growth pattern, capable of sexual reproduction 

but more often reproducing clonally, that is anchored by a 

heavy lateral rhizome and a fine, fibrous root mass (Ogle, 

1991). Flood scouring is believed to be essential to the 

survival of S. virginiana by eliminating arboreal competition. 

Although much of the above-ground parts of S. virginiana

may be destroyed in flood events, the belowground part is 

typically capable of regenerating a clone after disturbance 

(Ogle, 1991). 

Calamovilfa arcuata is a G2-ranked grass with scattered 

occurrences from Tennessee to Oklahoma, such as in cobble 

and gravel bars along the Obed River, Daddys Creek, and 

the Emory River (Schmalzer, 1989; Wolfe and others, 2007) 

as well as the Caney Fork, Big South Fork Cumberland, and 

Clear Fork Rivers (Bailey and Coe, 2001). Solidago arenicola

is a G2-ranked species inhabiting sandy alluvium in scourprone riparian areas. This goldenrod was originally described 

from a single population along the Locust Fork River in 

Alabama, but depending on taxonomic classification, may 

also be present along rivers in Tennessee such as the Obed 

(Floden, 2012).

In some cases, riverscour habitats appear to support taxa 

that are endemic not only to the ecosystem type but also to 

particular streams. For example, Pityopsis ruthii (fig. I–5C ), 

a federally listed endangered species, is a shade-intolerant 

herbaceous perennial known only from occurrences in scoured 

habitat along the Hiwassee and Ocoee Rivers in Polk County, 

Tennessee, with a total population of around 10,000 individuals (Major and others, 1999; Thomson and Schwartz, 

2006). Solidago plumosa is known from only a single population on scoured mafic bedrock along the Yadkin River in 

North Carolina, and is related to other rare goldenrods found 

in riverscour contexts, including Solidago arenicola (Peirson 

and others, 2012; Schafale, 2012). Eurybia saxicastellii is 

riverscour-endemic aster that is apparently limited to the 

Rockcastle River in southeastern Kentucky (Campbell and 

Medley, 1989), the Big South Fork Cumberland River, and the 

New River in the Big South Fork drainage in Scott County, 

Tennessee (Bailey and Coe, 2001; Taylor, 2003). These three 

species are all ranked G1.

Some taxa, such as Phyllanthopsis phyllanthoides12

and Cerastium arvense var. velutinum, which have adapted 

to conditions in barrens and other disturbance-maintained, 

nonforested environments, can also occupy riverscour 

habitats (Weakley, 2012). Certain disturbance-adapted or 

stress-adapted species might be considered “near-endemics” 

if they are restricted to riverscour systems and related 

ecosystems. Ptilimnium nodosum, for example, is a federally listed endangered species occupying riverscour sites 

including cobble and gravel bars and bedrock outcrops that 

are seasonally flooded but are above water during low-flow 

periods (Bartgis, 1997). This short-lived semiaquatic species 

reproduces sexually and asexually and forms meta-populations 

as ramets and seeds are dispersed by flowing water to form 

new populations at downstream sites (Bartgis, 1997; Wells, 

2012). Populations have been documented along tributaries 

to the Potomac River in western Maryland and northeastern 

West Virginia and in several other sites in Alabama, Georgia, 

North and South Carolina, Virginia, and Arkansas (Wells, 

2012). Two ecotypes of the species are essentially riverscour 

endemics; however, a third ecotype is found in pond habitat 

(Marcinko, 2007), such that the species as a taxonomic unit 

may be considered a near-endemic. Marshallia grandiflora

(fig. I–5D) is a G2-ranked forb, generally considered endemic 

to the Appalachian Mountains and the Cumberland Plateau, 

with occurrences along rivers such as the Clear Fork (Bailey 

and Coe, 2001), the Big South Fork Cumberland (Medley and 

Wofford, 1980), the Obed River and its tributaries (Schmalzer 

and others, 1985), and along rivers such as the Gauley in the 

Allegheny Mountains of West Virginia (Byers and others, 

2007; Vanderhorst and others, 2010). Although this species 

is primarily riverscour-associated, its historical range has 

also included bog occurrences in North Carolina (Franklin 

and Finnegan, 2006), and thus it may be more appropriately 

categorized as a near-endemic.

Riverscour communities also support plant taxa that 

are rare at the state or regional level, even if they are globally secure and are not riverscour endemics. For example, 

riverscour communities within the New River Gorge National 

River support populations of Aristida purpurascens var. 

purpurascens, Carex woodii, Commelina erecta var. angustifolia, Coreopsis pubescens var. robusta, Galactia volubilis, 

Hypericum virgatum, Melica mutica, and Piptochaetium 

avenaceum, all of which are State-rare in West Virginia 

(Vanderhorst and others, 2007).

12Species name according to the Integrated Taxonomic Information 

System (2015).
Chapter I 135

In some cases, regionally rare plants in riverscour 

communities may be biogeographically disjunct from home 

ranges in other regions or may represent substantial range 

expansions. For example, Packera paupercula, disjunct 

from the species main range in the northern United States 

and Canada, occupies riverscour habitats of the Daniel 

Boone National Forest (Taylor, 2003). The Cheat River in 

West Virginia supports populations of Triantha glutinosa, 

a northern species reaching the southern end of its range at 

this site (Byers and others, 2007). Rare and threatened species 

associated with riverscour communities in the Obed Wild 

and Scenic River in Tennessee include Helenium brevifolium, 

Polygonella americana, Leucothoe racemosa, and Sporobolus 

junceus, which are disjunct from the Coastal Plain (Schmalzer, 

1989; Shaw and Wofford, 2003; Wolfe and others, 2007). 

Conservation Considerations

As explained in this section, conservation of riverscour 

ecosystems necessarily involves assessment and evaluation 

beyond the immediate vicinity of riverscour habitat occurrence. In particular, anthropogenic threats, conservation 

strategies, and scientific investigations commonly encompass 

large portions of watersheds. Because riparian ecosystems 

are influenced by human and natural dynamics at a range of 

scales, their conservation commonly requires landscape-level 

analysis in addition to localized hydrologic assessment. 

Threats to Ecosystem Integrity

Riverscour ecosystems are relatively rare throughout 

their range in the southeastern United States, primarily 

because they are confined to geographically small corridors 

along only relatively large-order streams (Bailey and Coe, 

2001). Within the Cumberland Plateau, fewer than 500 acres 

(about 200 hectares) of riverscour habitats are estimated to 

be in existence (Murdock and others, 2007; U.S. Fish and 

Wildlife Service, 2011). Even for riverscour sites on protected 

public land, an array of threats to ecosystem integrity has 

been identified.

A primary threat to riverscour ecosystems is hydrologic 

alteration and the associated effects on sediment dynamics. 

Construction activities that generate sediment from erosion 

and infrastructure projects such as upstream impoundments 

that change sediment transport dynamics have the potential 

to adversely affect riverscour ecosystems by interfering 

with natural patterns of substrate accumulation and removal 

(Meijer, 1976; Ogle, 1992; Wolfe and others, 2007). Because 

the fundamental integrity of these ecosystems depends on 

particular flood-scouring disturbance regimes, anthropogenic 

flow alteration—especially flood control—has the potential to 

facilitate succession toward large woody species, resulting in 

displacement of certain rare herbaceous riverscour communities (Westervelt and others, 2006). Impoundments also can 

interfere with the dispersal of some rare riverscour endemic 

plant species by blocking the effective downstream transport 

of seeds and clonal fragments (Ogle, 1992).

Impoundments and resulting changes in flooding regime 

are believed to have harmed some globally rare riverscour 

communities, including the Hiwassee/Ocoee River boulder 

scour community (Major and others, 1999). Individual populations of some rare riverscour-endemic species, including 

Conradina verticillata and Spiraea virginiana, are believed to 

have been destroyed by the construction of dams (Nicholson, 

1986; Ogle, 1991). Forest succession, resulting from a 

disruption in the flood-scouring disturbance regime because of 

dam construction, has been identified as a primary threat to the 

Eastern Red-cedar-Virginia Pine flatrock woodland community of the New River Gorge National River (Mitchem, 2004; 

Vanderhorst and others, 2007). Managers have recognized 

road and trail construction within Daniel Boone National 

Forest on the Cumberland Plateau of Kentucky as a potential 

threat to riverscour ecosystems due to the potential for altering 

flow regimes and sediment transport and deposition dynamics 

(Taylor, 2003). Erosion-produced sedimentation from development within watersheds has also been identified as a threat 

to riverscour ecosystems on the Big South Fork Cumberland 

River and the Obed River (Murdock and others, 2007).

In addition to direct hydrologic alteration from impoundments, land-use changes over entire stream basins have 

the potential to result in more indirect forms of hydrologic 

alteration, for example by changing the relative permeability 

of the land surface. Also, regional water-use activities such as 

upstream water withdrawals and wastewater releases have the 

potential to affect flow regimes in ways that could threaten the 

ecological integrity of riverscour communities (Murdock and 

others, 2007; Wolfe and others, 2007). Basin-level hydrologic 

alteration—leading to changes in flood and scouring regime—

is suspected to have played a role in observed declines of 

some populations of C. verticillata; however, neither the 

degree of alteration nor the link to vegetation dynamics 

has been conclusively established (U.S. Fish and Wildlife 

Service, 2011).

Invasive exotic species are also frequently cited as a 

potential threat to riverscour communities; however, the 

patterns of invasion and the particular species responsible 

appear to vary based on geographic region and riverscour 

community type. The Eastern Red-cedar-Virginia Pine flatrock 

woodland community of the New River Gorge National River 

is currently threatened with encroachment by exotic species 

including Lonicera japonica and Rosa multiflora (Mitchem, 

2004; Vanderhorst and others, 2007). At two sites within 

the New River Gorge National River (Camp Brookside and 

Keeneys Creek), the exotics Sedum sarmentosum, Lespedeza 

cuneata, and L. bicolor have formed well-established colonies 

capable of trapping sediment and thus altering the successional 

dynamics of the ecosystem (Vanderhorst and others, 2007). 

Riverscour woodland communities can be threatened by the 

invasive species Lythrum salicaria, Polygonum cuspidatum, 

and Sorghum halepense, which disperse widely and can 
136 Insular Ecosystems of the Southeastern United States

become established in high-energy riparian corridors (Belden 

and others, 2003; Westervelt and others, 2006). Invasive 

species in twisted-sedge rocky creekbed communities include 

Microstegium vimineum and Polygonum caespitosum var. 

longisetum (Vanderhorst and others, 2007). Riverscour prairie 

communities are threatened by the exotics Sorghum halepense, 

Centaurea biebersteinii, and Lythrum salicaria (Westervelt 

and others, 2006; Vanderhorst and others, 2007). Within the 

Big South Fork National River and Recreation Area, Ailanthus 

altissima, Albizia julibrissin, Spiraea japonica, Ligustrum 

sinense, Polygonum cuspidatum, Lythrum salicaria, Alliaria

petiolata, and Microstegium vimineum have been identified 

as problematic invasive species in riparian habitats and 

have been targeted for control by the National Park Service 

(U.S. Fish and Wildlife Service, 2011). Invasive species in the 

Obed Wild and Scenic River in Tennessee include Lespedeza 

cuneata, Elaeagnus umbellata, Albizia julibrissin, Coronilla 

varia, and Rosa multiflora (Wolfe and others, 2007). 

In some cases, invasive exotic species have been 

named as threats to individual riverscour-adapted species. 

For example, competition from the exotic species Rosa 

multiflora, Spiraea japonica, and Polygonum cuspidatum

has been implicated in limiting population sizes of the rare 

riverscour-endemic Spiraea virginiana (Ogle, 1991). Woody 

encroachment (by both native and exotic species) into riverscour habitat is also believed to be a threat to the rare endemic 

Conradina verticillata on the Cumberland Plateau (U.S. Fish 

and Wildlife Service, 2011).

Riverscour communities are also commonly threatened 

by direct mechanical destruction related to recreational 

activities such as fishing, hiking, camping, horseback riding, 

and boating. Sites on publicly protected land are not immune 

from damage, and such problems have been identified within 

the New River Gorge National River, the Gauley River 

National Recreation Area, the Big South Fork National River 

and Recreation Area, and the Obed Wild and Scenic River 

(Vanderhorst and others, 2007; U.S. Fish and Wildlife Service, 

2011). Trampling is a substantial problem in some riverscour 

habitats such as cobble bars because the dry, open nature 

of these sites makes them attractive locations for canoeists 

looking to rest or camp (Nicholson, 1986). 

Because riverscour communities maintain such close 

contact with regional surface water, contamination may also 

be a threat in some areas. Urban runoff and sewage releases 

discharged into rivers, for example, often contain elevated 

levels of nutrients and contaminants (Wolfe and others, 2007). 

Water contamination can also result from mining activities 

within watersheds, especially acidic drainage and siltation 

from coal mines (Meijer, 1976; Murdock and others, 2007).

Conservation Strategies

Riverscour ecosystems currently exist on a number of 

sites on protected public land (table I–2). In recognition of 

the ecological importance of riverscour ecosystems and the 

suite of potential threats facing them, a number of agencies 

have undertaken conservation and management efforts. The 

Appalachian Highlands Network, an inventory and monitoring 

network of the National Park Service, has been actively monitoring riverscour habitats and mapping populations of rare or 

endemic plants within the Big South Fork National River and 

Recreation Area (fig. I–6) and the Obed Wild and Scenic River 

since 2005 (Murdock and others, 2007; U.S. Fish and Wildlife 

Service, 2011) and along the New River since 2013. 

