EPA/600/R-14/475F | January 2015 | epa.gov/research
United States
Environmental Protection
Agency
Connectivity of Streams & Wetlands
to Downstream Waters:
A Review & Synthesis of the Scientific Evidence
I
A - I
Office of Research and Development
NCEA (Washington DC, Cincinnati OH), NERL (Cincinnati OH, Las Vegas NV) and NHEERL (Corvallis OR)
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EPA/600/R-14/475F
January 2015
CONNECTIVITY OF STREAMS AND WETLANDS
TO DOWNSTREAM WATERS:
A REVIEW AND SYNTHESIS OFTHE
SCIENTIFIC EVIDENCE
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
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Errata for Connectivity of Streams and Wetlands to Downstream Waters: A Review and
Synthesis of the Scientific Evidence (EPA/600/R-14/475F)
Page 2-12, the last four lines on the page have been revised to include text that was missing
(see bolded text).
Hyporheic exchange occurs when water moves from stream or river into alluvial deposits and then
returns to the channels (Figures 2-6B and 2-6C; Bencala, 2005; Leibowitz et al., 2008). Hyporheic
exchange allows for the mixing of surface water and ground water. It occurs during both high- and
low-flow periods, and typically has relatively horizontal flowpaths at scales of meters to tens of meters
(Bencala, 2005) and vertical flowpaths with depths ranging from centimeters to tens of meters (Stanford
and Ward, 1988; Woessner, 2000 and references therein).
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DISCLAIMER
This document has been reviewed in accordance with U.S. Environmental Protection Agency policy and
approved for publication. Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis
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CONTENTS
PREFACE xii
AUTHORS AND REVIEWERS xiii
PHOTO CREDITS xix
ACKNOWLEDGMENTS xx
EXECUTIVE SUMMARY ES-1
BACKGROUND ES-1
SUMMARY OF MAJOR CONCLUSIONS ES-2
SUPPORT FOR MAJOR CONCLUSIONS ES-6
KEY FINDINGS FOR MAJOR CONCLUSIONS ES-7
CLOSING COMMENTS ES-15
CHAPTER 1. INTRODUCTION 1-1
1.1 Purpose 1-1
1.2 Scientific Context 1-2
1.2.1 Concepts of Connectivity in Hydrology and Ecology 1-2
1.2.2 Connectivity Gradients and Descriptors 1-3
1.2.3 Cumulative Effects of Streams and Wetlands on Downstream Waters 1-10
1.2.4 Effects of Human Activities on Connectivity 1-11
1.3 Report Approach 1-14
1.3.1 Selection and Screening of Scientific Materials 1-16
1.3.2 Report Structure 1-18
1.4 Summary 1-18
CHAPTER 2. AN INTEGRATED SYSTEMS PERSPECTIVE ON INTERACTIONS OF WATERSHEDS, STREAMS,
WETLANDS, AND DOWNSTREAM WATERS 2-1
2.1 Introduction 2-1
2.2 An Introduction to River Systems 2-2
2.2.1 River System Components 2-2
2.2.2 River System Hydrology 2-8
2.2.3 River Network Expansion and Contraction 2-18
2.3 Influence of Streams and Wetlands on Downstream Waters 2-22
2.3.1 Effects of Streams and Wetlands on Material Fluxes 2-22
2.3.2 Connectivity and Transport of Materials to and from Streams and Wetlands 2-26
2.3.2.1 Connectivity and Isolation 2-26
2.3.2.2 Spatial and Temporal Variability of Connectivity 2-29
2.4 Factors Influencing Connectivity 2-30
2.4.1 Climate-Watershed Characteristics 2-31
2.4.2 Spatial Distribution Patterns 2-38
2.4.3 Biota 2-40
2.4.4 Human Activities and Alterations 2-44
2.4.5 Interactions Among Factors 2-47
2.4.6 Quantifying Connectivity 2-49
2.4.6.1 Hydrologic and Chemical Connectivity 2-49
2.4.6.2 Biological Connectivity 2-50
2.4.6.3 Summary 2-51
CHAPTER 3. STREAMS: PHYSICAL, CHEMICAL, AND BIOLOGICAL CONNECTIONS TO RIVERS 3-1
3.1 Abstract 3-1
3.2 Introduction 3-2
3.3 Physical Connections 3-5
3.3.1 Water 3-5
3.3.2 Sediment 3-13
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3.3.3 Wood 3-17
3.3.4 Temperature (Heat Energy) 3-19
3.4 Chemical Connections 3-21
3.4.1 Nutrients 3-23
3.4.2 Dissolved and Particulate Organic Matter 3-28
3.4.3 Ions 3-32
3.4.4 Contaminants and Pathogens 3-33
3.5 Biological Connections 3-37
3.5.1 Invertebrates 3-38
3.5.2 Fishes 3-40
3.5.3 Genes 3-43
3.6 Streams: Synthesis and Implications 3-45
CHAPTER 4. WETLANDS: PHYSICAL, CHEMICAL, AND BIOLOGICAL CONNECTIONS TO RIVERS 4-1
4.1 Abstract 4-1
4.2 Introduction 4-3
4.3 Riparian/Floodplain Wetlands 4-3
4.3.1 Introduction 4-3
4.3.2 The Physical Influence of Riparian Areas on Streams 4-7
4.3.2.1 Hydrology 4-7
4.3.2.2 Geomorphology (Sediment-vegetation Interactions) 4-8
4.3.2.3 Temperature and Sunlight 4-9
4.3.3 The Chemical-nutrient Influence of Riparian Areas on Streams 4-10
4.3.3.1 Hyporheic/Soil Processing of Nutrients 4-11
4.3.3.2 Nitrogen 4-11
4.3.3.3 Phosphorus 4-12
4.3.3.4 Carbon and Allochthonous Inputs 4-13
4.3.3.5 Pesticides 4-14
4.3.3.6 Mercury 4-14
4.3.4 Biological Connections Between Riparian Areas and Streams 4-15
4.3.4.1 Vascular Plants and Phytoplankton 4-15
4.3.4.2 Vertebrates 4-17
4.3.4.3 Invertebrates 4-19
4.4 Non-floodplain Wetlands 4-20
4.4.1 Introduction 4-20
4.4.2 The Physical Influence of Non-floodplain Wetlands on Streams 4-21
4.4.2.1 Surface-water Connections 4-21
4.4.2.2 Ground-water Connections 4-22
4.4.2.3 Effects of Non-floodplain Wetlands on Streamflow 4-24
4.4.3 Effects of Non-floodplain Wetlands on Water Quality 4-26
4.4.3.1 Non-floodplain Wetlands as Sources for Downstream Waters 4-27
4.4.3.2 Non-floodplain Wetlands as Sinks and Transformers for
Downstream Waters 4-29
4.4.4 Biological Connections Between Non-floodplain Wetlands and Streams 4-30
4.4.5 Geographic Isolation of Non-floodplain Wetlands 4-35
4.5 Wetlands: Synthesis and Implications 4-39
4.5.1 Riparian/Floodplain Wetlands 4-39
4.5.2 Non-floodplain Wetlands 4-39
CHAPTER 5. APPLICATIONS AND DISCUSSION: CONNECTIVITY CASE STUDIES 5-1
5.1 Introduction 5-1
5.2 Carolina and Delmarva Bays 5-2
5.2.1 Connectivity and Consequences on Downstream Waters 5-2
5.2.2 Effects of Human Alteration 5-3
5.3 Oxbow Lakes 5-3
5.3.1 Connectivity and Consequences on Downstream Waters 5-3
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5.3.2 Effects of Human Alteration 5-4
5.4 Prairie Potholes 5-4
5.4.1 Connectivity and Consequences on Downstream Waters 5-4
5.4.2 Effects of Human Alteration 5-5
5.5 Prairie Streams 5-6
5.5.1 Connectivity and Consequences on Downstream Waters 5-6
5.5.2 Effects of Human Alteration 5-7
5.6 Southwestern Intermittent and Ephemeral Streams 5-7
5.6.1 Connectivity and Consequences on Downstream Waters 5-7
5.6.2 Effects of Human Alteration 5-8
5.7 Vernal Pools 5-8
5.7.1 Connectivity and Consequences on Downstream Waters 5-8
5.7.2 Effects of Human Alteration 5-9
5.8 Synthesis 5-9
CHAPTER 6. CONCLUSIONS 6-1
6.1 Major Conclusions and Key Findings 6-1
6.1.1 Conclusion 1: Streams 6-1
6.1.1.1 Conclusion 1, Key Findings 6-2
6.1.2 Conclusion 2: Riparian/Floodplain Wetlands and Open Waters 6-3
6.1.2.1 Conclusion 2, Key Findings 6-4
6.1.3 Conclusion 3: Non-floodplain Wetlands and Open Waters 6-5
6.1.3.1 Conclusion 3, Key Findings 6-6
6.1.4 Conclusion 4: Degrees and Determinants of Connectivity 6-8
6.1.4.1 Conclusion 4, Key Findings 6-9
6.1.5 Conclusion 5: Cumulative Effects 6-10
6.1.5.1 Conclusion 5, Key Findings 6-11
6.2 Strength of Evidence for Conclusions and Data Gaps in the Available Literature 6-12
CHAPTER 7. REFERENCES 7-1
APPENDIX A. GLOSSARY A-l
References A-15
APPENDIX B. CASE STUDIES B-l
B.I Case Study: Carolina and Delmarva Bays B-l
B.I.I Abstract B-l
B.1.2 Introduction B-l
B.1.2.1 Definition and Geographic Extent B-l
B.1.2.2 Geology B-2
B.l.2.3 Hydrology B-3
B.1.2.4 Water Chemistry B-3
B.1.2.5 Biological Communities B-4
B.I.3 Evidence of Connectivity B-5
B.1.3.1 Physical Connections B-5
B.l.3.2 Chemical Connections B-6
B.l.3.3 Biological Connections B-6
B.1.4 Carolina and Delmarva Bays: Synthesis and Implications B-7
B.2 Case Study: Oxbow Lakes B-8
B.2.1 Abstract B-8
B.2.2 Introduction B-8
B.2.2.1 Origin and Description B-8
B.2.3 Evidence B-9
B.2.3.1 Physical Connections B-9
B.2.3.2 Chemical Connections B-10
B.2.3.3 Biological Connections B-ll
B.2.4 Oxbow Lakes: Synthesis and Implications B-13
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B.3 Case Study: Prairie Potholes B-14
B.3.1 Abstract B-14
B.3.2 Introduction B-14
B.3.2.1 Hydrologic Dynamics B-15
B.3.2.2 Chemical Functions B-16
B.3.2.3 Ecological Characteristics B-16
B.3.3 Evidence B-17
B.3.3.1 Physical Connections B-17
B.3.3.2 Chemical Connections B-19
B.3.3.3 Biological Connections B-20
B.3.4 Prairie Potholes: Synthesis and Implications B-21
B.4 Case Study: Prairie Streams B-22
B.4.1 Abstract B-22
B.4.2 Introduction B-22
B.4.2.1 Geography and Climate B-22
B.4.2.2 Hydrology and Geomorphology B-24
B.4.2.3 Physicochemistry B-25
B.4.2.4 Ecology B-26
B.4.2.5 Human Alterations B-27
B.4.3 Evidence B-28
B.4.3.1 Physical Connections B-28
B.4.3.2 Chemical Connections B-31
B.4.3.3 Biological Connections B-34
B.4.4 Prairie Streams: Synthesis and Implications B-36
B.5 Case Study: Southwestern Intermittent and Ephemeral Streams B-37
B.5.1 Abstract B-37
B.5.2 Introduction B-38
B.5.3 Southwestern Rivers B-39
B.5.4 San Pedro River B-42
B.5.4.1 Basin Characteristics B-42
B.5.4.2 Ephemeral Stream Connections and Their Influence on the
San Pedro River B-45
B.5.5 Other Southwestern Rivers B-48
B.5.5.1 Physical Connections B-48
B.5.5.2 Human Alterations B-50
B.5.5.3 Biological Connections B-55
B.5.6 Southwestern Intermittent and Ephemeral Streams: Synthesis and
Implications B-59
B.6 Case Study: Vernal Pools B-60
B.6.1 Abstract B-60
B.6.2 Introduction B-60
B.6.2.1 Geography and Geology B-60
B.6.2.2 Temporal Dynamics B-61
B.6.2.3 Ecology B-62
B.6.3 Evidence B-62
B.6.3.1 Physical Connections B-62
B.6.3.2 Biological Connections B-64
B.6.4 Vernal Pools: Synthesis and Implications B-66
References B-67
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LIST OF TABLES
Table 1-1. Translating connectivity-related questions between policy and science 1-2
Table 1-2. Dimensions of watershed connectivity 1-4
Table 2-1. Functions by which streams and wetlands affect material and energy fluxes to downstream
waters 2-25
Table 3-1. Examples of mechanisms by which streams are connected to and influence downstream
waters, by functional type 3-47
Table 4-1. Examples of mechanisms by which riparian/floodplain wetlands and wetlands in non-
floodplain settings influence downstream waters, by functional type 4-4
Table 4-2. Partial list of amphibian and reptile species known to use both streams and non-floodplain
wetlands or other lentic waters 4-36
Table 4-3. Key conclusions on the effects of riparian/floodplain wetlands on rivers 4-40
Table 4-4. Key conclusions on the effects of non-floodplain wetlands on rivers 4-42
Table 6-1. Relative abundance of literature by functional category 6-15
Table 6-2. Relative abundance of literature by review topic area 6-16
Table B-l. California vernal pool inundation and hydrologic connectivity B-64
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LIST OF FIGURES
Figure 1-1A. Hydrologicflowpaths 1-5
Figure 1-1B. Biological flowpaths 1-6
Figure 1-2. Temporal dynamics of hydrologic flowpaths 1-7
Figure 1-3. Effects of human alterations on watershed connectivity 1-12
Figure 1-4. The role of connectivity in maintaining the physical, chemical, and biological integrity of
water 1-13
Figure 1-5. Waters and connections considered to be within scope for this report 1-15
Figure 1-6. Flow chart for screening and compiling literature 1-17
Figure 2-1. A generalized example of a river network within its watershed 2-3
Figure 2-2. Elements of a river system 2-5
Figure 2-3. Hypothetical cross-sections of (A) a headwater stream and (B) a large river within a river
network 2-6
Figure 2-4. Water below the land surface occurs in either the unsaturated or the saturated zone 2-10
Figure 2-5. Cross-section showing major hydrologic flowpaths in a regional-scale stream-watershed
system 2-11
Figure 2-6. Hyporheic zone flows 2-13
Figure 2-7. Hypothetical hydrographs illustrating maximum duration of flow (Dmax.q) for (A) perennial,
(B) intermittent, and (C) ephemeral streams 2-15
Figure 2-8. (A) Hypothetical hydrograph showing stormflow and baseflow responses to a rainfall event.