Knowledge Gaps

Although many important contributions have been made 

toward an understanding of riverscour ecosystems, they are 

still relatively poorly studied and several key knowledge 

gaps exist. For example, although it is generally understood 

that riverscour endemic species require specific disturbance 

regimes, the tolerance thresholds of flood scouring (intensity, 

frequency) are generally not known. Research opportunities 

exist to compare streamflow data with population demographic data for these endemic species to investigate the 

effects of changing disturbance regimes on species distribution 

(Ogle, 1992). Also, relatively few studies have been published 

on the fauna associated with riverscour vegetation. Hutchens 

and others (2009) found that in the Little Tennessee River in 

North Carolina, mats of Podostemum ceratophyllum provided 

important faunal habitat because they were associated with 

high abundance and biomass of benthic macroinvertebrates. A 

bee species, Anthidiellum notatum, previously unknown from 

Virginia, was discovered in a riverside outcrop prairie adjacent 

to the Potomac River in 2006, prompting the suggestion 

that “other rare bees might inhabit this globally rare plant 

community type” (Steury and others, 2009). More information 

on the fauna associated with other riverscour-adapted species 

could be helpful to elucidate larger ecosystem dynamics, such 

as trophic interactions and water-quality influences, in which 

these species may play important roles.

Possible Ecological Effects from Climate Change

Explicit discussions of climate change and its possible 

effects on riverscour ecosystems are limited in the literature. 

Even when climate change has been mentioned as a potential 

threat to these ecosystems, statements have typically included 

little elaboration or presentation of empirical evidence 

(for example, Byers and others, 2007). Rare plant species 

associated with riverscour habitat have been mentioned in 

association with several studies seeking to assess climatechange vulnerability for species of conservation concern 

at the state level. In West Virginia, for example, Byers and 

Norris (2011) assessed the climate-change vulnerability 

of Ptilimnium nodosum to be high and that of Spiraea 

virginiana and Marshallia grandiflora to be moderate. In 

Alabama, Davenport (2007) suggested that P. nodosum

“will need to shift with shifting climates and water sources” 

and “will be adversely affected if drier conditions prevail.” 
Chapter I 137

Table I–2. Selected occurrences of riverscour ecosystems on protected public land.

Stream Protected public land State Reference

New River New River Gorge National River West Virginia Grafton and McGraw (1976), Grafton 

and Grafton (1980), Suiter and Evans 

(1999), Mitchem (2004), Vanderhorst 

and others (2007)

Obed River, Clear Creek, and 

Daddys Creek

Obed Wild and Scenic River, 

Catoosa & Keyes-Harrison 

Wildlife Management Areas

Tennessee Schmalzer and others (1985), 

Schmalzer (1988, 1989), Wolfe 

and others (2007)

Cumberland, Clear Fork, and 

New Rivers

Big South Fork National River 

and Recreation Area

Tennessee and

Kentucky 

Medley and Wofford (1980), Bailey 

and Coe (2001), Shaw and Wofford 

(2003), Murdock and others (2007) 

Red River, Rockcastle River, 

and Cumberland River

Daniel Boone National Forest Kentucky Reed (2003), Taylor (2003)

Hiwassee and Ocoee Rivers Cherokee National Forest Tennessee Major and others (1999)

Shenandoah and Potomac Rivers Harpers Ferry National Historical Park West Virginia Vanderhorst (2000)

Potomac River Great Falls Park, unit of the George 

Washington Memorial Parkway

Virginia Steury and others (2008)

Chesapeake & Ohio Canal National 

Historical Park

Maryland Wiegand and Becker (1995)

Gauley River Gauley River National Recreation Area West Virginia Vanderhorst and others (2010)

Little River Little River Canyon National Preserve Alabama Rinehart (2008)

Figure I–6. Habitat monitoring at Leatherwood Ford cobble bar, Big South Fork Cumberland River, 

Big South Fork National River and Recreation Area, Tennessee. Photograph by Nora Murdock, 

National Park Service.
138 Insular Ecosystems of the Southeastern United States

These predictions, however, were made at the level of 

individual species rather than of plant assemblages or whole 

communities. Furthermore, these predictions did not include 

any detailed discussion of the mechanisms by which climate 

change would affect these species nor did they present new 

empirical evidence linking species population demographics 

to changes in habitat indicators.

In the absence of ecosystem-level empirical research 

concerning climate-change effects on riverscour communities, 

knowledge of the factors governing ecosystem integrity 

may be useful to generate hypotheses. Because riverscour 

ecosystems are largely maintained by flood-scour disturbance 

regimes, flow-regime alteration could be a key pathway by 

which changes in regional precipitation patterns might affect 

riverscour systems. Climate-change effects might be highly 

variable, however, from stream to stream and even among 

various riverscour habitats along the same stream. In many 

cases, natural flow regimes have already been substantially 

altered by upstream impoundments and by municipal water 

withdrawals and wastewater releases (Nicholson, 1986; Ogle, 

1991; Major and others, 1999; Murdock and others, 2007), 

such that it is unclear whether or how climate change might 

further affect flow regimes. Even for streams with relatively 

few direct forms of hydrologic alteration, any flow-regime 

changes resulting from shifts in regional precipitation patterns 

would likely have complex interactive effects with other 

sources of hydrologic alteration such as watershed-scale 

land-use change.

Theoretically, if climate change altered the flood-scour 

disturbance regime such that scour events become less 

frequent and (or) less intense, then the effects on scourmaintained communities might include woody encroachment 

by upland species, mirroring the effects produced by flood 

control from upstream impoundments. Conversely, an alteration of the scouring regime such that scour events become 

more frequent or more intense might elevate levels of stress 

beyond the tolerance thresholds of even riverscour-adapted 

endemic species. Even if the timing and intensity of flood 

scouring is not affected by climate change, alteration of 

low-flow conditions might moderate existing stress regimes 

in riverscour habitat. Seasonally xeric conditions, for 

example, are an important stress regime component for 

certain riverscour community types. An increase in summer 

base flow could theoretically expose such sites to invasion 

by drought-intolerant competitor species from surrounding 

upland forests (Wolfe and others, 2007). Additionally, because 

exotic species invasions can be substantially influenced by 

climate-change effects (Hellmann and others, 2008), regional 

changes in temperature or precipitation patterns could 

theoretically facilitate or exacerbate the expansion of exotic 

species into riverscour habitats. Evaluation of the likelihood 

of such scenarios would require research focused on linkages 

between changing climate conditions, flow regimes, population demographics of scour-adapted taxa, and ecological 

processes such as succession.

References

Albrecht, M., and Penagos, J., 2012, Seed germination 

ecology of three imperiled plants of rock outcrops in the 

southeastern United States: The Journal of the Torrey 

Botanical Society, v. 139, no. 1, p. 86–95.

Bailey, C., and Coe, F., 2001, The vascular flora of the 

riparian zones of the Clear Fork River and the New River 

in the Big South Fork National River and Recreation Area 

(BSFNRRA): Castanea, v. 66, no. 3, p. 252–274.

Bartgis, R., 1997, The distribution of the endangered plant 

Ptilimnium nodosum (Rose) Mathias (Apiaceae) in the 

Potomac River drainage: Castanea, v. 62, no. 1, p. 55–59.

Belden, A., Chazal, A., and Hobson, C., 2003, A natural 

heritage inventory of fourteen headwater sites in the 

Dragon Run Watershed: Richmond, Va., Virginia 

Department of Conservation and Recreation.

Byers, E., and Norris, S., 2011, Climate change vulnerability 

assessment of species of concern in West Virginia: Elkins, 

W. Va., West Virginia Divison of Natural Resources, 69 p.

Byers, E., Vanderhorst, J., and Streets, B., 2007, Classification and conservation assessment of high elevation wetland 

communities in the Allegheny Mountains of West Virginia: 

Elkins, W. Va., West Virginia Natural Heritage Program.

Campbell, J., and Medley, M., 1989, Aster saxicastellii

(Asteraceae), a new species from the Rockcastle River bars 

in southeastern Kentucky: Sida, v. 13, no. 3, p. 277–284.

Davenport, L., 2007, Climate change and its potential effects 

on Alabama’s plant life: Birmingham, Ala., Samford 

University, Vulcan Materials Center for Environmental 

Stewardship and Education.

Fenneman, N., and Johnson, D., 1946, Physical divisions of 

the United States: U.S. Geological Survey map prepared 

in cooperation with the Physiographic Commission, 

scale 1:7,000,000. 

Fleming, G., and Patterson, K., 2012, Natural communities 

of Virginia—Ecological groups and community types: 

Richmond, Va., Virginia Department of Conservation and 

Recreation, Division of Natural Heritage, 36 p.

Floden, A., 2012, Notes on two rare Solidago (Asteraceae) 

in Tennessee—S. Arenicola and S. simplex: Phytoneuron, 

v. 63, p. 1–4.

Franklin, M.A., and Finnegan, J.T., 2006, Natural Heritage 

Program list of rare plant species of North Carolina 2006: 

Raleigh, N.C., North Carolina Natural Heritage Program, 

Office of Conservation and Community.
Chapter I 139

Franklin, S., Faber-Langendoen, D., Jennings, M., 

Keeler-Wolf, T., Loucks, O., Peet, R., Roberts, D., and 

McKerrow, A., 2012, Building the United States National 

Vegetation Classification: Annali di Botanica, v. 2, p. 1–9.

Grafton, W., and Grafton, E., 1980, Literature review of 

geology/soils, plants and animal information pertinent to the 

New River Gorge National River, West Virginia, unpublished report submitted to staff at the New River Gorge 

National River.

Grafton, W., and McGraw, C., 1976, The vascular flora of 

New River Gorge, West Virginia: Beckley, West Virginia, 

Center for Extension and Continuing Education, West 

Virginia University, 19 p.

Grund, S., and Zimmerman, E., 2009, Youghiogheny River 

scour—An unusual place for unlikely plants: Keystone 

WILD! Notes, Pennsylvania Department of Conservation 

and Natural Resources, p. 15–17.

Hellmann, J.J., Byers, J.E., Bierwagen, B.G., and Dukes, 

J.S., 2008, Five potential consequences of climate change 

for invasive species: Conservation Biology, v. 22, no. 3, 

p. 534–543.

Hupp, C., 1983, Vegetation pattern on channel features in 

the Passage Creek Gorge, Virginia: Castanea, v. 48, no. 2, 

p. 62–72.

Huston, M., 1979, A general hypothesis of species diversity: 

American Naturalist, v. 113, no. 1, p. 81–101.

Hutchens, J., Wallace, J., and Romaniszyn, E., 2009, Role of 

Podostemum ceratophyllum Michx. in structuring benthic 

macroinvertebrate assemblages in a southern Appalachian 

river: Journal of the North American Benthological Society, 

v. 23, no. 4, p. 713–727.

Integrated Taxonomic Information System, 2015, The 

Integrated Taxonomic Information System, accessed 

October 5, 2015, at http://www.itis.gov.

Jennings, M., Faber-Langendoen, D., Loucks, O., Peet, R.,

and Roberts, D., 2009, Standards for associations and 

alliances of the U.S. National Vegetation Classification: 

Ecological Monographs, v. 79, no. 2, p. 173–199.

Lea, C., 2000, Plant communities of the Potomac Gorge and 

their relationship to fluvial factors: Fairfax, Virginia, 

George Mason University, Masters Thesis, 219 p.

Major, C., Bailey, C., Donaldson, J., McCoy, R., Nordman, C.,

Williams, M., and Withers, D., 1999, An ecological inventory of selected sites in the Cherokee National Forest: 

Tennessee Department of Environment and Conservation, 

Tennessee Division of Natural Heritage.

Marcinko, S., 2007, Pattern and process in rare plant conservation—An assessment of the southeastern U.S.: Chapel 

Hill, North Carolina, University of North Carolina, 

Masters Thesis, 88 p.

Medley, M., and Wofford, B., 1980, Thuja occidentalis L. 

and other noteworthy collections from the Big South Fork 

of the Cumberland River in McCreary County, Kentucky: 

Castanea, v. 45, no. 3, p. 213–215.

Meijer, W., 1976, A note on Podostemum ceratophyllum

Michx., as an indicator of clean streams in and around the 

Appalachian Mountains: Castanea, v. 41, no. 4, p. 319–324.

Mitchem, D., 2004, Characterization of the vegetation and soil 

of the forest communities at Camp Brookside in Summers 

County, West Virginia: Blacksburg, Va., Virginia Polytechnic Institute, Masters Thesis, 51 p.

Mouw, J., 2009, Influences of flooding and hyporheic 

exchange on floodplain plant richness and productivity: 

River Research and Applications, v. 25, p. 929–945.

Murdock, N., Hughes, J., and Emmott, R., 2007, Understanding hydrologic links between “river prairies” and other 

threatened riparian resources of the Cumberland Plateau, 

in Selleck, J., ed., Natural resource year in review—2006: 

Denver, Colo., National Park Service, Natural Resource 

Program Center, publication D-1859, p. 45–46.

Naiman, R., Decamps, H., and Pollock, M., 1993, The role 

of riparian corridors in maintaining regional biodiversity: 

Ecological Applications, v. 3, no. 2, p. 209–212.

Nicholson, R., 1986, To the arks with rabbitbane—Plant 

conservation at the Arnold Arboretum: Arnoldia, v. 46, 

no. 3, p. 23–25.

Ogle, D., 1991, Spiraea virginiana Britton—II. Ecology and 

species biology: Castanea, v. 56, no. 4, p. 297–303.

Ogle, D., 1992, Virginia spiraea (Spiraea virginiana Britton) 

recovery plan: Newton Corner, Mass., U.S. Fish and Wildlife Service, 39 p.

Peirson, J., Reznicek, A., and Semple, J., 2012, Polyploidy, 

infraspecific cytotype variation, and speciation in goldenrods—The cytogeography of Solidago subsect. Humiles

(Asteraceae) in North America: Taxon, v. 61, no. 1, 

p. 197–210.

Pyne, M., and Withers, D., 1996, Terrestrial and subterranean 

natural heritage survey of TVA Columbia Project lands: 

Tennessee Department of Conservation, Division of Natural 

Heritage, Report No. 96-001.