(B) Expansion and contraction of flowing water in a stream network following a rainfall
event 2-16
Figure 2-9. Characteristics of U.S. streams by watershed, in terms of percent of total stream length as
(A) perennial, (B) intermittent, and (C) headwater streams 2-17
Figure 2-10. Extent and connectivity of streams with flowing water, wetlands, and other water bodies in
(A) Spring Valley Creek, OR and (B) Spoon Creek, OR during dry summer (left) and wet
winter (right) conditions 2-19
Figure 2-11. Stormflow moves downstream through the river network and interacts with lower stream
reaches, floodplains, and alluvial aquifers 2-20
Figure 2-12. Landsat 5 satellite images of the Mississippi River along the borders of Tennessee,
Kentucky, Missouri, and Arkansas on (A) May 12, 2006 and (B) May 10, 2011 2-21
Figure 2-13. The direction and magnitude of interactions between surface water and ground water can
dramatically change during large hydrologic events, including floods 2-23
Figure 2-14. Illustration of the sequential transformation of materials as they move through the river
network, via either downstream transport with water flow (solid black arrows) or via aerial
or terrestrial movements (dashed black arrows) 2-28
Figure 2-15. Map of annual runoff in contiguous United States showing locations of five example
streams that illustrate daily runoff patterns and total annual runoff depths 2-32
Figure 2-16. Generalized hydrologic landscape forms 2-33
Figure 2-17. Major hydrologic flowpaths for hillslopes with combinations of permeable and
impermeable soils and geologic formations 2-35
Figure 2-18. Types of hydrologic connections between non-floodplain wetlands and streams or rivers 2-37
Figure 2-19. Major types of basin shapes and network configurations 2-39
Figure 2-20. Examples of different landscapes showing interspersion of wetlands and streams or
rivers 2-41
Figure 2-21. Comparison of percent wetland loss between (A) the 1780s and mid-1980s with (B) the
distribution of artificially drained agricultural land in 1985 2-46
Figure 3-1. Longitudinal pattern of flow along (A) River Derwent and (B) River Trent, illustrating stepped
increases in flow associated with tributary inflows 3-8
Figure 3-2. Time series of rainfall and streamflow observations in the Rio Puerco and Rio Grande,
6-18 September 2003 3-9
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Figure 3-3. Airborne thermal infrared remotely sensed water temperature in the mainstem and at
tributary confluences of the North Fork John Day River, OR, on 4 August 1998 3-22
Figure 3-4. (A) A dendritic network with multilevel hierarchical structure, and (B) a uninodal network
with all headwater streams feeding directly into a river mainstem 3-44
Figure 5-1. Relative positioning of streams, riparian and floodplain waters, and non-floodplain waters
along a gradient of connectivity 5-12
Figure B-l. Aerial photograph of Carolina bays within a region of the upper Coastal Plain of South
Carolina B-2
Figure B-2. Map of the United States showing physiographic subregions and major rivers of the Great
Plains B-23
Figure B-3. Map showing the location of Kings Creek and N01B, intermittent tributaries to the Kansas
River B-29
Figure B-4. Hydrographs (instantaneous and daily mean) showing propagation of the 13 May 1995
(Julian date 133) flood downstream from headwater sites (N01B and Kings Creek) to the
Kansas River at Wamego B-30
Figure B-5. Upper: Geographic distribution of intermittent and ephemeral (red) and perennial (blue)
streams in the Continental United States and two example watersheds in Arizona and
Michigan/Ohio/Indiana from the National Hydrography Dataset (NHD) stream map B-38
Figure B-6. 2003 calendar year hydrographs from (a) the White River near Fort Apache, AZ and (b) the
San Pedro River near Tombstone, AZ B-40
Figure B-7. San Pedro River basin map showing major physiographic features and current and
historical perennial reaches B-43
Figure B-8. Perennial (blue) and nonperennial (red) streams in the San Pedro Basin from the U.S.-
Mexico border to its confluence with the Gila River based on USGS National Hydrography
Dataset (NHD) stream map (http://nhd.usgs.gov/) B-44
Figure B-9. Generalized east-west section and stratigraphic units in the middle San Pedro watershed B-45
Figure B-10. Storm rainfall and downstream hydrographs with decreasing runoff volume and peak rate
due to channel transmission losses as measured by in the USDA-ARS Walnut Gulch
Experimental Watershed (WGEW) and the impact of this storm runoff on the San Pedro
River in SE Arizona B-47
Figure B-ll. Aerial photograph showing ephemeral tributaries to Cienega Creek, a perennial stream,
flowing through the small community of Vail, southeast of Tucson, AZ B-51
Figure B-12. Change in riparian vegetation along the Santa Cruz River, Tucson, AZ, as the result of
water-level declines in the regional aquifer B-55
Figure B-13. Aerial photograph showing dense corridor of vegetation lining ephemeral washes in
southeastern Arizona B-56
LIST OF BOXES
Box 3-1. Urban Streams 3-3
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LIST OF ABBREVIATIONS AND ACRONYMS
CWA Clean Water Act
DEM digital elevation model
DOC dissolved organic carbon
FROM fine particulate organic matter
GW groundwaterflowpath
HUC Hydrologic Unit Code
NHD National Hydrography Dataset
PPR prairie pothole region
USDA-ARS United States Department of Agriculture, Agricultural Research Services
U.S. ACE United States Army Corps of Engineers
U.S. EPA United States Environmental Protection Agency
USGS United States Geological Survey
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LIST OF UNITS AND VARIABLES
A drainage area
C Celsius
c scaling power constant
cm centimeter
d day
g gram
ha hectare
hr hour
kg kilogram
km kilometer
L liter
m meter
mg milligram
Mg megagram
mm millimeter
|jM micromolar
N metric normal temperature and pressure
ng nanogram
Pg petagram
Q discharge
s second
t metric ton
TO shear stress
V velocity
yr yea r
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This report was prepared by the National Center for Environmental Assessment, the National Health and
Environmental Effects Research Laboratory, and the National Exposure Research Laboratory, in the U.S.
Environmental Protection Agency's (U.S. EPA's) Office of Research and Development. It reviews and
evaluates evidence from peer-reviewed sources that were published or in press by December 2014.
Throughout this document, terms are used with their generally recognized scientific meaning. We have
provided definitions of technical terms in the Glossary (Appendix A). Two previous drafts prepared on 1
February 2011 and 12 July 2011 were reviewed by U.S. EPA and U.S. Army Corps of Engineers staff.
Additional comments were received from scientists in government, academic, nonprofit, and private
industry organizations listed in the Reviewers section who reviewed all or part of the 1 February 2011
preliminary draft. A draft prepared on 11 October 2011 was independently peer reviewed by a panel of
11 topic experts, listed in the Reviewers section, on 30 January 2012. An external review draft released
in September 2013 (600/R-11/098B) was reviewed by U.S. EPA staff and a panel of the U.S. EPA's
Science Advisory Board (SAB) that convened 16-18 December 2013 (SAB report number EPA-SAB-15-
001, available online at www.epa.gov/sabl. The 27 topic experts comprising the SAB panel are listed in
the Reviewers section. In addition, comments from the public were received through the docket or at
the SAB panel meeting. Comments from these sources were considered and used to improve the clarity
and scientific rigor of the document.
Stream and Wetland Connectivity:
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January 2015
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AUTHORS AND REVIEWERS
AUTHORS
Laurie C. Alexander, PhD
U.S. Environmental Protection Agency
Office of Research and Development, National Center for Environmental Assessment
Washington, DC
Bradley Autrey, MS, JD
U.S. Environmental Protection Agency
Office of Research and Development, National Exposure Research Laboratory
Cincinnati, OH
Julie DeMeester, PhD
American Association for the Advancement of Science Fellow
U.S. Environmental Protection Agency
Office of Research and Development, National Center for Environmental Assessment
Washington, DC
Ken M. Fritz, PhD
U.S. Environmental Protection Agency
Office of Research and Development, National Exposure Research Laboratory
Cincinnati, OH
Heather E. Golden, PhD
U.S. Environmental Protection Agency
Office of Research and Development, National Exposure Research Laboratory
Cincinnati, OH
David C. Goodrich, PhD
U.S. Department of Agriculture
Agricultural Research Service
Tucson, AZ
William G. Kepner, MS, MPA
U.S. Environmental Protection Agency
Office of Research and Development, National Exposure Research Laboratory,
Environmental Sciences Division
Las Vegas, NV
Stream and Wetland Connectivity:
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January 2015
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AUTHORS (continued)
Charles R. Lane, PhD
U.S. Environmental Protection Agency
Office of Research and Development, National Exposure Research Laboratory
Cincinnati, OH
Stephen D. LeDuc, PhD
U.S. Environmental Protection Agency
Office of Research and Development, National Center for Environmental Assessment
Washington, DC
Scott G. Leibowitz, PhD
U.S. Environmental Protection Agency
Office of Research and Development, National Health and Environmental Effects Research
Laboratory
Corvallis, OR
Michael G. McManus, PhD
U.S. Environmental Protection Agency
Office of Research and Development, National Center for Environmental Assessment
Cincinnati, OH
Amina I. Pollard, PhD
U.S. Environmental Protection Agency
Office of Water, Office of Wetlands, Oceans, and Watersheds
Washington, DC
Hadas Raanan Kiperwas, PhD
Oak Ridge Institute for Science and Education Fellow, U.S. Environmental Protection Agency
Office of Water, Office of Wetlands, Oceans, and Watersheds
Washington, DC
Caroline E. Ridley, PhD
U.S. Environmental Protection Agency
Office of Research and Development, National Center for Environmental Assessment
Washington, DC
Kate Schofield, PhD
U.S. Environmental Protection Agency
Office of Research and Development, National Center for Environmental Assessment
Washington, DC
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
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AUTHORS (continued)
Melanie Vanderhoof, PhD
Oak Ridge Institute for Science and Education Fellow, U.S. Environmental Protection Agency
Office of Research and Development, National Center for Environmental Assessment
Washington, DC
Parker J. Wigington, PhD
U.S. Environmental Protection Agency
Office of Research and Development, National Health and Environmental Effects Research
Laboratory
Corvallis, OR
PEER CONSULTATION REVIEWERS
Robert T. Brooks, PhD
U.S. Department of Agriculture Forest Service, Amherst, MA
William H. Eldridge, PhD
Stroud Water Research Center, Avondale, PA
Keith B. Gido, PhD
Kansas State University, Manhattan, KS
Arthur J. Gold, PhD
University of Rhode Island, Kingston, RI
Tracie-Lynn Nadeau, PhD
U.S. Environmental Protection Agency, Portland, OR
Denis Newbold, PhD
Stroud Water Research Center, Avondale, PA
Michael J. Paul, PhD
TetraTech Inc., Owings Mills, MD
Donald Rosenberry, PhD
U.S. Geological Survey, Lakewood, CO
Doug Samson, PhD
The Nature Conservancy, Bethesda, MD
Rebecca Sharitz, PhD
Savannah River Ecology Laboratory, Aiken, GA
Kirk 0. Winemiller, PhD
Texas A&M University, College Station, TX
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EXTERNAL PEER REVIEW PANEL
David J. Cooper, PhD
Colorado State University, Fort Collins, CO
William G. Crumpton, PhD
Iowa State University, Ames, IA
Kenneth W. Cummins, PhD
Humboldt State University, Arcata, CA
Walter K. Dodds, PhD (Chair)
Kansas State University, Manhattan, KS
James W. La Baugh, PhD
U.S. Geological Survey, Reston, VA
Mark C. Rains, PhD
University of South Florida, Tampa, FL
John S. Richardson, PhD
University of British Columbia, Vancouver, BC
Joel W. Snodgrass, PhD
Towson University, Towson, MD
Arnold van der Valk, PhD
Iowa State University, Ames, IA
Mark S. Wipfli, PhD
U.S. Geological Survey, Fairbanks, AK
William R. Wise, PhD
University of Florida, Gainesville, FL
U.S. ENVIRONMENTAL PROTECTION AGENCY SCIENCE ADVISORY BOARD
PANEL FOR THE REVIEW OF THE EPA WATER BODY CONNECTIVITY REPORT
Amanda D. Rodewald, PhD (Chair)
Cornell University, Ithaca, NY
Allison Aldous, PhD
The Nature Conservancy, Portland, OR
Genevieve Ali, PhD
University of Manitoba, Winnipeg, MB, Canada
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U.S. ENVIRONMENTAL PROTECTION AGENCY SCIENCE ADVISORY BOARD
PANEL FOR THE REVIEW OF THE EPA WATER BODY CONNECTIVITY REPORT (continued)
J. David Allan, PhD
University of Michigan, Ann Arbor, MI
Lee Benda, PhD
Earth Systems Institute, Mt. Shasta, CA
Emily S. Bernhardt, PhD
Duke University, Durham, NC
Robert P. Brooks, PhD
Pennsylvania State University, University Park, PA
Kurt Fausch, PhD
Colorado State University, Fort Collins, CO
Siobhan Fennessy, PhD
Kenyon College, Gambler, OH
Michael Gooseff, PhD
Colorado State University, Fort Collins, CO
Judson Harvey, PhD
U.S. Geological Survey, Reston, VA
Charles Hawkins, PhD
Utah State University, Logan, UT
Lucinda B. Johnson, PhD
University of Minnesota Duluth, Duluth, MN
Michael Josselyn, PhD
Wetlands Research Associates, Inc., San Rafael, CA
LatifKalin, PhD
Auburn University, Auburn, AL
Kenneth Kolm, PhD
Hydrologic Systems Analysis, LLC, Golden, CO
Judith L. Meyer, PhD
University of Georgia, Lopez Island, WA
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U.S. ENVIRONMENTAL PROTECTION AGENCY SCIENCE ADVISORY BOARD
PANEL FOR THE REVIEW OF THE EPA WATER BODY CONNECTIVITY REPORT (continued)
Mark Murphy, PhD
Hassayampa Associates, Tucson, AZ
Duncan Patten, PhD
Arizona State University, Bozeman, MT
Mark Rains, PhD
University of South Florida, Tampa, FL
Ramesh Reddy, PhD
University of Florida, Gainesville, FL
Emma Rosi-Marshall, PhD
Gary Institute of Ecosystem Studies, Millbrook, NY
Jack Stanford, PhD
University of Montana, Poison, MT
Mazeika Sullivan, PhD
The Ohio State University, Columbus, OH
Jennifer Tank, PhD
University of Notre Dame, Notre Dame, IN
Maurice Valett, PhD
University of Montana, Missoula, MT
Ellen Wohl, PhD
Colorado State University, Fort Collins, CO
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PHOTO CREDITS
Front cover, Executive Summary, Chapter 5, and References
Back cover, Executive Summary, and References
Chapter 3
All other photos (U.S. EPA)
Nambe Lake, New Mexico (L.C.
Alexander, U.S. EPA)
Children in Delaware inland wetland
(Hennis H. Bartow, Delaware Center for
the Inland Bays)
Mayfly (Heptagenia culacanthd) (David
H. Funk, Stroud Water Research Center)
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ACKNOWLEDGMENTS
We gratefully acknowledge support provided by ICF International in Durham, NC and Tetra Tech, Inc. in
Owings Mills, MD for preparation and production of draft and final reports; Eastern Research Group, Inc.
in Lexington, MA for organizing and managing an independent peer review of a draft report; and Ms. Iris
Goodman, Dr. Thomas Armitage, and Dr. Angela Nugent at the U.S. Environmental Protection Agency,
Science Advisory Board (SAB) in Washington, DC for their service as Designated Federal Officers for the
SAB review of this report.
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BACKGROUND
The objective of the Clean Water Act is to restore and maintain the chemical, physical, and biological
integrity of the nation's waters. The U.S. Environmental Protection Agency's (U.S. EPA's) Office of
Research and Development developed this report to inform rulemaking by the U.S. EPA and U.S. Army
Corps of Engineers (U.S. ACE) on the definition of "waters of the United States" under the Clean Water
Act (CWA). Its purpose is to summarize current scientific understanding about the connectivity and
mechanisms by which streams and wetlands, singly or in aggregate, affect the physical, chemical, and
biological integrity of downstream waters. The focus of the review is on surface and shallow subsurface
connections of small or temporary streams, nontidal wetlands, and certain open waters. Because this
report is a technical review of peer-reviewed scientific literature, it neither considers nor sets forth legal
standards for CWA jurisdiction, nor does it establish EPA policy.
The report is organized into six chapters. Chapter 1 outlines the purpose, scientific context, and
approach of the report. Chapter 2 describes the components of a river system and watershed; the types
of physical, chemical, and biological connections that link those components; the factors that influence
connectivity at various temporal and spatial scales; and methods for quantifying connectivity. Chapter 3
reviews literature on connectivity in stream networks in terms of physical, chemical, and biological
connections and their resulting effects on downstream waters. Chapter 4 reviews literature on the
connectivity and effects of nontidal wetlands and certain open waters on downstream waters. Chapter 5
applies concepts and evidence from previous chapters to six case studies from published literature on
Carolina and Delmarva bays, oxbow lakes, prairie potholes, prairie streams, southwestern streams, and
vernal pools. Chapter 6 summarizes key findings and conclusions, identifies data gaps, and briefly
discusses research approaches that could fill those gaps. A glossary of scientific terms used in the report
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and detailed case studies of selected systems (summarized in Chapter 5) are included in Appendix A and
Appendix B, respectively.
SUMMARY OF MAJOR CONCLUSIONS
Based on the review and synthesis of more than 1,200 publications from the peer reviewed scientific
literature, the evidence supports five major conclusions. Citations have been omitted from the text to
improve readability; please refer to individual chapters for supporting publications and additional
information.
Conclusion 1: Streams
The scientific literature unequivocally demonstrates that streams, individually or cumulatively,
exert a strong influence on the integrity of downstream waters. All tributary streams, including
perennial, intermittent, and ephemeral streams, are physically, chemically, and biologically
connected to downstream rivers via channels and associated alluvial deposits where water and
other materials are concentrated, mixed, transformed, and transported. Streams are the
dominant source of water in most rivers, and the majority of tributaries are perennial,
intermittent, or ephemeral headwater streams. Headwater streams also convey water into local
storage compartments such as ponds, shallow aquifers, or stream banks, and into regional and
alluvial aquifers; these local storage compartments are important sources of water for
maintaining baseflow in rivers. In addition to water, streams transport sediment, wood, organic
matter, nutrients, chemical contaminants, and many of the organisms found in rivers. The
literature provides robust evidence that streams are biologically connected to downstream
waters by the dispersal and migration of aquatic and semiaquatic organisms, including fish,
amphibians, plants, microorganisms, and invertebrates, that use both upstream and
downstream habitats during one or more stages of their life cycles, or provide food resources to
downstream communities. In addition to material transport and biological connectivity,
ephemeral, intermittent, and perennial flows influence fundamental biogeochemical processes
by connecting channels and shallow ground water with other landscape elements. Physical,
chemical, and biological connections between streams and downstream waters interact via
integrative processes such as nutrient spiraling, in which stream communities assimilate and
chemically transform large quantities of nitrogen and other nutrients that otherwise would be
transported directly downstream, increasing nutrient loads and associated impairments due to
excess nutrients in downstream waters.