Reed, T., 2003, Viability assessment report for general forest 

habitat association, Daniel Boone National Forest, unpublished report submitted to the National Forest Service, 

U.S. Department of Agriculture, 24 p.
140 Insular Ecosystems of the Southeastern United States

Rinehart, N., 2008, Watershed condition assessment for Little 

River Canyon National Preserve, Alabama: Bowling Green, 

Kentucky, Western Kentucky University, Masters Thesis, 

166 p.

Rood, S.B., Goater, L., Mahoney, J.M., Pearce, C.M., and 

Smith, D.G., 2007, Floods, fire, and ice—Disturbance 

ecology of riparian cottonwoods: Canadian Journal of 

Botany, v. 85, no. 11, p. 1019–1032.

Schafale, M., 2012, Guide to the natural communities of 

North Carolina, fourth approximation: Raleigh, N.C., 

North Carolina Natural Heritage Program, 208 p.

Schmalzer, P., 1988, Vegetation of the Obed River Gorge 

System, Cumberland Plateau, Tennessee: Castanea, v. 53, 

no. 1, p. 1–32.

Schmalzer, P., 1989, Vegetation and flora of the Obed River 

gorge system, Cumberland Plateau, Tennessee: Journal of 

the Tennessee Academy of Science, v. 64, no. 3, p. 161–168.

Schmalzer, P., Patrick, T., and DeSelm, H., 1985, Vascular 

flora of the Obed Wild and Scenic River, Tennessee: 

Castanea, v. 50, no. 2, p. 71–88.

Shaw, J., and Wofford, B., 2003, Woody plants of Big South 

Fork National River and Recreation Area, Tennessee and 

Kentucky and floristic comparison of selected Southern 

Appalachian woody: Castanea, v. 68, no. 2, p. 119–134.

Steury, B., Droege, S., and Oberg, E., 2009, Bees (Hymenoptera: Anthophila) of a riverside outcrop prairie in Fairfax 

County, Virginia: Banisteria, v. 34, p. 17–24.

Steury, B., Fleming, G., and Strong, M., 2008, An emendation 

of the vascular flora of Great Falls Park, Fairfax County, 

Virginia: Castanea, v. 73, no. 2, p. 123–149.

Suiter, D., and Evans, D., 1999, Vascular flora and rare 

species of New River Gorge National River, West Virginia: 

Castanea, v. 64, no. 1, p. 23–49.

Taylor, D., 2003, Viability assessment report for riparian 

habitat association, Daniel Boone National Forest, unpublished report submitted to the National Forest Service, 

U.S. Department of Agriculture, 19 p.

Thomson, D.M., and Schwartz, M.W., 2006, Using population 

count data to assess the effects of changing river flow on 

an endangered riparian plant: Conservation Biology, v. 20, 

no. 4, p. 1132–1142.

U.S. Fish and Wildlife Service, 2011, Cumberland rosemary 

(Conradina verticillata) 5-year review—Summary and 

evaluation: Cookeville, Tenn., U.S. Fish and Wildlife 

Service, 14 p.

Vanderhorst, J., 2000, Plant communities of Harper’s Ferry 

National Historical Park—Analysis, characterization, and 

mapping: Elkins, W. Va., West Virginia Division of Natural 

Resources, West Virginia Natural Heritage Program, 30 p.

Vanderhorst, J.P., Jeuck, J., and Gawler, S.C., 2007, Vegetation 

classification and mapping of New River Gorge National 

River, West Virginia: Philadelphia, Pa., U.S. National Park 

Service, technical report NPS/NER/NRTR—2007/092, 

396 p.

Vanderhorst, J., Streets, B., Arcaro, Z., and Gawler, S., 2010, 

Vegetation classification and mapping of Gauley River 

National Recreation Area, West Virginia,: Philadelphia, Pa., 

U.S. National Park Service, technical report NPS/NER/

NRTR—2010/148, 288 p.

Weakley, A., 2012, Flora of the southern and mid-Atlantic 

States: Chapel Hill, N.C., University of North Carolina 

Herbarium, 377 p.

Wells, E.F., 2012, Seed germination and reproductive 

strategies in federally endangered Harperella (Harperella nodosum Rose, Apiaceae): Castanea, v. 77, no. 3, 

p. 218–223.

Westervelt, K., Largay, E., Coxe, R., McAvoy, W., Perles, S.,

Podniesinski, G., Sneddon, L., and Walz, K., 2006, A 

guide to the natural communities of the Delaware Estuary, 

Version I: Arlington, Va., NatureServe, 371 p.

Wiegand, R., and Becker, P., 1995, Rare plant survey along

the C&O Canal National Historical Park and Potomac 

River: Maryland Department of Natural Resources, 

Annapolis, Maryland.

Wolfe, W.J., Fitch, K.C., and Ladd, D.E., 2007, Alluvial 

bars of the Obed Wild and Scenic River, Tennessee: 

U.S. Geological Survey Scientific Investigations

 Map 2007–2972, 6 p.

Zimmerman, E., 2011, Floodplain scour community 

factsheet, terrestrial and palustrine plant communities of 

Pennsylvania, Pennsylvania Natural Heritage Program, 

accessed November 14, 2015, at http://www.naturalheritage.

state.pa.us/Community.aspx?=16011. 
Chapter J. Conclusions and Implications

Each of the ecosystems reviewed in the preceding 

chapters exists as an “archipelago” of spatially discrete habitat 

occurrences (islands) embedded within a matrix of contrasting 

ecosystems. Although these outcrops, grasslands, wetlands, 

and riparian zones are diverse in their geography, geomorphology, physical environments, and ecological processes, 

several unifying themes are common to all of them:

1. The contributions of these ecosystems to regional and 

global biodiversity—in terms of rare, endemic, and 

biogeographically disjunct plant and animal taxa—are 

substantial and likely disproportionate to their limited 

geographic areas;

2. Complex interactions of abiotic stress and disturbance 

factors help maintain ecological integrity in these 

ecosystems, with community composition commonly 

reflecting the interactive effects of multiple forms of 

stress and disturbance;

3. Each of these ecosystems has a history of anthropogenic 

damage, with common threats including land-use 

change, hydrologic alteration, invasive exotic species, 

and human disruption of historical disturbance regimes;

4. Despite decades of botanical and ecological research on 

these ecosystems, fundamental knowledge gaps persist 

concerning key determinants of ecosystem function; 

and in particular 

5. Climate-change effects have been poorly studied in 

these ecosystems, and empirically derived predictions 

of climate-change effects are exceedingly rare.

These conclusions have a number of implications for 

nature conservation in the southeastern United States and 

beyond. A primary implication is that preservation of the 

rare habitats provided by insular ecosystems is an important 

component of regional and global biodiversity conservation. 

The concentrations of rare, endemic, and disjunct taxa in 

insular ecosystems arguably qualify them as “biodiversity 

hotspots” (Myers and others, 2000) at a regional scale. Indeed, 

recent discoveries of new plants and animals—for example 

by DeBiase and Taylor (1997), Rogers and others (2004), Kral 

and Moffett (2009), and Hill (2010)—suggest that the full 

complement of biodiversity has yet to be described in these 

ecosystems. Furthermore, it should be noted that the total 

biodiversity supported by insular ecosystems is considerably 

greater than the limited subset described in this report. Within 

the southeastern United States, a number of additional insular 

ecosystems (ecosystems meeting the definitional criteria 

itemized in chapter A) are also noteworthy for their strong 

contributions to regional and global biodiversity (table J–1). 

Similar to the ecosystems discussed in the chapters of this 

report, these islands of distinct habitat tend to support rare, 

endemic, and disjunct plant taxa. Also common to these 

ecosystems are general trends of geographic rarity combined 

with threats from human activities. 

Table J–1. Examples of additional insular ecosystems in the southeastern United States.

Ecosystem description Reference

Ice ponds at cliff bases in the Blue Ridge of South Carolina Hill (1999)

Ephemeral ponds of the Cumberland Plateau Scheffers and others (2006)

Spray cliffs of the Chattooga Basin Zartman and Pittillo (1998)

Cliffs in river gorges of the Cumberland Plateau Nepal (2010), Boggess (2013)

Bayhead swamps of interior Louisiana Smith (1996)

Chalk bluffs and outcrops near the Black Prairie region in Mississippi Morris and others (1993)

Sandstone outcrops on the Cumberland Plateau, in the Atahama Grit region

of Georgia, and along the Fall Line in South Carolina 

Quarterman and others (1993)

Sandstone rockhouses of the Appalachian and Interior Low Plateaus Walck and others (1996)

Serpentine barrens of the eastern Appalachians and the Piedmont Tyndall and Hull (1999)

Southern Appalachian peat bogs Steward and Nilsen (1993)

Georgia Eocene chalk prairies Echols and Zomlefer (2010)

Cypress domes of peninsular Florida Kurz and Wagner (1953)

Sandhill vegetation patches of peninsular Florida Laessle (1958)
142 Insular Ecosystems of the Southeastern United States

Of course, insular ecosystems are by no means limited to 

the southeastern United States and in fact are probably a global 

phenomenon (table J–2). Across several continents, insular 

ecosystems (of which the examples presented here are illustrative but by no means fully representative) provide important 

contributions to regional and global biodiversity and are 

generally under threat from anthropogenic forces.

A second implication of these conclusions is that no 

single management approach is likely to be sufficient for 

conserving biodiversity across all insular ecosystems. As 

described in previous chapters, many ecosystems have been 

subjected to such degrees of qualitative degradation and 

quantitative destruction that they can be conceived of as 

“endangered ecosystems,” a term applied by Noss and others 

(1995) to several insular ecosystems including limestone cedar 

glades, pocosins, Carolina (Delmarva) bays, Appalachian bogs, 

and upland wetlands of the Tennessee Highland Rim. Some 

anthropogenic threats, such as development and conversion 

to agriculture, may be effectively mitigated by traditional 

conservation approaches, such as public land ownership and 

habitat conservation plans for rare taxa. Mitigating threats 

from larger-scale processes, such as regional hydrologic 

alteration, invasive species, and contamination of air and 

water, will likely require additional conservation measures and 

the coordination of diverse groups of stakeholders. Because 

many ecosystem occurrences have been severely degraded 

but not destroyed, restoration efforts offer some promise for 

expanding and improving habitat for rare taxa (Vandevender, 

2006; Duncan and others, 2008; De Steven and others, 2010). 

As was discussed for several ecosystems in previous chapters, 

these restoration efforts commonly require holistic assessment 

of landscape-scale as well as site-specific environmental 

conditions and ecological processes (Pavlovic, 1994; Bruland 

and others, 2003; VanZandt and others, 2005; Vepraskas and 

others, 2007). In cases where suitable but unoccupied habitat 

exists for rare taxa, management options may include reintroduction efforts to establish new populations (Johnson, 1996; 

Kutner and Morse, 1996; Morse, 1996; Wiser and others, 1998; 

Nordman, 2004; Albrecht and McCue, 2010). 

A third and related implication is that best management 

practices for insular ecosystems may differ from those for 

geographically continuous and widespread ecosystems. By 

definition, insular ecosystems provide naturally fragmented 

habitat to their constituent organisms. As a result, some taxa 

of conservation concern may be subject to the commonly 

observed effects of small and isolated populations, including 

barriers to gene flow, reduced population genetic diversity, 

and increased vulnerability to localized extinctions (Godt 

and others, 1996). Where these factors are at work, effective 

protection of biodiversity in insular ecosystems (metaphorical 

archipelagos) may require the preservation of discrete habitat 

occurrences (islands) ranging across the full spectrum of areal 

size, geographic range, and environmental conditions. The 

need to preserve habitat diversity within any given category 

of insular ecosystem—rather than simply conserving a few 

large “representative” examples—stems from two primary 

observations. First, evidence suggests that communities in 

small islands are not simply representative subsets of those in 

larger islands because of differences in ecologically relevant 

environmental conditions that are associated with island size 

(Snodgrass and others, 2000; Sharitz, 2003). Indeed, small 

islands such as “dry-end wetlands” in geographically isolated 

upland depressions likely serve essential ecological functions 

that are not redundant to their larger counterparts (Whigham, 

1999). Second, small islands may be essential as “stepping 

stones” to facilitate gene flow between larger islands, especially 

for plant taxa incapable of long-distance dispersal (Mann and 

others, 1999; Leibowitz, 2003; Tiner, 2003; Cofer and others, 

2008). Thus, effective management of insular ecosystems may 

necessitate “portfolio diversification” in terms of the individual 

habitat occurrences that are targeted for protection.

All of the conclusions above highlight the need for 

ongoing ecological research in insular ecosystems. Investigations of the factors and processes that regulate biodiversity in 

these ecosystems—including stress and disturbance regimes, 

successional dynamics, relationships across trophic levels, and 

human influences—have the potential to contribute substantially toward the emerging field of conservation biogeography 

Table J–2. Examples of insular ecosystems outside the southeastern United States.

Ecosystem description Reference

Athabasca sand dunes in Saskatchewan, Canada Purdy and others (1994), (Hill, 1999)

Ironstone outcrops in Brazil Jacobi and others (2007)

Rupestrian fields in Brazil Ribeiro and Fernandes (2000)

Chalk barrens and riparian canyons in the Pikes Peak region of Colorado, USA Kelso and others (2001)

Serpentine barrens in California, USA Harrison and others (2006)

Calcareous alvars in several European regions Pärtel and others (2001)

Vernal pools occurring globally in regions with Mediterranean climate Keeley and Zedler (1998), Zedler (2003)

Granitic and gneissic inselbergs in several African countries, in Brazil, and in 

western Australia

Porembski and Barthlott (2000)

Quartzite and dolomite outcrops and grasslands in the Transvaal Escarpment, 

South Africa

Matthews and others (1993)
Chapter J 143

(Whittaker and others, 2005; Ladle and Whittaker, 2011). 

Likewise, a fusion of investigative approaches from the fields 

of biogeography and conservation biology may help elucidate 

the determinants of community structure within islands and of 

distribution patterns for rare taxa across archipelagos. Identification of these determinants and improved understanding of 

their dynamic interactions may prove critical to conserving 

insular ecosystems as regional climatic conditions change.

Profound and well-documented ecological effects of 

climate change at a global level (Parmesan and Yohe, 2003; 

Parmesan, 2006; Walther, 2010) suggest that more vigorous 

and focused efforts are needed to anticipate such changes 

in insular ecosystems and to recognize them as they occur. 

In particular, hypothesis-driven studies are needed that 

mechanistically link long-term demographic trends of sensitive populations to changing environmental conditions; for 

example, see McCarthy and Evans (2000) and Wolfe and others 

(2004). Also important are efforts to discern and interpret 

differences in community structure along climatic gradients; 

for example, see Wiser (1998). For some ecosystems, the stress 

and disturbance factors described in the previous chapters 

may represent reasonable starting points for hypothesis 

generation and testing. For example, community structure is 

strongly linked to hydrologic conditions in a variety of insular 

ecosystems such as depression wetlands and rock outcrops. 

Soil surface temperatures and (or) temperature-dependent 

evapotranspiration dynamics are important in some cases, 

as are weather-related disturbances such as storms, fires, and 

droughts. All of these factors—as well as their interactive 

effects—may be affected by climate change over the coming 

decades (Easterling and Evans, 2000; Hanson and Weltzin, 

2000; Trenberth and others, 2013; Mitchell and others, 2014). 

The process of linking changing climatic conditions to 

ecological outcomes will necessarily be complex and subject 

to multiple sources of uncertainty. A fundamental source of 

uncertainty concerns the magnitude, velocity, and nature of 

regional climate change (Chen and others, 2003; Monier 

and others, 2014). Uncertainty is compounded at each of 

several additional levels of analysis and interpretation that are 

required for prediction of ecological outcomes. Changes in 

regional climate may produce shifts in localized microclimatic 

conditions that vary across microhabitats (Kutner and Morse, 

1996; Ashcroft, 2010; Storlie and others, 2014). Interactive 

effects between multiple types of abiotic change (for example, 

increased soil surface temperature combined with increased 

drought frequency) may shift localized stress regimes in ways 

that could not be predicted by examining each abiotic change 

independently. Based on idiosyncratic stress-tolerance capacities, different plant taxa may respond differently to changing 

stress regimes (Ahmad and Prasad, 2011; Panda and others, 

2013), with consequences for interspecific interactions that 

may be difficult to predict. At a global level, cases have been 

documented in which phenophase change occurred at different 

rates between taxa linked by mutualistic relationships such as 

pollination, leading to a breakdown of interspecific interactions 

known as phenological mismatch (Parmesan, 2006; Fabina 

and others, 2010; Walther, 2010; Thackeray and others, 

2010). Evaluating the potential for such situations in insular 

ecosystems requires detailed knowledge concerning interspecific interactions and the abiotic conditions that regulate 

them. All of these sources of uncertainty highlight the need for 

increased scientific investigation of insular ecosystems and the 

webs of interactions that regulate their ecological function. 

References

Ahmad, P., and Prasad, M., eds., 2011, Environmental 

adaptations and stress tolerance of plants in the era of 

climate change: New York, Springer.

Albrecht, M., and McCue, K., 2010, Changes in demographic 

processes over long time scales reveal the challenge of 

restoring an endangered plant: Restoration Ecology, v. 18, 

no. S2, p. 235–243.

Ashcroft, M.B., 2010, Identifying refugia from climate 

change: Journal of Biogeography, v. 37, p. 1407–1413.

Boggess, L., 2013, Cliff ecology of the Big South Fork 

National River and Recreation Area: Appalachian State 

University, Masters Thesis.

Bruland, G., Hanchey, M., and Richardson, C., 2003, Effects 

of agriculture and wetland restoration on hydrology, soils, 

and water quality of a Carolina bay complex: Wetlands 

Ecology and Management, v. 11, p. 141–156.

Chen, M., Pollard, D., and Barron, E., 2003, Comparison of 

future climate change over North America simulated by two 

regional models: Journal of Geophysical Research, v. 108, 

no. D12, p. ACL 3–1—ACL 3–19.

Cofer, M., Walck, J., and Hidayati, S., 2008, Species richness and exotic species invasion in Middle Tennessee cedar 

glades in relation to abiotic and biotic factors: The Journal 

of the Torrey Botanical Society, v. 135, no. 4, p. 540–553.

DeBiase, A., and Taylor, B., 1997, Aglaodiaptomus atomicus, 

a new species (Crustacea: Copepoda: Calanoida: Diaptomidae) from freshwater wetland ponds in South Carolina, 

USA, and a redescription of A. saskatchewanensis

(Wilson 1958): Proceedings of the Biological Society 

of Washington, v. 110, no. 4, p. 569–580.

De Steven, D., Sharitz, R., and Barton, C., 2010, Ecological 

outcomes and evaluation of success in passively restored 

southeastern depressional wetlands: Wetlands, v. 30, no. 6, 

p. 1129–1140.

Duncan, R., Anderson, C., Sellers, H., and Robbins, E., 2008, 

The effect of fire reintroduction on endemic and rare plants 

of a southeastern glade ecosystem: Restoration Ecology, 

v. 16, no. 1, p. 39–49.
144 Insular Ecosystems of the Southeastern United States

Easterling, D., and Evans, J., 2000, Observed variability and 

trends in extreme climate events—A brief review: Bulletin 

of the American Meteorological Society, v. 81, no. 3, 

p. 417–425.

Echols, S., and Zomlefer, W., 2010, Vascular plant flora of 

the remnant blackland prairies in Oaky Woods Wildlife 

Management Area, Houston County, Georgia: Castanea, 

v. 75, no. 1, p. 78–100.

Fabina, N.S., Abbott, K.C., and Gilman, R.T., 2010, 

Sensitivity of plant–pollinator–herbivore communities

to changes in phenology: Ecological Modelling, v. 221, 

no. 3, p. 453–458.

Godt, M., Johnson, B., and Hamrick, J., 1996, Genetic 

diversity and population size in four rare southern Appalachian plant species: Conservation Biology, v. 10, no. 3, 

p. 796–805.

Hanson, P.J., and Weltzin, J.F., 2000, Drought disturbance 

from climate change—Response of United States forests: 

The Science of the Total Environment, v. 262, p. 205–220.

Harrison, S., Safford, H., and Grace, J., 2006, Regional 

and local species richness in an insular environment—

Serpentine plants in California: Ecological Monographs, 

v. 76, no. 1, p. 41–56.

Hill, J., 2010, A new species of Melanoplus (Orthoptera: 

Acrididae: Melanoplinae) from the cedar glades of 

Tennessee, USA: Journal of Orthoptera Research, v. 19, 

no. 2, p. 341–345.

Hill, S., 1999, The relict flora of ice ponds in South Carolina: 

Castanea, v. 64, no. 1, p. 14–22.

Jacobi, C.M., Carmo, F.F., Vincent, R.C., and Stehmann, J.R., 

2007, Plant communities on ironstone outcrops—A diverse 

and endangered Brazilian ecosystem: Biodiversity and 

Conservation, v. 16, no. 7, p. 2185–2200.

Johnson, B., 1996, Southern Appalachian rare plant reintroductions on granite outcrops, in Faulk, D., Miller, C., and 

Olwell, M., eds., Restoring diversity—Strategies for reintroduction of endangered plants: Washington, D.C., Island 

Press, p. 433–444.

Keeley, J., and Zedler, P., 1998, Characterization and 

global distribution of vernal pools, in Witham, C., 

Bauder, E., Belk, D., Ferren, W., and Ornduff, R., eds., 

Ecology, Conservation, and Management of Vernal Pool 

Ecosystems—Proceedings from a 1996 Conference of the 

California Native Plant Society, Sacramento, California, 

p. 1–14.

Kelso, S., Hall, C., and Maentz, G., 2001, The role of landscape anomalies in regional plant conservation, in Southwestern Rare and Endangered Plants, Proceedings of the 

Third Conference: Flagstaff, Arizona, p. 13–19.

Kral, R., and Moffett, J., 2009, Xyris spathifolia (Xyridaceae), 

a new xyrid from the Ketona dolomite-limestone glades 

of Alabama: Journal of the Botanical Research Institute 

of Texas, v. 3, no. 2, p. 469–478.

Kurz, H., and Wagner, K., 1953, Factors in cypress dome 

development: Ecology, v. 34, no. 1, p. 157–164.

Kutner, L., and Morse, L., 1996, Reintroduction in a changing 

climate, in Faulk, D., Miller, C., and Olwell, M., eds., 

Restoring diversity—Strategies for reintroduction of endangered plants: Washington, D.C., Island Press, p. 23–48.

Ladle, R., and Whittaker, R., eds., 2011, Conservation 

biogeography: Oxford, United Kingdom, Blackwell, 320 p.

Laessle, A., 1958, The origin and successional relationship 

of sandhill vegetation and sand-pine scrub: Ecological 

Monographs, v. 28, no. 4, p. 361–387.

Leibowitz, S., 2003, Isolated wetlands and their functions—

An ecological perspective: Wetlands, v. 23, no. 3, 

p. 517–531.

Mann, L., King, A., Dale, V., Hargrove, W., Washington-Allen, 

R., Pounds, L., and Ashwood, T., 1999, The role of soil 

classification in geographic information system modeling 

of habitat pattern—Threatened calcareous ecosystems: 

Ecosystems, v. 2, p. 524–538.

Matthews, W.S., van Wyk, A.E., and Bredenkamp, G.J., 

1993, Endemic flora of the north-eastern Transvaal 

Escarpment, South Africa: Biological Conservation, 

v. 63, no. 1, p. 83–94.

McCarthy, S., and Evans, J., 2000, Population dynamics of 

overcup oak (Quercus lyrata) in a seasonally flooded karst 

depression: Journal of the Torrey Botanical Society, v. 127, 

no. 1, p. 9–18.

Mitchell, R.J., Liu, Y., O’Brien, J.J., Elliott, K.J., Starr, G., 

Miniat, C.F., and Hiers, J.K., 2014, Future climate and fire 

interactions in the southeastern region of the United States: 

Forest Ecology and Management, v. 327, p. 316–326.

Monier, E., Gao, X., Scott, J.R., Sokolov, A.P., and 

Schlosser, C.A., 2014, A framework for modeling 

uncertainty in regional climate change: Climatic Change, 

v. 131, no. 1, p. 51–66.

Morris, M., Bryson, C., and Warren, R., 1993, Rare vascular 

plants and associate plant communities from the Sand Creek 

Chalk Bluffs, Oktibbeha County, Mississippi: Castanea, 

v. 58, no. 4, p. 250–259.

Morse, L., 1996, Plant rarity and endangerment in North 

America, in Faulk, D., Miller, C., and Olwell, M., eds., 

Restoring diversity—Strategies for reintroduction of endangered plants: Washington, D.C., Island Press, p. 7–22.
Chapter J 145

Myers, N., Mittermeier, R.A., Mittermeier, C.G., Fonseca, 

G.A.B., and Kent, J., 2000, Biodiversity hotspots for 

conservation priorities: Nature, v. 403, p. 853–858.

Nepal, B., 2010, Characterization of the vegetational 

communities associated with ancient Juniperus virginiana L.

stands in the Obed Wild and Scenic River Gorge: Appalachian State University, Masters Thesis.

Nordman, C., 2004, Vascular plant community classification 

for Stones River National Battlefield: Durham, N.C., 

NatureServe. 

Noss, R., LaRoe, E., and Scott, J., 1995, Endangered ecosystems of the United States—A preliminary assessment of

loss and degradation: Washington, D.C., U.S. Department

of the Interior. 

Panda, B.B., Achary, V.M.M., Mahanty, S., and Panda, K.K., 

2013, Plant adaptation to abiotic and genotoxic stress—

Relevance to climate change and evolution, in Tuteja, N., 

and Gill, S.S., eds., Climate change and plant abiotic stress 

tolerance: Weinheim, Germany, Wiley-VCH Verlag GmbH 

& Co. KGaA.

Parmesan, C., 2006, Ecological and evolutionary responses

to recent climate change: Annual Review of Ecology, 

Evolution, and Systematics, v. 37, no. 2006, p. 637–669.

Parmesan, C., and Yohe, G., 2003, A globally coherent fingerprint of climate change impacts across natural systems: 

Nature, v. 421, no. 6918, p. 37–42.

Pärtel, M., Moora, M., and Zobel, M., 2001, Variation in 

species richness within and between calcareous (alvar) 

grassland stands—The role of core and satellite species: 

Plant Ecology, v. 157, p. 203–211.

Pavlovic, N., 1994, Disturbance-dependent persistence of 

rare plants—Anthropogenic impacts and restoration 

implications, in Bowles, M., and Whelan, C., eds., 

Restoration of endangered species—Conceptual issues, 

planning and implementation: Melbourne, Australia, 

Cambridge University Press, p. 159–193.

Porembski, S., and Barthlott, W., 2000, Granitic and gneissic 

outcrops (inselbergs) as centers of diversity for desiccationtolerant vascular plants: Plant Ecology, v. 151, no. 1, 

p. 19–28.

Purdy, B., Bayer, R., and Macdonald, S., 1994, Variation, 

breeding system evolution, and conservation of the narrow 

sand dune endemic Stellaria arenicola and the widespread 

S. longipes (Caryophyllaceae): American Journal of Botany, 

v. 81, no. 7, p. 904–911.

Quarterman, E., Burbanck, M., and Shure, D., 1993, Rock 

outcrop communities—Limestone, sandstone, and granite, 

in Martin, W., Boyce, S., and Echternacht, A., eds., 

Biodiversity of the southeastern United States—Upland 

terrestrial communities: New York, John Wiley and Sons, 

Inc., p. 35–86.

Ribeiro, K., and Fernandes, G., 2000, Patterns of abundance

of a narrow endemic species in a tropical and infertile 

montane habitat: Plant Ecology, v. 147, p. 205–218.

Rogers, D., Jensen, J., and Floyd, T., 2004, Eubranchipus 

stegosus n. sp. (Crustacea: Anostraca), a bizarre new

species of fairy shrimp from the southeastern United States: 

Hydrobiologia, v. 528, p. 201–207.