Conclusion 2: Riparian/Floodplain Wetlands and Open Waters
The literature clearly shows that wetlands and open waters in riparian areas and floodplains are
physically, chemically, and biologically integrated with rivers via functions that improve
downstream water quality, including the temporary storage and deposition of channel-forming
sediment and woody debris, temporary storage of local ground water that supports baseflow in
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rivers, and transformation and transport of stored organic matter. Riparian/floodplain wetlands
and open waters improve water quality through the assimilation, transformation, or
sequestration of pollutants, including excess nutrients and chemical contaminants such as
pesticides and metals, that can degrade downstream water integrity. In addition to providing
effective buffers to protect downstream waters from point source and nonpoint source
pollution, these systems form integral components of river food webs, providing nursery habitat
for breeding fish and amphibians, colonization opportunities for stream invertebrates, and
maturation habitat for stream insects. Lateral expansion and contraction of the river in its
floodplain result in an exchange of organic matter and organisms, including fish populations that
are adapted to use floodplain habitats for feeding and spawning during high water, that are
critical to river ecosystem function. Riparian/floodplain wetlands and open waters also affect
the integrity of downstream waters by subsequently releasing (desynchronizing) floodwaters
and retaining large volumes of stormwater, sediment, and contaminants in runoff that could
otherwise negatively affect the condition or function of downstream waters.
Conclusion 3: Non-floodplain Wetlands and Open Waters
Wetlands and open waters in non-floodplain landscape settings (hereafter called "non-
floodplain wetlands") provide numerous functions that benefit downstream water integrity.
These functions include storage of floodwater; recharge of ground water that sustains river
baseflow; retention and transformation of nutrients, metals, and pesticides; export of organisms
or reproductive propagules to downstream waters; and habitats needed for stream species. This
diverse group of wetlands (e.g., many prairie potholes, vernal pools, playa lakes) can be
connected to downstream waters through surface-water, shallow subsurface-water, and
ground-water flows and through biological and chemical connections.
In general, connectivity of non-floodplain wetlands occurs along a gradient (Conclusion 4), and
can be described in terms of the frequency, duration, magnitude, timing, and rate of change of
water, material, and biotic fluxes to downstream waters. These descriptors are influenced by
climate, geology, and terrain, which interact with factors such as the magnitudes of the various
functions within wetlands (e.g., amount of water storage or carbon export) and their proximity
to downstream waters to determine where wetlands occur along the connectivity gradient. At
one end of this gradient, the functions of non-floodplain wetlands clearly affect the condition of
downstream waters if a visible (e.g., channelized) surface-water or a regular shallow subsurface-
water connection to the river network is present. For non-floodplain wetlands lacking a
channelized surface or regular shallow subsurface connection (i.e., those at intermediate points
along the gradient of connectivity), generalizations about their specific effects on downstream
waters from the available literature are difficult because information on both function and
connectivity is needed. Although there is ample evidence that non-floodplain wetlands provide
hydrologic, chemical, and biological functions that affect material fluxes, to date, few scientific
studies explicitly addressing connections between non-floodplain wetlands and river networks
have been published in the peer-reviewed literature. Even fewer publications specifically focus
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on the frequency, duration, magnitude, timing, or rate of change of these connections. In
addition, although areas that are closer to rivers and streams have a higher probability of being
connected than areas farther away when conditions governing the type and quantity of flows—
including soil infiltration rate, wetland storage capacity, hydraulic gradient, etc.—are similar,
information to determine if this similarity holds is generally not provided in the studies we
reviewed. Thus, current science does not support evaluations of the degree of connectivity for
specific groups or classes of wetlands (e.g., prairie potholes or vernal pools). Evaluations of
individual wetlands or groups of wetlands, however, could be possible through case-by-case
analysis.
Some effects of non-floodplain wetlands on downstream waters are due to their isolation, rather
than their connectivity. Wetland "sink" functions that trap materials and prevent their export to
downstream waters (e.g., sediment and entrained pollutant removal, water storage) result
because of the wetland's ability to isolate material fluxes. To establish that such functions
influence downstream waters, we also need to know that the wetland intercepts materials that
otherwise would reach the downstream water. The literature we reviewed does provide limited
examples of direct effects of wetland isolation on downstream waters, but not for classes of
wetlands (e.g., vernal pools). Nevertheless, the literature we reviewed enables us to conclude
that sink functions of non-floodplain wetlands, which result in part from their relative isolation,
will affect a downstream water when these wetlands are situated between the downstream
water and known point or nonpoint sources of pollution, and thus intersect flowpaths between
the pollutant source and downstream waters.
Conclusion 4: Degrees and Determinants of Connectivity
Watersheds are integrated at multiple spatial and temporal scales by flows of surface water and
ground water, transport and transformation of physical and chemical materials, and movements
of organisms. Although all parts of a watershed are connected to some degree—by the
hydrologic cycle or dispersal of organisms, for example—the degree and downstream effects of
those connections vary spatially and temporally, and are determined by characteristics of the
physical, chemical, and biological environments and by human activities.
Stream and wetland connections have particularly important consequences for downstream
water integrity. Most of the materials—broadly defined as any physical, chemical, or biological
entity—in rivers, for example, originate from aquatic ecosystems located upstream or elsewhere
in the watershed. Longitudinal flows through ephemeral, intermittent, and perennial stream
channels are much more efficient for transport of water, materials, and organisms than diffuse
overland flows, and areas that concentrate water provide mechanisms for the storage and
transformation, as well as transport, of materials.
Connectivity of streams and wetlands to downstream waters occurs along a continuum that can
be described in terms of the frequency, duration, magnitude, timing, and rate of change of water,
material, and biotic fluxes to downstream waters. These terms, which we refer to collectively as
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connectivity descriptors, characterize the range over which streams and wetlands vary and shift
along the connectivity gradient in response to changes in natural and anthropogenic factors and,
when considered in a watershed context, can be used to predict probable effects of different
degrees of connectivity over time. The evidence unequivocally demonstrates that the stream
channels and riparian/floodplain wetlands or open waters that together form river networks
are clearly connected to downstream waters in ways that profoundly influence downstream
water integrity. The connectivity and effects of non-floodplain wetlands and open waters are
more variable and thus more difficult to address solely from evidence available in peer-
reviewed studies.
Variations in the degree of connectivity influence the range of functions provided by streams
and wetlands, and are critical to the integrity and sustainability of downstream waters.
Connections with low values of one or more descriptors (e.g., low-frequency, low-duration
streamflows caused by flash floods) can have important downstream effects when considered in
the context of other descriptors (e.g., large magnitude of water transfer). At the other end of the
frequency range, high-frequency, low-magnitude vertical (surface-subsurface) and lateral flows
contribute to aquatic biogeochemical processes, including nutrient and contaminant
transformation and organic matter accumulation. The timing of an event can alter both
connectivity and the magnitude of its downstream effect. For example, when soils become
saturated by previous rainfall events, even low or moderate rainfall can cause streams or
wetlands to overflow, transporting water and materials to downstream waters. Fish that use
nonperennial or perennial headwater stream habitats to spawn or rear young, and invertebrates
that move into seasonally inundated floodplain wetlands prior to emergence, have life cycles
that are synchronized with the timing of flows, temperature thresholds, and food resource
availability in those habitats.
Conclusion 5: Cumulative Effects
The incremental effects of individual streams and wetlands are cumulative across entire
watersheds and therefore must be evaluated in context with other streams and wetlands.
Downstream waters are the time-integrated result of all waters contributing to them. For
example, the amount of water or biomass contributed by a specific ephemeral stream in a given
year might be small, but the aggregate contribution of that stream over multiple years, or by all
ephemeral streams draining that watershed in a given year or over multiple years, can have
substantial consequences on the integrity of the downstream waters. Similarly, the downstream
effect of a single event, such as pollutant discharge into a single stream or wetland, might be
negligible but the cumulative effect of multiple discharges could degrade the integrity of
downstream waters.
In addition, when considering the effect of an individual stream or wetland, all contributions and
functions of that stream or wetland should be evaluated cumulatively. For example, the same
stream transports water, removes excess nutrients, mitigates flooding, and provides refuge for
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fish when conditions downstream are unfavorable; if any of these functions is ignored, the
overall effect of that stream would be underestimated.
SUPPORT FOR MAJOR CONCLUSIONS
This report synthesizes a large body of scientific literature on the connectivity and mechanisms by
which streams, wetlands, and open waters, singly or in aggregate, affect the physical, chemical, and
biological integrity of downstream waters. The major conclusions reflect the strength of evidence
currently available in the peer-reviewed scientific literature for assessing the connectivity and
downstream effects of water bodies identified in Chapter 1 of this report.
The conclusions of this report were corroborated by two independent peer reviews by scientists
identified in the front matter of this report.
The term connectivity is defined in this report as the degree to which components of a watershed are
joined and interact by transport mechanisms that function across multiple spatial and temporal scales.
Connectivity is determined by the characteristics of both the physical landscape and the biota of the
specific system. Our review found strong evidence supporting the central roles of the physical, chemical,
and biological connectivity of streams, wetlands, and open waters—encompassing varying degrees of
both connection and isolation—in maintaining the structure and function of downstream waters,
including rivers, lakes, estuaries, and oceans. Our review also found strong evidence demonstrating the
various mechanisms by which material and biological linkages from streams, wetlands, and open waters
affect downstream waters, classified here into five functional categories (source, sink, refuge, lag, and
transformation; discussed below), and modify the timing of transport and the quantity and quality of
resources available to downstream ecosystems and communities. Thus, the currently available literature
provided a large body of evidence for assessing the types of connections and functions by which streams
and wetlands produce the range of observed effects on the integrity of downstream waters.
We identified five categories of functions by which streams, wetlands, and open waters influence the
timing, quantity, and quality of resources available to downstream waters:
• Source: the net export of materials, such as water and food resources;
• Sink: the net removal or storage of materials, such as sediment and contaminants;
• Refuge: the protection of materials, especially organisms;
• Transformation: the transformation of materials, especially nutrients and chemical
contaminants, into different physical or chemical forms; and
• Lag: the delayed or regulated release of materials, such as storm water.
These functions are not mutually exclusive; for example, the same stream or wetland can be both a
source of organic matter and a sink for nitrogen. The presence or absence of these functions, which
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depend on the biota, hydrology, and environmental conditions in a watershed, can change over time; for
example, the same wetland can attenuate runoff during storm events and provide ground-water
recharge following storms. Further, some functions work in conjunction with others; a lag function can
include transformation of materials prior to their delayed release. Finally, effects on downstream waters
should consider both actual function and potential function. A potential function represents the capacity
of an ecosystem to perform that function under suitable conditions. For example, a wetland with high
capacity for denitrification is a potential sink for nitrogen, a nutrient that becomes a contaminant when
present in excessive concentrations. In the absence of nitrogen, this capacity represents the wetland's
potential function. If nitrogen enters the wetland (e.g., from fertilizer in runoff), it is removed from the
water; this removal represents the wetland's actual function. Both potential and actual functions play
critical roles in protecting and restoring downstream waters as environmental conditions change.
The evidence unequivocally demonstrates that the stream channels and riparian/floodplain wetlands or
open waters that together form river networks are clearly connected to downstream waters in ways
that profoundly influence downstream water integrity. The body of literature documenting connectivity
and downstream effects was most abundant for perennial and intermittent streams, and for
riparian/floodplain wetlands. Although less abundant, the evidence for connectivity and downstream
effects of ephemeral streams was strong and compelling, particularly in context with the large body of
evidence supporting the physical connectivity and cumulative effects of channelized flows that form and
maintain stream networks.
As stated in Conclusion 3, the connectivity and effects of wetlands and open waters that lack visible
surface connections to other water bodies are more difficult to address solely from evidence available in
the peer-reviewed literature. The limited evidence currently available shows that these systems have
important hydrologic, water-quality, and habitat functions that can affect downstream waters where
connections to them exist; the literature also provides limited examples of direct effects of non-
floodplain wetland isolation on downstream water integrity. Currently available peer-reviewed
literature, however, does not identify which types or classes of non-floodplain wetlands have or lack the
types of connections needed to convey the effects on downstream waters of functions, materials, or
biota provided by those wetlands.
KEY FINDINGS FOR MAJOR CONCLUSIONS
This section summarizes key findings for each of the five major conclusions, above and in Chapter 6 of
the report. Citations have been omitted from the text to improve readability; please refer to individual
chapters for supporting publications and additional information.
Conclusion 1, Streams: Key Findings
• Streams are hydrologically connected to downstream waters via channels that convey surface
and subsurface water either year-round (i.e., perennial flow), weekly to seasonally (i.e.,
intermittent flow), or only in direct response to precipitation (i.e., ephemeral flow). Streams are
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the dominant source of water in most rivers, and the majority of tributaries are perennial,
intermittent, or ephemeral headwater streams. For example, headwater streams, which are the
smallest channels where streamflows begin, are the cumulative source of approximately 60% of
the total mean annual flow to all northeastern U.S. streams and rivers.
• In addition to downstream transport, headwaters convey water into local storage compartments
such as ponds, shallow aquifers, or stream banks, and into regional and alluvial aquifers. These
local storage compartments are important sources of water for maintaining baseflow in rivers.
Streamflow typically depends on the delayed (i.e., lagged) release of shallow ground water from
local storage, especially during dry periods and in areas with shallow ground-water tables and
pervious subsurfaces. For example, in the southwestern United States, short-term shallow
ground-water storage in alluvial floodplain aquifers, with gradual release into stream channels,
is a major source of annual flow in rivers.
• Infrequent, high-magnitude events are especially important for transmitting materials from
headwater streams in most river networks. For example, headwater streams, including
ephemeral and intermittent streams, shape river channels by accumulating and gradually or
episodically releasing stored materials such as sediment and large woody debris. These
materials help structure stream and river channels by slowing the flow of water through
channels and providing substrate and habitat for aquatic organisms.
• There is strong evidence that headwater streams function as nitrogen sources (via export) and
sinks (via uptake and transformation) for river networks. For example, one study estimated that
rapid nutrient cycling in small streams with no agricultural or urban impacts removed 20-40%
of the nitrogen that otherwise would be delivered to downstream waters. Nutrients are
necessary to support aquatic life, but excess nutrients lead to eutrophication and hypoxia, in
which over-enrichment causes dissolved oxygen concentrations to fall below the level necessary
to sustain most aquatic animal life in the stream and streambed. Thus, the influence of streams
on nutrient loads can have significant repercussions for hypoxia in downstream waters.
• Headwaters provide habitat that is critical for completion of one or more life-cycle stages of
many aquatic and semiaquatic species capable of moving throughout river networks. Evidence
is strong that headwaters provide habitat for complex life-cycle completion; refuge from
predators, competitors, parasites, or adverse physical conditions in rivers (e.g., temperature or
flow extremes, low dissolved oxygen, high sediment); and reservoirs of genetic- and species-
level diversity. Use of headwater streams as habitat is especially critical for the many species
that migrate between small streams and marine environments during their life cycles (e.g.,
Pacific and Atlantic salmon, American eels, certain lamprey species). The presence of these
species within river networks provides robust evidence of biological connections between
headwaters and larger rivers; because these organisms also transport nutrients and other
materials as they migrate, their presence also provides evidence of biologically mediated
chemical connections. In prairie streams, many fishes swim upstream into tributaries to release
eggs, which develop as they are transported downstream.
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• Human alterations affect the frequency, duration, magnitude, timing, and rate of change of
connections between headwater streams, including ephemeral and intermittent streams, and
downstream waters. Human activities and built structures (e.g., channelization, dams, ground-
water withdrawals) can either enhance or fragment longitudinal connections between
headwater streams and downstream waters, while also constraining lateral and vertical
exchanges and tightly controlling the temporal dimension of connectivity. In many cases,
research on human alterations has enhanced our understanding of the headwater stream-
downstream water connections and their consequences. Recognition of these connections and
effects has encouraged the development of more sustainable practices and infrastructure to
reestablish and manage connections, and ultimately to protect and restore the integrity of
downstream waters.
Conclusion 2, Riparian/Floodplain Wetlands and Open Waters: Key Findings
• Riparian areas and floodplains connect upland and aquatic environments through both surface
and subsurface hydrologic flowpaths. These areas are therefore uniquely situated in watersheds
to receive and process waters that pass over densely vegetated areas and through subsurface
zones before the waters reach streams and rivers. When pollutants reach a riparian or
floodplain wetland, they can be sequestered in sediments, assimilated into wetland plants and
animals, transformed into less harmful or mobile forms or compounds, or lost to the
atmosphere. Wetland potential for biogeochemical transformations (e.g., denitrification) that
can improve downstream water quality is influenced by local factors, including anoxic
conditions and slow organic matter decomposition, shallow water tables, wetland plant
communities, permeable soils, and complex topography.