Scheffers, B.R., Berton, J., Harris, C., and Haskell, D.G., 

2006, Avifauna associated with ephemeral ponds on the 

Cumberland Plateau: Journal of Field Ornithology, v. 77, 

no. 2, p. 178–183.

Sharitz, R., 2003, Carolina bay wetlands—Unique habitats 

of the southeastern United States: Wetlands, v. 23, no. 3, 

p. 550–562.

Smith, L., 1996, The rare and sensitive natural wetland plant 

communities of interior Louisiana: Baton Rouge, La., 

Louisiana Natural Heritage Program, Department of Wildlife and Fisheries, 40 p.

Snodgrass, J., Komoroski, M., Bryan, A., and Burger, J., 2000, 

Relationships among isolated wetland size, hydroperiod, 

and amphibian species richness—Implications for wetland 

regulations: Conservation Biology, v. 14, no. 2, p. 414–419.

Steward, C., and Nilsen, E., 1993, Association of edaphic 

factors and vegetation in several isolated Appalachian peat 

bogs: Bulletin of the Torrey Botanical Club, v. 120, no. 2, 

p. 128–135.

Storlie, C., Merino-Viteri, A., Phillips, B., VanDerWal, J., 

Welbergen, J., and Williams, S., 2014, Stepping inside the 

niche—Microclimate data are critical for accurate assessment of species’ vulnerability to climate change: Biology 

Letters, v. 10, no. 9, 4 p.

Thackeray, S.J., Sparks, T.H., Frederiksen, M., Burthe, S., 

Bacon, P.J., Bell, J.R., Botham, M.S., Brereton, T.M., 

Bright, P.W., Carvalho, L., Clutton-Brock, T., Dawson, A., 

Edwards, M., Elliott, J.M., and others, 2010, Trophic level 

asynchrony in rates of phenological change for marine, 

freshwater and terrestrial environments: Global Change 

Biology, v. 16, no. 12, p. 3304–3313.

Tiner, R.W., 2003, Geographically isolated wetlands of the 

United States: Wetlands, v. 23, no. 3, p. 494 –516.
146 Insular Ecosystems of the Southeastern United States

Trenberth, K.E., Dai, A., van der Schrier, G., Jones, P.D., 

Barichivich, J., Briffa, K.R., and Sheffield, J., 2013, Global 

warming and changes in drought: Nature Climate Change, 

v. 4, no. 1, p. 17–22.

Tyndall, R., and Hull, J., 1999, Vegetation, flora, and plant 

physiological ecology of serpentine barrens of eastern 

North America, in Anderson, R., Fralish, J., and Baskin, J., 

eds., Savannas, barrens, and rock outcrop plant communities of North America: New York, Cambridge University 

Press, p. 67– 82.

Vandevender, J., 2006, Native plant restoration at Stones 

River National Battlefield, in People, Places and Parks, 

Proceedings of the 2005 George Wright Society Conference 

on Parks, Protected Areas and Cultural Sites: The George 

Wright Society, Inc., p. 146 –148.

VanZandt, P., Collins, E., Losos, J., and Chase, J., 2005, 

Implications of food web interactions for restoration of 

Missouri Ozark glade habitats: Restoration Ecology, v. 13, 

no. 2, p. 312–317.

Vepraskas, M., White, J., Amoozegar, A., Hesterberg, D., 

Huffman, R., Skaggs, R., Gregory, J., Broome, S., Caldwell, P.,

Davis, K., Abit, S., Smith, H., Zelasko, A., and Brownfield, C.,

2007, Methodology to assess soil, hydrologic, and site 

parameters that affect wetland restoration, phase 2: Raleigh, 

North Carolina, North Carolina State University, 241 p.

Walck, J., Baskin, J., Baskin, C., and Francis, S., 1996, Sandstone rockhouses of the eastern United States, with particular reference to the ecology and evolution of the endemic 

plant taxa: The Botanical Review, v. 62, no. 4, p. 311–362.

Walther, G., 2010, Community and ecosystem responses to 

recent climate change: Philosophical Transactions of the 

Royal Society, v. 365, p. 2019–2024.

Whigham, D., 1999, Ecological issues related to wetland 

preservation, restoration, creation and assessment: Science 

of the Total Environment, v. 240, p. 31– 40.

Whittaker, R.J., Araújo, M.B., Jepson, P., Ladle, R.J., 

Watson, J.E.M., and Willis, K.J., 2005, Conservation 

biogeography—Assessment and prospect: Diversity and 

Distributions, v. 11, p. 3–24.

Wiser, S., 1998, Comparison of southern Appalachian highelevation outcrop plant communities with their northern 

Appalachian counterparts: Journal of Biogeography, v. 25, 

no. 3, p. 501–513.

Wiser, S., Peet, R., and White, P., 1998, Prediction of rareplant occurrence—A southern Appalachian example: 

Ecological Applications, v. 8, no. 4, p. 909–920.

Wolfe, W.J., Evans, J.P., McCarthy, S., Gain, W.S., and 

Bryan, B.A., 2004, Tree-regeneration and mortality 

patterns and hydrologic change in a forested karst 

wetland—Sinking Pond, Arnold Air Force Base, Tennessee: 

U.S. Geological Survey Water-Resources Investigations 

Report 03– 4217, 62 p. 

Zartman, C., and Pittillo, J., 1998, Spray cliff communities of 

the Chattooga Basin: Castanea, v. 63, no. 3, p. 217–240.

Zedler, P.H., 2003, Vernal pools and the concept of “isolated 

wetlands”: Wetlands, v. 23, no. 3, p. 597– 607.
Glossary

Allelopathy The process by which one plant 

inhibits the growth or germination of another 

by releasing chemicals into the environment.

Autecology The study of the ecological 

relationships between an individual species 

and its environment.

C3 photosynthesis The most common metabolic pathway for carbon fixation by plants, 

common in mesic environments.

C4 photosynthesis A metabolic pathway 

for carbon fixation by plants, most commonly 

found in environments characterized by 

high insolation, high temperatures, and low 

moisture availability.

CAM photosynthesis The Crassulacean 

acid metabolism pathway for carbon fixation 

by plants, most commonly found in environments characterized by high insolation, high 

temperatures, and low moisture availability.

Calcareous Characterized by the presence 

of calcium carbonate derived from sedimentary carbonate rocks such as limestone 

and dolomite.

Climax The final stage of succession 

attainable by a plant community based on 

the environmental conditions in a particular 

place at a particular time.

Cryptogam A plant that reproduces by 

spores without flowers or seeds, including 

ferns and mosses.

Community A group consisting of populations of two or more different species 

inhabiting the same geographic area at the 

same time.

Congener An organism belonging to the 

same taxonomic group as another.

Disjunct An isolated population of plants 

or animals far from their typical geographic range.

Dispersal The movement of individual 

organisms or their reproductive structures 

from one geographic area to another.

Distribution The spatial arrangement of a 

biological taxon.

Disturbance An event or episodic series of 

events, such as fires and floods, producing 

intense environmental stress over a relatively 

short period of time.

Ecotone A transitional zone between 

ecological communities with different 

environmental conditions.

Edaphic Related to soil.

Endemic Primarily restricted to a 

particular ecosystem.

G1 Conservation status ranking indicating 

that an association or taxonomic unit is 

critically imperiled globally and at very high 

risk of extinction.

G2 Conservation status ranking indicating 

that an association or taxonomic unit is imperiled globally and at high risk of extinction.

G3 Conservation status ranking indicating 

that an association or taxonomic unit is 

vulnerable globally and at moderate risk 

of extinction.

Geographic range The geographic area 

occupied by a species, also known as the 

species range. 

Hemicryptophyte A perennial plant with 

reproductive structures located at the soil 

surface.

Hydroperiod The amount of time a wetland 

is covered by or saturated with water.

Hydrophytic Vegetation that thrives in 

saturated or inundated conditions.

Insular ecosystem An ecosystem characterized by (1) geographic discreteness, 

(2) narrow geographic distribution tied to 

geology, elevation, or similar environmental 

factors, (3) biogeographic endemism or 

disjunction, (4) high levels of abiotic stress 

factors, and (5) steep environmental and 

ecological boundaries.
148 Insular Ecosystems of the Southeastern United States

Life-form spectra Plant classifications based 

on position of perennating tissue relative to 

the ground surface.

Life history The sequence of events related 

to an organism’s survival and reproduction.

Mesic Environmental conditions characterized by moderate amounts of moisture, 

as well as species adapted to those environmental conditions.

Obligate heliophyte A plant that requires 

high levels of light and is susceptible to 

reduced fitness in shaded conditions.

Phenophase An observable phase in the 

annual life cycle of an organism or species.

Phenology The seasonal cycle of phenophases and their relationships with environmental conditions.

Population A group of individuals of the 

same species inhabiting the same geographic 

area at the same time.

Richness The number of distinct species

in an ecological community.

S1 Conservation status ranking indicating 

that an association or taxonomic unit is 

critically imperiled at the state level.

S2 Conservation status ranking indicating

that an association or taxonomic unit is 

imperiled at the state level.

S3 Conservation status ranking indicating 

that an association or taxonomic unit is 

vulnerable at the state level.

Saturation A hydrologic condition whereby 

soil pore space is completely filled by water.

Seepage The slow discharge of groundwater at the land surface.

Seral stage An intermediate community 

as an ecosystem progresses toward climax 

in the process of succession.

Stress regime The patterns of variation 

and interaction of abiotic stress factors 

influencing ecological processes such as 

competition and succession.

Succession The evolution over time of 

ecosystem structure and composition.

Taxa Taxonomic groups of any rank, such 

as varieties, sub-species, species, genera, 

families, or classes.

Therophyte An annual plant that survives 

unfavorable seasons in the form of seeds only.

Woody encroachment The process by 

which woody plants such as shrubs and 

trees colonize areas previously occupied 

by herbaceous plants.

Xeric Environmental conditions characterized by lack of soil moisture.
Appendix 1 149

Appendix 1. Ecological System Names According to the International Terrestrial 

Ecological Systems Classification 

The International Terrestrial Ecological Systems 

Classification (ITESC) is a system used for the classification and mapping of terrestrial ecosystems (NatureServe, 

2003). The ecosystem categories described in the preceding 

chapters all generally correspond to one or more ecological systems according to the ITESC (table 1–1). In some 

cases, lists of ITESC systems are not exhaustive because of 

imperfect correspondence between ITESC system classification and ecosystem conceptualization by previous studies. 

In some cases, only a certain subset of vegetation communities associated with a particular ITESC system are associated 

with the ecological system described in this report. Detailed 

descriptions of ITESC systems are available in NatureServe 

Explorer (NatureServe, 2015).

Table 1–1. Selected insular ecosystems of the southeastern United States, with selected corresponding ecological system 

names according to the International Terrestrial Ecological Systems Classification (ITESC).

Ecological system ITESC ecological system name(s) ITESC identifier

Piedmont granite outcrops Southern Piedmont Granite Flatrock and Outcrop CES202.329

Limestone cedar glades Nashville Basin Limestone Glade and Woodland CES202.334

Xeric limestone prairies Ozark Prairie and Woodland (subset) CES202.326

Western Highland Rim Prairie and Barrens (subset) CES202.352

Eastern Highland Rim Prairie and Barrens (subset) CES202.354

Pennyroyal Karst Plain Prairie and Barrens (subset) CES202.355

Central Interior Highlands Calcareous Glade and Barrens (subset) CES202.691

Southern Ridge and Valley Calcareous Glade and Woodland (subset) CES202.024

Alabama Ketona Glade and Woodland CES202.338

Mid-Appalachian shale barrens Appalachian Shale Barrens CES202.598

High-elevation outcrops and balds Southern Appalachian Grass and Shrub Bald CES202.294

Southern Appalachian Granitic Dome (subset) CES202.297

Southern Appalachian Rocky Summit (subset) CES202.327

Carolina bays Atlantic Coastal Plain Clay-Based Carolina Bay Wetland CES203.245

Atlantic Coastal Plain Peatland Pocosin and Canebrake (subset) CES203.267

Karst-depression wetlands Central Interior Highlands and Appalachian Sinkhole and Depression Pond CES202.018

East Gulf Coastal Plain Depression Pond (subset) CES203.558

Southern Atlantic Coastal Plain Depression Pond (subset) CES203.262

Riverscour ecosystems Cumberland Riverscour CES202.036

Southern Piedmont Small Floodplain and Riparian Forest (subset) CES202.323

Southern Piedmont Large Floodplain Forest (subset) CES202.324

Central Appalachian River Floodplain (subset) CES202.608

Central Appalachian Stream and Riparian (subset) CES202.609

South-Central Interior Large Floodplain Forest (subset) CES202.705

South-Central Interior Small Stream and Riparian (subset) CES202.706
150 Insular Ecosystems of the Southeastern United States

Appendix 2. Component Associations According to the International Terrestrial 

Ecological Systems Classification 

Ecological systems are coarse-level categories under 

the International Terrestrial Ecological Systems Classification (ITESC). For each ecological system, component 

plant associations are finer-scale categories (NatureServe, 

2003). For each of the ecosystem categories described in the 

preceding chapters, table 2–1 lists selected plant associations 

of conservation concern based on the ITESC. Component 

associations are included if they are ranked as critically 

imperiled globally (G1), imperiled globally (G2), or vulnerable globally (G3), according to NatureServe (2015).

Table 2–1. Selected plant associations of conservation concern in insular ecosystems of the southeastern United States, with 

association names and identifiers from the International Terrestrial Ecological Systems Classification (ITESC).—Continued

[G1, critically imperiled globally; G2, imperiled globally; G3, vulnerable globally; all rankings G1–G3 are based on NatureServe (2015)]

Ecological system ITESC association name ITESC ID

Piedmont granite outcrops Virginia Piedmont Granitic Flatrock Glade (G2) CEGL003857

Pine/Granitic Flatrock Border Woodland (G3) CEGL003993

Granite Flatrock Complex, Perennial Zone (G3) CEGL004298

Piedmont Granitic Flatrock Glade, Seasonal Pool (G1) CEGL004342

Piedmont Granitic Flatrock Glade, Annual Succulent Zone (G3) CEGL004344

Limestone cedar glades Nashville Basin Post Oak Woodland (G2) CEGL003712

Red-cedar–Blue Ash Limestone Woodland (G3) CEGL003754

Southern Limestone Glade Margin Shrubland (G3) CEGL003938

Limestone Seep Glade (G2) CEGL004169

Limestone Glade Streamside Meadow (G2) CEGL004292

Limestone Annual Grass Glade (G3) CEGL004340

Interior Low Plateaus Limestone Glade Ephemeral Pool (G3) CEGL004346

Kentucky Glade Seep (G3) CEGL004669

Central Limestone Glade (G2) CEGL005131

Outer Bluegrass Dolomite Glade (G1) CEGL007772

Xeric limestone prairies Chinkapin Oak–Ash/Little Bluestem Woodland (G3) CEGL002143

Ozark Limestone Glade (G2) CEGL002251

Ozark Dolomite Glade (G3) CEGL002398

Dry Coosa Valley Barrens (G1) CEGL004045

Highland Rim Dry-Mesic Prairie (G1) CEGL004063

Post Oak–White Oak Dry-Mesic Barrens (G1) CEGL004217

Southern Ridge and Valley Annual Grass Glade (G2) CEGL004339

Interior Low Plateaus Limestone Glade Ephemeral Pool (G3) CEGL004346
Appendix 2 151

Table 2–1. Selected plant associations of conservation concern in insular ecosystems of the southeastern United States, with 

association names and identifiers from the International Terrestrial Ecological Systems Classification (ITESC).—Continued

[G1, critically imperiled globally; G2, imperiled globally; G3, vulnerable globally; all rankings G1–G3 are based on NatureServe (2015)]

Ecological system ITESC association name ITESC ID

Xeric limestone prairies—

Continued

Bluegrass Cat Prairie (G1) CEGL004464

Kentucky Glade Seep (G3) CEGL004669

Western Highland Rim Post Oak Barrens (G2) CEGL004686

Moulton and Tennessee Valley Limestone Hill Barrens (G2) CEGL004738

Western Highland Rim Blackjack Oak Barrens (G2) CEGL004756

Southeastern Highland Rim Blackjack Oak Barrens (G2) CEGL004922

Central Limestone Glade (G2) CEGL005131

Central Dry-Mesic Limestone–Dolomite Prairie (G1) CEGL005280

Chinkapin Oak Limestone–Dolomite Savanna (G2) CEGL005284

Chinkapin Oak/Prairie Willow/Rattlesnake-master Woodland (G1) CEGL006239

Outer Bluegrass Dolomite Glade (G1) CEGL007772

Arkansas Cherokee Prairie, Dry Phase (G2) CEGL007827

Ozark Ashe’s Juniper Glade Woodland (G2) CEGL007833

Dry Calcareous Forest/Woodland (White Ash–Shagbark Hickory Type) (G1) CEGL008458

Ozark Annual Grass Glade (G3) CEGL008563

Arkansas Cherokee Prairie, Xeric Phase (G1) CEGL008564

Mid-Appalachian 

shale barrens

Blue Ridge Acidic Shale Woodland (G2) CEGL003624

Central Appalachian Circumneutral Barrens (G2) CEGL006037

North-Central Appalachian Acidic Shale Woodland (G3) CEGL006288

Chestnut Oak–Virginia Pine/Ragwort Acidic Shale Woodland (Southern Type) (G3) CEGL006562

Blue Ridge Calcareous Shale Slope Woodland (G2) CEGL007720

Central Appalachian Virginia Pine/Sparse Herbs Shale Woodland (G3) CEGL008525

Central Appalachian Chestnut Oak/Mixed Herbs Shale Woodland (G3) CEGL008526

Central Appalachian Shale Ridge Prairie/Bald (G2) CEGL008530

High-elevation outcrops 

and balds

Southern Appalachian High-Elevation Rocky Summit (Little Bluestem Type) (G1) CEGL004074

Southern Appalachian High-Elevation Rocky Summit (High Peak Type) (G1) CEGL004277

Southern Appalachian High-Elevation Rocky Summit (Typic Type) (G2) CEGL004279

Southern Appalachian Spikemoss Granitic Dome (G2) CEGL004283

High-Elevation Granitic Dome (High Peak Lichen Type) (G2) CEGL004386

Southern Appalachian Biltmore Sedge Granitic Dome (G2) CEGL004523

Southern Blue Ridge Quartzite Ledge (G1) CEGL007010

High-Elevation Outcrop Barrens (Black Chokeberry Igneous/Metamorphic Type) (G1) CEGL008508

High-Elevation Outcrop Barrens (Greenland Stitchwort Igneous/Metamorphic Type) (G1) CEGL008509
152 Insular Ecosystems of the Southeastern United States

Table 2–1. Selected plant associations of conservation concern in insular ecosystems of the southeastern United States, with 

association names and identifiers from the International Terrestrial Ecological Systems Classification (ITESC).—Continued

[G1, critically imperiled globally; G2, imperiled globally; G3, vulnerable globally; all rankings G1–G3 are based on NatureServe (2015)]

Ecological system ITESC association name ITESC ID

Carolina bays Carolina Longleaf Pine/Mixed Scrub Oak Sandhill (G3) CEGL003591

Pond Pine–Loblolly-bay/Shining Fetterbush Woodland (G3) CEGL003671

Atlantic Coastal Plain Pond-cypress Savanna (G2) CEGL003733

Coastal Plain Small Depression Swamp (Pocosin Shrub Type) (G2) CEGL003734

Small Depression Pond (Cutgrass Prairie Type) (G2) CEGL004047

Small Depression Drawdown Meadow (Typic Subtype) (G2) CEGL004105

Small Depression Pond (Typic Marsh Subtype) (G3) CEGL004127

Pond-cypress Savanna (G2) CEGL004441

Peatland Atlantic White-cedar Forest (G2) CEGL006146

Coastal Plain Depression Swamp (Mixed Subtype) (G3) CEGL007420

Karst-depression wetlands Sinkhole Pond Marsh (G3) CEGL002413

Pin Oak Mixed Hardwood Depression Forest (G3) CEGL002432

Interior Highland Maidencane Pond (G1) CEGL004126

Interior Oak–Swamp Cottonwood Pond Forest (G2) CEGL004421

Bluegrass Basin Swamp White Oak Forest (G1) CEGL004422

Water Tupelo Sinkhole Pond Swamp (G1) CEGL004712

Highland Rim Pond (Woolgrass Bulrush–Threeway Sedge Type) (G2) CEGL004719

Buttonbush Sinkhole Pond Swamp (G3) CEGL004742

Depression Pond (Spikerush–Creeping Rush Type) (G2) CEGL004748

Highland Rim Pond (Narrow Plumegrass Type) (G2) CEGL004750

Sinking Pond Overcup Oak Swamp (G1) CEGL004975

Highland Rim Pond (Pickerelweed–Arrowhead Type) (G1) CEGL004986

Highland Rim Barrens Depression Willow Oak Forest (G2) CEGL007364

Highland Rim Upland Depression Flatwoods (G1) CEGL007405

Swamp Tupelo Sagpond Forest (G1) CEGL004116

Maple Flats Barratt’s Sedge Marsh (G1) CEGL007857

Shenandoah Valley Sinkhole Pond (Typic Type) (G1) CEGL007858

Goldenclub Pond (G1) CEGL007859

Moulton Valley Buttonbush Pond (G1) CEGL008439

Montane Herbaceous Pond (Aquatic Sedge –Threeway Sedge Type) (G1) CEGL008542

Coastal Plain Vernal Pool Depression (G2) CEGL004100

Coastal Plain Limesink Pond (Rosette Grass Type) (G2) CEGL004105

Atlantic Coastal Plain Maidencane Limesink Pond (G3) CEGL004127

Coastal Plain Beaksedge Depression (G3) CEGL004132

Coastal Plain Limesink Pond (Threadleaf Beaksedge Type) (G2) CEGL004131

Buttonbush Sinkhole Pond Swamp (G3) CEGL004742

Southeastern Coastal Plain Pond Herbaceous Vegetation (Plumegrass Type) (G2) CEGL004752
Appendix 2 153

Table 2–1. Selected plant associations of conservation concern in insular ecosystems of the southeastern United States, with 

association names and identifiers from the International Terrestrial Ecological Systems Classification (ITESC).—Continued

[G1, critically imperiled globally; G2, imperiled globally; G3, vulnerable globally; all rankings G1–G3 are based on NatureServe (2015)]

Ecological system ITESC association name ITESC ID

Riverscour ecosystems Appalachian/Cumberland Sycamore-Birch Riverscour Woodland (G3) CEGL003725

Rocky Bar and Shore (Alder-Yellowroot Type) (G3) CEGL003895

Riverside Rock Outcrop and Prairie Complex (G3) CEGL004033

Rocky Bar and Shore (Twisted Sedge Type) (G3) CEGL004103

Cahaba Riverwash Prairie (G1) CEGL004149

Rocky-shoal Spiderlily–American Water-willow Herbaceous Vegetation (G1) CEGL004285

Eel-grass River Channel (G3) CEGL004333

Yadkin River Mafic Scour Prairie (G1) CEGL004459

Duck River Scour Prairie (G2) CEGL004739

Midwestern Brush Prairie Gravel Wash (G2) CEGL005175

River Birch Low Floodplain Forest (G3) CEGL006184

Piedmont/Central Appalachian River-Scour Woodland (G1) CEGL006218

Fall-line Riverwash Bedrock Prairie (G3) CEGL006283

Central Appalachian/Piedmont Bedrock Floodplain Woodland (G2) CEGL006476

Ridge and Valley Gravel-Wash Prairie (G2) CEGL006477

Piedmont/Central Appalachian Riverside Outcrop Prairie (G1) CEGL006478

Piedmont Riverscour Shrubland (G1) CEGL006484

Potomac Gorge Riverside Outcrop Barrens (G2) CEGL006491

Acidic Sandstone Appalachian Rivershore Prairie (G2) CEGL006623

Appalachian/Northern Piedmont Riverside Outcrop Woodland (G1) CEGL008449

Hiwassee/Ocoee River Boulder Scour Vegetation (G2) CEGL008455

Cumberland Riverside Scour Prairie (G2) CEGL008471

Bushy St. John’s-wort–Hazel Alder/Alabama Warbonnet–Tennessee 

Yellow-eyed-grass Shrubland (G1)

CEGL008494

Alabama St. John’s-wort–Alder Shrubland (G1) CEGL008495

Hiwassee/Ocoee Bedrock Scour Vegetation (G2) CEGL008496

Barbara’s-buttons Riverscour Prairie (G1) CEGL006598
154 Insular Ecosystems of the Southeastern United States

Appendix 3. Selected Plant Taxa of Conservation Concern in Insular Ecosystems

of the Southeastern United States

Tables 3–1 through 3–8 list selected globally rare plant 

taxa that have been documented in the insular ecosystems 

described in the preceding chapters. Scientific names are in 

accordance with those used by NatureServe (2015). Naming 

revisions, reclassifications, synonyms, and basionyms are 

available in the International Plant Names Index (2015), 

the Integrated Taxonomic Information System (2015), and 

Natural Resources Conservation Service (2015). 

As discussed in the preceding chapters, consensus lists 

of endemic and biogeographically disjunct taxa are generally 

lacking for these ecosystems. Thus, notations of endemic and 

disjunct taxa in these tables are based on a preponderance 

of evidence from the cited publications. Conservation status 

information is based on global rankings by NatureServe 

(2015). For details on methodology for conservation status 

ranking assignments, see Faber-Langendoen and others 

(2012). In cases where conservation status ranks were 

uncertain or straddled two rankings (for example, a G3/G2 

listed species), the more conservative (lower numbered) 

status ranking was used (G2 in this example).

Table 3–1. Selected plant taxa of conservation concern in Piedmont granite outcrops. 

[All global rankings are based on NatureServe (2015). Species names denoted by * indicate recent reclassifications (Mason 

Brock, Austin Peay State University, written commun., 2015). Reference: 1, Shure (1999); 2, Burbanck and Platt (1964); 

3, McVaugh (1943); 4, Belden and others (2003); 5, Baskin and Baskin (1988); 6, Wyatt and Fowler (1977); 7, Spaulding 

(2013); 8, Quarterman and others (1993); 9, Matthews and Murdy (1969). Endemic/disjunct: D, biogeographic disjunction—

indicating that occurrences within this ecosystem are disjunct from primary ranges in other regions or represent substantial 

range expansions (see Brown and others,1996); E, endemic taxa primarily restricted to the ecosystem (see Estill and Cruzan, 

2001); NE, near-endemic (primarily restricted to this ecosystem and to related ecosystems in similar contexts); —, not endemic 

or disjunct. Conservation status: G1, critically imperiled globally; G2, imperiled globally; G3, vulnerable globally; LE, listed 

as endangered under the U.S. Endangered Species Act; LT, listed as threatened under the U.S. Endangered Species Act]

Species name Reference Endemic/disjunct Conservation status

Amphianthus pusillus* 1, 2, 3, 8 E G2, LT

Cuscuta harperi 7 NE G2

Cyperus granitophilus 1, 2, 3, 4, 5, 8 E G3

Draba aprica* 8 D G3

Isoetes melanospora 1, 3, 6, 8, 9 E G1, LE

Isoetes tegetiformans 1, 8, 9 E G1, LE

Portulaca smallii 1, 2, 3, 4, 6, 8 E G3

Pycnanthemum curvipes — G3

Quercus georgiana 1, 3, 8 E G3

Rhus michauxii — G2, LE

Rhynchospora saxicola 1, 3 NE G3

Sedum pusillum 1, 3, 6, 8 E G3

Talinum mengesii* 5 NE G3
Appendix 3 155

Table 3–2. Selected plant taxa of conservation concern in limestone cedar glades. 