• Riparian/floodplain wetlands can reduce flood peaks by storing and desynchronizing
floodwaters. They can also maintain river baseflows by recharging alluvial aquifers. Many
studies have documented the ability of riparian/floodplain wetlands to reduce flood pulses by
storing excess water from streams and rivers. One review of wetland studies reported that
riparian wetlands reduced or delayed floods in 23 of 28 studies. For example, peak discharges
between upstream and downstream gaging stations on the Cache River in Arkansas were
reduced 10-20% primarily due to floodplain water storage.
• Riparian areas and floodplains store large amounts of sediment and organic matter from
upstream and from upland areas. For example, riparian areas have been shown to remove
80-90% of sediments leaving agricultural fields in North Carolina.
• Ecosystem function within a river system is driven in part by biological connectivity that links
diverse biological communities with the river system. Movements of organisms that connect
aquatic habitats and their populations, even across different watersheds, are important for the
survival of individuals, populations, and species, and for the functioning of the river ecosystem.
For example, lateral expansion and contraction of the river in its floodplain result in an exchange
of matter and organisms, including fish populations that are adapted to use floodplain habitats
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for feeding and spawning during high water. Wetland and aquatic plants in floodplains can
become important seed sources for the river network, especially if catastrophic flooding scours
vegetation and seed banks in other parts of the channel. Many invertebrates exploit temporary
hydrologic connections between rivers and floodplain wetland habitats, moving into these
wetlands to feed, reproduce, or avoid harsh environmental conditions and then returning to the
river network. Amphibians and aquatic reptiles commonly use both streams and
riparian/floodplain wetlands to hunt, forage, overwinter, rest, or hide from predators. Birds can
spatially integrate the watershed landscape through biological connectivity.
Conclusion 3, Non-floodplain Wetlands and Open Waters: Key Findings
• Water storage by wetlands well outside of riparian or floodplain areas can affect streamflow.
Hydrologic models of prairie potholes in the Starkweather Coulee subbasin (North Dakota) that
drains to Devils Lake indicate that increasing the volume of pothole storage across the subbasin
by approximately 60% caused simulated total annual streamflow to decrease 50% during a
series of dry years and 20% during wet years. Similar simulation studies of watersheds that feed
the Red River of the North in North Dakota and Minnesota demonstrated qualitatively
comparable results, suggesting that the ability of potholes to modulate streamflow could be
widespread across eastern portions of the prairie pothole region. This work also indicates that
reducing water storage capacity of wetlands by connecting formerly isolated potholes through
ditching or drainage to the Devils Lake and Red River basins could increase stormflow and
contribute to downstream flooding. In many agricultural areas already crisscrossed by extensive
drainage systems, total streamflow and baseflow are increased by directly connecting potholes
to stream networks. The impacts of changing streamflow are numerous, including altered flow
regime, stream geomorphology, habitat, and ecology. The presence or absence of an effect of
prairie pothole water storage on streamflow depends on many factors, including patterns of
precipitation, topography, and degree of human alteration. For example, in parts of the prairie
pothole region with low precipitation, low stream density, and little human alteration,
hydrologic connectivity between prairie potholes and streams or rivers is likely to be low.
• Non-floodplain wetlands act as sinks and transformers for various pollutants, especially
nutrients, which at excess levels can adversely impact human and ecosystem health and pose a
serious pollution problem in the United States. In one study, sewage waste waters were applied
to forested wetlands in Florida for 4.5 years; more than 95% of the phosphorus, nitrate,
ammonium, and total nitrogen were removed by the wetlands during the study period, and
66-86% of the nitrate removed was attributed to the process of denitrification. In another
study, sizeable phosphorus retention (0.3 to 8.0 mg soluble reactive P nr2 d-1) occurred in
marshes that comprised only 7% of the lower Lake Okeechobee basin area in Florida. A non-
floodplain bog in Massachusetts was reported to sequester nearly 80% of nitrogen inputs from
various sources, including atmospheric deposition, and prairie pothole wetlands in the upper
Midwest were found to remove >80% of the nitrate load via denitrification. A large prairie
marsh was found to remove 86% of nitrate, 78% of ammonium, and 20% of phosphate through
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assimilation and sedimentation, sorption, and other mechanisms. Together, these and other
studies indicate that onsite nutrient removal by non-floodplain wetlands is substantial and
geographically widespread. The effects of this removal on rivers are generally not reported in
the literature.
Non-floodplain wetlands provide unique and important habitats for many species, both common
and rare. Some of these species require multiple types of waters to complete their full life cycles,
including downstream waters. Abundant or highly mobile species play important roles in
transferring energy and materials between non-floodplain wetlands and downstream waters.
Biological connections are likely to occur between most non-floodplain wetlands and
downstream waters through either direct or stepping stone movement of amphibians,
invertebrates, reptiles, mammals, and seeds of aquatic plants, including colonization by invasive
species. Many species in those groups that use both stream and wetland habitats are capable of
dispersal distances equal to or greater than distances between many wetlands and river
networks. Migratory birds can be an important vector of long-distance dispersal of plants and
invertebrates between non-floodplain wetlands and the river network, although their influence
has not been quantified. Whether those connections are of sufficient magnitude to impact
downstream waters will either require estimation of the magnitude of material fluxes or
evidence that these movements of organisms are required for the survival and persistence of
biota that contribute to the integrity of downstream waters.
Spatial proximity is one important determinant of the magnitude, frequency and duration of
connections between wetlands and streams that will ultimately influence the fluxes of water,
materials and biota between wetlands and downstream waters. However, proximity alone is not
sufficient to determine connectivity, due to local variation in factors such as slope and
permeability.
The cumulative influence of many individual wetlands within watersheds can strongly affect the
spatial scale, magnitude, frequency, and duration of hydrologic, biological and chemical fluxes or
transfers of water and materials to downstream waters. Because of their aggregated influence,
any evaluation of changes to individual wetlands should be considered in the context of past and
predicted changes (e.g., from climate change) to other wetlands within the same watershed.
Non-floodplain wetlands can be hydrologically connected directly to river networks through
natural or constructed channels, nonchannelized surface flows, or subsurface flows, the latter of
which can travel long distances to affect downstream waters. A wetland surrounded by uplands
is defined as "geographically isolated." Our review found that, in some cases, wetland types such
as vernal pools and coastal depressional wetlands are collectively—and incorrectly—referred to
as geographically isolated. Technically, the term "geographically isolated" should be applied only
to the particular wetlands within a type or class that are completely surrounded by uplands.
Furthermore, "geographic isolation" should not be confused with functional isolation, because
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geographically isolated wetlands can still have hydrologic, chemical, and biological connections
to downstream waters.
• Non-floodplain wetlands occur along a gradient of hydrologic connectivity-isolation with
respect to river networks, lakes, or marine/estuarine water bodies. This gradient includes, for
example, wetlands that serve as origins for stream channels that have permanent surface-water
connections to the river network; wetlands with outlets to stream channels that discharge to
deep ground-water aquifers; geographically isolated wetlands that have local ground-water or
occasional surface-water connections to downstream waters; and geographically isolated
wetlands that have minimal hydrologic connection to other water bodies (but which could
include surface and subsurface connections to other wetlands). This gradient can exist among
wetlands of the same type or in the same geographic region.
• Caution should be used in interpreting connectivity for wetlands that have been designated as
"geographically isolated" because (1) the term can be applied broadly to a heterogeneous group
of wetlands, which can include wetlands that are not actually geographically isolated; (2)
wetlands with permanent channels could be miscategorized as geographically isolated if the
designation is based on maps or imagery with inadequate spatial resolution, obscured views,
etc.; and (3) wetland complexes could have connections to downstream waters through stream
channels even if individual wetlands within the complex are geographically isolated. For
example, a recent study examined hydrologic connectivity in a complex of wetlands on the Texas
Coastal Plain. The wetlands in this complex have been considered to be a type of geographically
isolated wetland; however, collectively they are connected both geographically and
hydrologically to downstream waters in the area: During an almost 4-year study period, nearly
20% of the precipitation that fell on the wetland complex flowed out through an intermittent
stream into downstream waters. Thus, wetland complexes could have connections to
downstream waters through stream channels even when the individual wetland components
are geographically isolated.