[All global rankings are based on NatureServe (2015). Sub-species or variety; names denoted by * indicate recent reclassifications 

(Mason Brock, Austin Peay State University, written commun., 2015). Reference: 1, Quarterman (1950); 2, Baskin and others (1968); 

3, Baskin and Baskin (1989); 4, Baskin and Baskin (1999); 5, Quarterman and others (1993); 6, Baskin and Baskin (2003). Endemic/

disjunct: E, endemic taxa primarily restricted to the ecosystem (see Estill and Cruzan, 2001); NE, near-endemic (primarily restricted 

to this ecosystem and to related ecosystems in similar contexts); NE (LG/XLP), near endemic (primarily restricted to limestone cedar 

glades and xeric limestone prairies); —, not endemic or disjunct. Conservation status: G1, critically imperiled globally; G2, imperiled 

globally; G3, vulnerable globally; LE, listed as endangered under the U.S. Endangered Species Act; LT, listed as threatened under the 

U.S. Endangered Species Act; PE, proposed endangered under the U.S. Endangered Species Act; PT, proposed threatened under the 

U.S. Endangered Species Act; T1, critically imperiled sub-species or variety; T3, vulnerable]

Species name Reference Endemic/disjunct Conservation status

Astragalus bibullatus 3, 4, 6 E G1, LE

Astragalus tennesseensis 2, 3, 4, 5, 6 NE G3

Dalea foliosa 1, 2, 3, 4, 6 NE G2, LE

Dalea gattingeri 1, 2, 3, 4, 5, 6 NE (LG/XLP) G3

Delphinium alabamicum 4, 6 E G2

Echinacea tennesseensis 2, 3, 4, 5, 6 NE (LG/XLP) G2

Leavenworthia alabamica 2, 4, 5, 6 NE (LG/XLP) G2

Leavenworthia crassa 2, 4, 5, 6 NE G2, PE

Leavenworthia exigua var. exigua 2, 3, 4, 5, 6 NE (LG/XLP) T3

Leavenworthia exigua var. laciniata 2, 4, 5, 6 NE (LG/XLP) T1, PT

Leavenworthia exigua var. lutea 2, 3, 4, 5, 6 NE (LG/XLP) T1

Lesquerella lyrata* 2, 4, 6 NE G1, LT

Onosmodium molle ssp. molle* 2, 3, 4, 5, 6 NE (LG/XLP) T3

Phacelia dubia var. interior 2, 3, 4, 6 E T3

Phlox bifida ssp. stellaria 5, 6 — T3

Solidago gattingeri 3, 6 — G3

Talinum calcaricum* 2, 3, 4, 5, 6 NE G3

Trifolium calcaricum 4, 6 NE (LG/XLP) G1
156 Insular Ecosystems of the Southeastern United States

Table 3–3. Selected plant taxa of conservation concern in xeric limestone prairies. 

[Sll global rankings are based on NatureServe (2015). Species names denoted by * indicate recent reclassifications (Mason Brock, 

Austin Peay State University, written commun., 2015). Reference: 1, Lawless and others (2006); 2, Allison and Stevens (2001); 

3, Nelson and others (2013); 4, Collins and Wieboldt (1992); 5, Webb and others (1997). Endemic/disjunct: E (KDG), endemic to the 

Ketona dolomite glades of Bibb County, Alabama; NE, near-endemic (primarily restricted to this ecosystem and to related ecosystems 

in similar contexts); NE (KDG), near-endemic to the Ketona dolomite glades of Bibb County, Alabama; NE (LG/XLP), near endemic 

(primarily restricted to limestone cedar glades and xeric limestone prairies); —, not endemic or disjunct. Conservation status: C, candidate for listing under the U.S. Endangered Species Act; G1, critically imperiled globally; G2, imperiled globally; G3, vulnerable 

globally; LE, listed as endangered under the U.S. Endangered Species Act; LT, listed as threatened under the U.S. Endangered Species 

Act; PT, proposed threatened under the U.S. Endangered Species Act; T1, critically imperiled sub-species or variety; T2, imperiled 

sub-species or variety; T3, vulnerable sub-species or variety]

Species name Reference Endemic/disjunct Conservation status

Agalinis auriculata 3 — G3

Arabis georgiana 2 — G1, PT

Astragalus tennesseensis 2 NE G3

Carex juniperorum 1, 3 — G3

Castilleja kraliana 1, 2 E (KDG) G2

Coreopsis grandiflora var. inclinata 1, 2 E (KDG) T2

Croton alabamensis 2 — G3

Dalea cahaba 1, 2 E (KDG) G2

Dalea gattingeri 3, 5 NE (LG/XLP) G3

Delphinium treleasei 1, 3 — G3

Echinacea paradoxa var. paradoxa 1, 3 — G2

Echinacea tennesseensis 3 NE (LG/XLP) G2

Erigeron strigosus var. dolomiticola 1, 2 E (KDG) T2

Helianthus eggertii 3 — G3

Helianthus smithii 2 — G2

Leavenworthia alabamica 3, 5 NE (LG/XLP) G2

Leavenworthia exigua var. exigua 2, 3 NE (LG/XLP) T3

Leavenworthia exigua var. laciniata 3 NE (LG/XLP) T1, PT

Leavenworthia exigua var. lutea 3 NE (LG/XLP) T1

Liatris oligocephala 1, 2 E (KDG) G1

Marshallia mohrii 2 — G3, LT

Onosmodium decipiens* 1, 2 E (KDG) G2

Onosmodium molle ssp. molle* 1, 3, 5 NE (LG/XLP) T3

Phlox pulchra 2 — G2

Prenanthes barbata 3 — G3

Scutellaria alabamensis 2 NE (KDG) G2

Sedum nevii 2 — G3

Silene regia 2 — G3

Silphium glutinosum 1, 2 E (KDG) G2

Silphium pinnatifidum 3, 5 — G3

Solanum carolinense var. hirsutum 2 — T1

Solidago gattingeri 1, 3 — G3

Spigelia gentianoides 1, 2 E (KDG) G1, LE

Symphyotrichum georgianum 2 — G3, C

Talinum calcaricum 1, 3 NE G3

Trifolium calcaricum 4 NE (LG/XLP) G1
Appendix 3 157

Table 3–4. Selected plant taxa of conservation concern in mid-Appalachian shale barrens. 

[All global rankings are based on NatureServe (2015). Species names denoted by * indicate recent reclassifications (Mason 

Brock, Austin Peay State University, written commun., 2015). Reference: 1, Platt (1951); 2, Keener (1983); 3, Artz (1948); 

4, Braunschweig and others (1999); 5, Clarkson (1966); 6, U.S. Forest Service (2011). Endemic/disjunct: E, endemic taxa 

primarily restricted to the ecosystem (see Estill and Cruzan, 2001); NE, near-endemic (primarily restricted to this ecosystem 

and to related ecosystems in similar contexts). Conservation status: G1, critically imperiled globally; G2, imperiled globally; 

G3, vulnerable globally; LE, listed as endangered under the U.S. Endangered Species Act]

Species name Reference Endemic/disjunct Conservation status

Allium oxyphilum 1, 2, 4, 6 NE G2

Arabis serotina* 1, 2, 4, 5, 6 E G2, LE

Clematis coactilis 1, 2, 4, 6 NE G3

Clematis viticaulis 1, 2, 3, 4, 6 E G1

Oenothera argillicola 1, 2, 3, 4, 5, 6 E G3

Phlox buckleyi 1, 2, 4, 5, 6 NE G2

Taenidia montana 1, 2, 5, 6 NE G3

Trifolium virginicum 1, 2, 3, 4, 5, 6 NE G3

Table 3–5. Selected plant taxa of conservation concern in high-elevation outcrops and balds. 

[All global rankings are based on NatureServe (2015). Species names denoted by * indicate recent reclassifications (Mason 

Brock, Austin Peay State University, written commun., 2015). Reference: 1, Wiser (1994); 2, White and Sutter (1999); 3, Wiser 

and White (1999). Endemic/disjunct: E, endemic taxa primarily restricted to the ecosystem (see Estill and Cruzan, 2001); NE, 

near-endemic (primarily restricted to this ecosystem and to related ecosystems in similar contexts); —, not endemic or disjunct. 

Conservation status: G1, critically imperiled globally; G2, imperiled globally; G3, vulnerable globally LE, listed as endangered 

under the U.S. Endangered Species Act; LT, listed as threatened under the U.S. Endangered Species Act]

Species name Reference Endemic/disjunct Conservation status

Abies fraseri 3 — G2

Calamagrostis cainii 3 E G1

Carex biltmoreana 1, 3 NE G3

Carex misera 1, 2, 3 E G3

Carex ruthii 2, 3 — G3

Delphinium exaltatum 2 — G3

Geum geniculatum 2 E G1

Geum radiatum 1, 2, 3 E G2, LE

Glyceria nubigena 2, 3 — G2

Houstonia purpurea var. montana 1, 2, 3 E T2, LE

Hypericum buckleii 1, 3 E G3

Hypericum graveolens 2, 3 — G3

Hypericum mitchellianum 2, 3 — G3

Krigia montana 1, 2, 3 NE G3

Liatris helleri 1, 3 NE G2, LT

Lilium grayi 2 — G3

Prenanthes roanensis* 2, 3 — G3

Rhododendron vaseyi 2, 3 — G3

Rugelia nudicaulis 2, 3 NE G3

Selaginella tortipila* 1, 3 — G3

Packera millefolia * 1, 3 E G2

Solidago spithamaea 1, 3 E G2, LT

Stachys clingmanii 2 NE G2
158 Insular Ecosystems of the Southeastern United States

Table 3–6. Selected plant taxa of conservation concern in Carolina bays. 

[All global rankings are based on NatureServe (2015). Species names denoted by * indicate recent reclassifications (Mason Brock, 

Austin Peay State University, written commun., 2015). Reference: 1, De Steven and Toner (1997); 2, South Carolina Department

of Natural Resources (1999); 3, NatureServe Explorer Ecological System Comprehensive Report for CES203.245: Atlantic 

Coastal Plain Clay-Based Carolina Bay Wetland (NatureServe, 2015); 4, Bennett and Nelson (1991). Endemic/disjunct: NE,

near-endemic (primarily restricted to this ecosystem and to related ecosystems in similar contexts); —, not endemic or disjunct. 

Conservation status: C, candidate for listing under the U.S. Endangered Species Act; G1, critically imperiled globally; G2, imperiled globally; G3, vulnerable globally; LE, listed as endangered under the U.S. Endangered Species Act;]

Species name Reference Endemic/disjunct Conservation status

Coreopsis rosea 1, 2, 3 — G3

Croton elliottii 1, 3 — G2

Dichanthelium hirstii 3 — G1, C

Dionaea muscipula 2, 4 — G3

Echinodorus parvulus 2, 3, 4 — G3

Eupatorium paludicola 3 — G2

Hypericum harperi 2, 3, 4 — G3

Lindera melissifolia 3 — G2, LE

Litsea aestivalis 2, 3, 4 — G3

Lobelia boykinii 2, 3, 4 — G2

Ludwigia spathulata 1, 3, 4 — G2

Myriophyllum laxum 3 — G3

Oxypolis canbyi 1, 2, 3, 4 NE G2, LE

Peltandra sagittifolia 2, 3, 4 — G3

Polygonum hirsutum* 3 — G3

Rhexia aristosa 1, 2, 3, 4 — G3
Appendix 3 159

Table 3–7. Selected plant taxa of conservation concern in karst-depression wetlands of the Interior Low Plateaus,

Shenandoah Valley region of the Valley and Ridge, and the Dougherty Plain.

[All global rankings are based on NatureServe (2015). Reference: 1, Homoya and Hedge (1985). Reference: 2, Fleming and 

Van Alstine (1999); 3, Buhlmann and others (1999); 4, Sutter and Kral (1994). Region: DP, Dougherty Plain; ILP, Interior Low 

Plateaus; VR, Shenandoah Valley region of the Valley and Ridge. Endemic/disjunct: D, occurrences within this ecosystem are disjunct 

from primary ranges in other regions or represent substantial range expansions (see Brown and others, 1996); E, endemic taxa primarily restricted to the ecosystem (see Estill and Cruzan, 2001); NE, near-endemic (primarily restricted to this ecosystem and to related 

ecosystems in similar contexts); —, not endemic or disjunct. Conservation status: G1, critically imperiled globally; G2, imperiled 

globally; G3, vulnerable globally; LT, listed as threatened under the U.S. Endangered Species Act]

Species name Reference Region Endemic/disjunct Conservation status

Carex decomposita 1 ILP D G3

Helenium virginicum 2, 3 VR D G3, LT

Helonias bullata 2, 3 VR D G3, LT

Hypericum lissophloeus 4 DP E G2

Isoetes virginica 2 VR NE G1

Rhexia salicifolia 4 DP NE G2

Xyris longisepala 4 DP NE G2

Xyris scabrifolia 4 DP — G3

Table 3–8. Selected plant taxa of conservation concern in riverscour ecosystems. 

[All global rankings are based on NatureServe (2015). Reference: 1, Vanderhorst and others (2007); 2, Vanderhorst and others (2010); 

3, Schmalzer (1989); 4, Shaw and Wofford (2003); 5, Taylor (2003); 6, Byers and others (2007); 7, Major and others (1999); 8, 

Floden (2012); 9, Campbell and Medley (1989); 10, Schafale (2012); 11, Bailey and Coe (2001); 12, Ogle (1991); 13, Bartgis (1997); 

14, Wells (2012). Endemic/disjunct: E, endemic taxa primarily restricted to the ecosystem (see Estill and Cruzan, 2001); NE, nearendemic (primarily restricted to this ecosystem and to related ecosystems in similar contexts); —, not endemic or disjunct. Conservation status: G1, critically imperiled globally; G2, imperiled globally; G3, vulnerable globally; LE, listed as endangered under the 

U.S. Endangered Species Act; LT, listed as threatened under the U.S. Endangered Species Act]

Species name Reference Endemic/disjunct Conservation status

Berberis canadensis 3, 4 — G3

Calamovilfa arcuata 3, 4, 11 E G2

Conradina verticillata 3, 4, 5, 11 E G3, LT

Eurybia saxicastellii 5, 9, 11 E G1

Fothergilla major 4 — G3

Marshallia grandiflora 2, 3, 4, 5, 6, 11 NE G2

Pityopsis ruthii 7 E G1, LE

Ptilimnium nodosum 13, 14 NE G2, LE

Pycnanthemum torrei 1 — G2

Solidago arenicola 8 E G2

Solidago plumosa 10 E G1

Spiraea virginiana 2, 3, 4, 5, 11, 12 E G2, LT

Vitis rupestris 1 NE G3
160 Insular Ecosystems of the Southeastern United States

References Cited in Appendixes

Allison, J., and Stevens, T., 2001, Vascular flora of Ketona 

dolomite outcrops in Bibb County, Alabama: Castanea, 

v. 66, no. 1, p. 154–205.

Artz, L., 1948, Plants of the shale barrens of the tributaries

of the James River in Virginia: Castanea, v. 13, no. 4, 

p. 141–145.

Bailey, C., and Coe, F., 2001, The vascular flora of the riparian zones of the Clear Fork River and the New River in 

the Big South Fork National River and Recreation Area 

(BSFNRRA): Castanea, v. 66, no. 3, p. 252–274.

Bartgis, R., 1997, The distribution of the endangered plant 

Ptilimnium nodosum (Rose) Mathias (Apiaceae) in the 

Potomac River drainage: Castanea, v. 62, no. 1, p. 55–59.

Baskin, J., and Baskin, C., 1988, Endemism in rock outcrop 

plant communities of unglaciated eastern United States—

An evaluation of the roles of the edaphic, genetic and light 

factors: Journal of Biogeography, v. 15, no. 5, p. 829–840.

Baskin, J., and Baskin, C., 1989, Cedar glade endemics in 

Tennessee, and a review of their autecology: Journal of the 

Tennessee Academy of Science, v. 64, no. 3, p. 63–74.

Baskin, J., and Baskin, C., 1999, Cedar glades of the southeastern United States, in Anderson, R., Fralish, J., and 

Baskin, J., eds., Savannas, barrens, and rock outcrop plant 

communities of North America: New York, Cambridge 

University Press, p. 206–219.

Baskin, J., and Baskin, C., 2003, The vascular flora of cedar 

glades of the southeastern United States and its phytogeographical relationships: Journal of the Torrey Botanical 

Society, v. 130, no. 2, p. 101–118.

Baskin, J., Quarterman, E., and Caudle, C., 1968, Preliminary 

check-list of the herbaceous vascular plants of cedar glades: 

Journal of the Tennessee Academy of Science, v. 43, no. 3, 

p. 65–71.

Belden, A., Chazal, A., and Hobson, C., 2003, A natural 

heritage inventory of fourteen headwater sites in the 

Dragon Run Watershed: Richmond, Va., Virginia Department of Conservation and Recreation.

Bennett, S., and Nelson, J., 1991, Distribution and status of 

Carolina bays in South Carolina: Columbia, S.C., South 

Carolina Wildlife and Marine Resources Department, 

Nongame and Heritage Trust Publications, 98 p.

Braunschweig, S., Nilson, E., and Wieboldt, T., 1999, The 

mid-Appalachian shale barrens, in Anderson, R., Fralish, J., 

and Baskin, J.M., eds., Savannas, barrens, and rock outcrop 

plant communities of North America: New York, Cambridge University Press, p. 83–98.

Brown, J., Stevens, G., and Kaufman, D., 1996, The geographic range—Size, shape, boundaries, and internal 

structure: Annual Review of Ecology and Systematics, 

v. 27, no. 1996, p. 597–623.

Buhlmann, K., Mitchell, J., and Smith, L., 1999, Descriptive 

ecology of the Shenandoah Valley sinkhole pond system in 

Virginia: Banisteria, v. 13, p. 23–51.

Burbanck, M., and Phillips, D., 1983, Evidence of plant 

succession on granite outcrops of the Georgia Piedmont: 

American Midland Naturalist, v. 109, no. 1, p. 94–104.

Burbanck, M., and Platt, R., 1964, Granite outcrop communities of the Piedmont Plateau in Georgia: Ecology, v. 45, 

no. 2, p. 292–306.

Byers, E., Vanderhorst, J., and Streets, B., 2007, Classification

and conservation assessment of high elevation wetland 

communities in the Allegheny Mountains of West Virginia: 

Elkins, W. Va., West Virginia Natural Heritage Program.

Campbell, J., and Medley, M., 1989, Aster saxicastellii

(Asteraceae), a new species from the Rockcastle River bars 

in southeastern Kentucky: Sida, v. 13, no. 3, p. 277–284.

Clarkson, R., 1966, The vascular flora of the Monongahela 

national forest, West Virginia: Castanea, v. 31, no. 1, 

p. 1–119.

Collins, J., and Wieboldt, T., 1992, Trifolium calcaricum

(Fabaceae), a new clover from limestone barrens of eastern 

United States: Castanea, v. 57, no. 4, p. 282–286.

De Steven, D., and Toner, M., 1997, Gradient analysis and 

classification of Carolina bay vegetation—A framework for 

bay wetlands conservation and restoration: New Ellenton, 

S.C., U.S. Forest Service, Savannah River Forest Station, 

62 p.

Estill, J., and Cruzan, M., 2001, Phytogeography of rare 

plant species endemic to the southeastern United States: 

Castanea, v. 66, no. 1, p. 3–23.

Faber-Langendoen, D., Nichols, J., Master, L., Snow, K., 

Tomaino, A., Bittman, R., Hammerson, G., Heidel, B., 

Ramsay, L., Teucher, A., and Young, B., 2012, NatureServe 

conservation status assessments—Methodology for assigning ranks: Arlington, Va., NatureServe.

Fleming, G., and Van Alstine, N., 1999, Plant communities 

and floristic features of sinkhole ponds and seepage wetlands in southeastern Augusta County, Virginia: Banisteria, 

v. 13, p. 67–94.

Floden, A., 2012, Notes on two rare Solidago (Asteraceae) 

in Tennessee—S. Arenicola and S. simplex: Phytoneuron, 

v. 63, p. 1–4.
Appendix 3 161

Homoya, M., and Hedge, C., 1985, The upland sinkhole 

swamps and ponds of Harrison County, Indiana: Proceedings of the Indiana Academy of Science, v. 92, p. 383–388.

International Plant Names Index, 2015, The international plant 

names index, accessed October 5, 2015, at http://www.ipni.org. 

Integrated Taxonomic Information System, 2015, The 

Integrated Taxonomic Information System, accessed 

October 5, 2015, at http://www.itis.gov.

Keener, C., 1983, Distribution and biohistory of the endemic 

flora of the mid-Appalachian shale barrens: The Botanical 

Review, v. 49, no. 1, p. 65–115.

Lawless, P., Baskin, J., and Baskin, C., 2006, Xeric limestone 

prairies of eastern United States—Review and synthesis: 

The Botanical Review, v. 72, no. 3, p. 235–272.

Major, C., Bailey, C., Donaldson, J., McCoy, R., Nordman, C.,

Williams, M., and Withers, D., 1999, An ecological inventory of selected sites in the Cherokee National Forest: 

Tennessee Department of Environment and Conservation, 

Tennessee Division of Natural Heritage.

Matthews, J., and Murdy, W., 1969, A study of Isoetes common to the granite outcrops of the southeastern Piedmont, 

United States: Botanical Gazette, v. 130, no. 1, p. 53–61.

McVaugh, R., 1943, The vegetation of the granitic flat-rocks 

of the southeastern United States: Ecological Monographs, 

v. 13, no. 2, p. 119–166.

Natural Resources Conservation Service, 2015, The USDA 

plants database: Greensboro, N.C., U.S. Department of 

Agriculture, Natural Resources Conservation Service, 

National Plant Data Team, accessed September 25, 2015, 

at http://plants.usda.gov.

NatureServe, 2003, A working classification of terrestrial 

ecological systems in the conterminous United States—

International terrestrial ecological systems classification: 

Arlington, Va., NatureServe.

NatureServe, 2015, NatureServe Explorer—An online 

encyclopedia of life [web application], ver. 7.1, accessed 

October 7, 2015, at http://www.natureserve.org/explorer.

Nelson, P., Fitzgerald, J., Larson, K., McCoy, R., Schotz, A., 

Taft, J., Witsell, T., and Yahn, B., 2013, Central hardwoods 

joint venture glade conservation assessment for the Interior 

Highlands and Interior Low Plateaus of the central hardwoods region: Central Hardwoods Joint Venture.

Ogle, D., 1991, Spiraea virginiana Britton—II. Ecology and 

species biology: Castanea, v. 56, no. 4, p. 297–303.

Platt, R., 1951, An ecological study of the mid-Appalachian 

shale barrens and of the plants endemic to them: Ecological 

Monographs, v. 21, no. 4, p. 269–300.

Quarterman, E., 1950, Ecology of cedar glades—I. 

Distribution of glade flora in Tennessee: Bulletin of 

the Torrey Botanical Club, v. 77, no. 1, p. 1–9.

Quarterman, E., Burbanck, M., and Shure, D., 1993, Rock 

outcrop communities—Limestone, sandstone, and granite, 

in Martin, W., Boyce, S., and Echternacht, A., eds., 

Biodiversity of the southeastern United States—Upland 

terrestrial communities: New York, John Wiley and Sons, 

Inc., p. 35–86.

Schafale, M., 2012, Guide to the natural communities of 

North Carolina, fourth approximation: Raleigh, N.C., 

North Carolina Natural Heritage Program, 208 p.

Schmalzer, P., 1989, Vegetation and flora of the Obed River 

gorge system, Cumberland Plateau, Tennessee: Journal of 

the Tennessee Academy of Science, v. 64, no. 3, p. 161–168.

Shaw, J., and Wofford, B., 2003, Woody plants of Big South 

Fork National River and Recreation Area, Tennessee and 

Kentucky and floristic comparison of selected Southern 

Appalachian woody: Castanea, v. 68, no. 2, p. 119–134.

Shure, D., 1999, Granite outcrops of the southeastern 

United States., in Anderson, R., Fralish, J., and Baskin, J., 

eds., Savannas, barrens, and rock outcrop plant communities of North America: New York, Cambridge University 

Press, p. 99–118.

South Carolina Department of Natural Resources, 1999, 

Advance identification of Carolina bays for South Carolina

wetlands protection: Columbia, S.C., South Carolina 

Department of Natural Resources, Land, Water and

Conservation Division, 46 p.

Spaulding, D., 2013, Key to the dodders (Cuscuta, Convolvulaceae) of Alabama and adjacent States: Phytoneuron, v. 74, 

p. 1–15.

Sutter, R., and Kral, R., 1994, The ecology, status, and 

conservation of two non-alluvial wetland communities in 

the south Atlantic and eastern Gulf coastal plain, USA: 

Biological Conservation, v. 68, p. 235–243.

Taylor, D., 2003, Viability assessment report for riparian

habitat association, Daniel Boone National Forest, 

unpublished report submitted to the National Forest Service, 

U.S. Department of Agriculture, 19 p.

U.S. Forest Service, 2011, Draft environmental impact 

statement for the revised land and resource management 

plan, George Washington National Forest: Roanoake, Va., 

U.S. Department of Agriculture, Forest Service.

Vanderhorst, J.P., Jeuck, J., and Gawler, S.C., 2007, Vegetation 

classification and mapping of New River Gorge National 

River, West Virginia: Philadelphia, Pa., U.S. National Park 

Service, technical report NPS/NER/NRTR—2007/092, 

396 p.
162 Insular Ecosystems of the Southeastern United States

Vanderhorst, J., Streets, B., Arcaro, Z., and Gawler, S., 2010, 

Vegetation classification and mapping of Gauley River 

National Recreation Area, West Virginia,: Philadelphia, Pa., 

U.S. National Park Service, technical report NPS/NER/

NRTR—2010/148, 288 p.

Webb, D., DeSelm, H., and Dennis, W., 1997, Studies of 

prairie barrens of northwestern Alabama: Castanea, v. 62, 

no. 3, p. 173–184.

Wells, E.F., 2012, Seed germination and reproductive strategies in federally endangered Harperella (Harperella nodosum Rose, Apiaceae): Castanea, v. 77, no. 3, p. 218–223.

White, P., and Sutter, R., 1999, Managing biodiversity in 

historic habitats—Case history of the southern Appalachian 

grassy balds, in Peine, J., ed., Ecosystem management for 

sustainability—Principles and practices illustrated by a 

regional biosphere reserve cooperative: Washington, D.C., 

Lewis Publishers, p. 375–395.

Wiser, S., 1994, High-elevation cliffs and outcrops of the 

southern Appalachians—Vascular plants and biogeography: 

Castanea, v. 59, no. 2, p. 85–116.

Wiser, S., and White, P., 1999, High elevation outcrops and 

barrens of the southern Appalachian mountains, in Anderson, R., Fralish, J., and Baskin, J., eds., Savannas, barrens, 

and rock outcrop plant communities of North America: 

New York, Cambridge University Press, p. 119–132.

Wyatt, R., and Fowler, N., 1977, The vascular flora and 

vegetation of the North Carolina granite outcrops: Bulletin 

of the Torrey Botanical Club, v. 104, no. 3, p. 245–253.
Manuscript approved May 20, 2016

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 Edited by Kimberly A. Waltenbaugh

 Lafayette Publishing Service Center

 Illustrations and layout by Caryl J. Wipperfurth

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please contact:

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Cartwright and Wolfe—Insular Ecosystems of the Southeastern United States—Professional Paper 1828

ISSN 1044-9612 (print)

ISSN 2330-7102 (online)

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