Conclusion 4, Degrees and Determinants of Connectivity: Key Findings
• The surface-water and ground-water flowpaths (hereafter, hydrologic flowpaths), along which
water and materials are transported and transformed, determine variations in the degree of
physical and chemical connectivity. These flowpaths are controlled primarily by variations in
climate, geology, and terrain within and among watersheds and over time. Climate, geology, and
terrain are reflected locally in factors such as rainfall and snowfall intensity, soil infiltration
rates, and the direction of ground-water flows. These local factors interact with the landscape
positions of streams and wetlands relative to downstream waters, and with functions (such as
the removal or transformation of pollutants) performed by those streams and wetlands to
determine connectivity gradients.
• Gradients of biological connectivity (i.e., the active or passive movements of organisms through
water or air and over land that connect populations) are determined primarily by species
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assemblages, and by features of the landscape (e.g., climate, geology, terrain) that facilitate or
impede the movement of organisms. The temporal and spatial scales at which biological
pathways connect aquatic habitats depend on characteristics of both the landscape and species,
and overland transport or movement can occur across watershed boundaries. Dispersal is
essential for population persistence, maintenance of genetic diversity, and evolution of aquatic
species. Consequently, dispersal strategies reflect aquatic species' responses and adaptations to
biotic and abiotic environments, including spatial and temporal variation in resource availability
and quality. Species' traits and behaviors encompass species-environment relationships over
time, and provide an ecological and evolutionary context for evaluating biological connectivity
in a particular watershed or group of watersheds.
• Pathways for chemical transport and transformation largely follow hydrologic flowpaths, but
sometimes follow biological pathways (e.g., nutrient transport from wetlands to coastal waters
by migrating waterfowl, upstream transport of marine-derived nutrients by spawning of
anadromous fish, uptake and removal of nutrients by emerging stream insects).
• Human activities alter naturally occurring gradients of physical, chemical, and biological
connectivity by modifying the frequency, duration, magnitude, timing, and rate of change of
fluxes, exchanges, and transformations. For example, connectivity can be reduced by dams,
levees, culverts, water withdrawals, and habitat destruction, and can be increased by effluent
discharges, channelization, drainage ditches and tiles, and impervious surfaces.
Conclusion 5, Cumulative Effects: Key Findings
• Structurally and functionally, stream-channel networks and the watersheds they drain are
fundamentally cumulative in how they are formed and maintained. Excess water from
precipitation that is not evaporated, taken up by organisms, or stored in soils and geologic
layers moves downgradient by gravity as overland flow or through channels carrying sediment,
chemical constituents, and organisms. These channels concentrate surface-water flows and are
more efficient than overland (i.e., diffuse) flows in transporting water and materials, and are
reinforced over time by recurrent flows.
• Connectivity between streams and rivers provides opportunities for materials, including
nutrients and chemical contaminants, to be transformed chemically as they are transported
downstream. Although highly efficient at the transport of water and other physical materials,
streams are dynamic ecosystems with permeable beds and banks that interact with other
ecosystems above and below the surface. The exchange of materials between surface and
subsurface areas involves a series of complex physical, chemical, and biological alterations that
occur as materials move through different parts of the river system. The amount and quality of
such materials that eventually reach a river are determined by the aggregate effect of these
sequential alterations that begin at the source waters, which can be at some distance from the
river. The opportunity for transformation of material (e.g., biological uptake, assimilation, or
beneficial transformation) in intervening stream reaches increases with distance to the river.
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Nutrient spiraling, the process by which nutrients entering headwater streams are transformed
by various aquatic organisms and chemical reactions as they are transported downstream, is
one example of an instream alteration that exhibits significant beneficial effects on downstream
waters. Nutrients (in their inorganic form) that enter a headwater stream (e.g., via overland
flow) are first removed from the water column by streambed algal and microbial populations.
Fish or insects feeding on algae and microbes take up some of those nutrients, which are
subsequently released back into the stream via excretion and decomposition (i.e., in their
organic form), and the cycle is repeated. In each phase of the cycling process—from dissolved
inorganic nutrients in the water column, through microbial uptake, subsequent transformations
through the food web, and back to dissolved nutrients in the water column—nutrients are
subject to downstream transport. Stream and wetland capacities for nutrient cycling have
important implications for the form and concentration of nutrients exported to downstream
waters.
Cumulative effects across a watershed must be considered when quantifying the frequency,
duration, and magnitude of connectivity, to evaluate the downstream effects of streams and
wetlands. For example, although the probability of a large-magnitude transfer of organisms
from any given headwater stream in a given year might be low (i.e., a low-frequency connection
when each stream is considered individually), headwater streams are the most abundant type of
stream in most watersheds. Thus, the overall probability of a large-magnitude transfer of
organisms is higher when considered for all headwater streams in a watershed—that is, a high-
frequency connection is present when headwaters are considered cumulatively at the
watershed scale, compared with probabilities of transport for streams individually. Similarly, a
single pollutant discharge might be negligible but the cumulative effect of multiple discharges
could degrade the integrity of downstream waters. Riparian open waters (e.g., oxbow lakes),
wetlands, and vegetated areas cumulatively can retain up to 90% of eroded clays, silts, and
sands that otherwise would enter stream channels. The larger amounts of snowmelt and
precipitation cumulatively held by many wetlands can reduce the potential for flooding at
downstream locations. For example, wetlands in the prairie pothole region cumulatively stored
about 11-20% of the precipitation in one watershed.
The combination of diverse habitat types and abundant food resources cumulatively makes
floodplains important foraging, hunting, and breeding sites for fish, aquatic life stages of
amphibians, and aquatic invertebrates. The scale of these cumulative effects can be extensive;
for example, coastal ibises travel up to 40 km to obtain food from freshwater floodplain
wetlands for nesting chicks, which cannot tolerate salt levels in local food resources until they
fledge.
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CLOSING COMMENTS
The structure and function of downstream waters highly depend on materials—broadly defined as any
physical, chemical, or biological entity—that originate outside of the downstream waters. Most of the
constituent materials in rivers, for example, originate from aquatic ecosystems located upstream in the
drainage network or elsewhere in the drainage basin, and are transported to the river through
flowpaths illustrated in the introduction to this report. Thus, the effects of streams, wetlands, and open
waters on rivers are determined by the presence of (1) physical, chemical, or biological pathways that
enable (or inhibit) the transport of materials and organisms to downstream waters; and (2) functions
within the streams, wetlands, and open waters that alter the quantity and quality of materials and
organisms transported along those pathways to downstream waters.
The strong hydrologic connectivity of river networks is apparent in the existence of stream channels
that form the physical structure of the network itself. Given the evidence reviewed in this report, it is
clear that streams and rivers are much more than a system of physical channels for efficiently conveying
water and other materials downstream. The presence of physical channels, however, is a compelling line
of evidence for surface-water connections from tributaries, or water bodies of other types, to
downstream waters. Physical channels are defined by continuous bed-and-bank structures, which can
include apparent disruptions (such as by bedrock outcrops, braided channels, flow-through wetlands)
associated with changes in the material and gradient over and through which water flows. The
continuation of bed and banks downgradient from such disruptions is evidence of the surface
connection with the channel that is upgradient of the perceived disruption.
Although currently available peer-reviewed literature does not identify which types of non-floodplain
wetlands have or lack the types of connections needed to convey functional effects to downstream
waters, additional information (e.g., field assessments, analysis of existing or new data, reports from
local resource agencies) could be used in case-by-case analysis of non-floodplain wetlands. Importantly,
information from emerging research into the connectivity of non-floodplain wetlands, including studies
of the types identified in Section 4.5.2 of this report, could close some of the current data gaps in the
near future. Recent scientific advances in the fields of mapping, assessment, modeling, and landscape
classification indicate that increasing availability of high-resolution data sets, promising new
technologies for watershed-scale analyses, and methods for classifying landscape units by hydrologic
behavior can facilitate and improve the accuracy of connectivity assessments. Emerging research that
expands our ability to detect and monitor ecologically relevant connections at appropriate scales,
metrics to accurately measure effects on downstream integrity, and management practices that apply
what we already know about ecosystem function will contribute to our ability to identify waters of
national importance and maintain the long-term sustainability and resiliency of valued water resources.
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1.1 Purpose
The objective of the Clean Water Act is to restore and maintain the chemical, physical, and biological
integrity of the nation's waters. The U.S. Environmental Protection Agency's (U.S. EPA's) Office of
Research and Development developed this report to inform rulemaking by the U.S. EPA and U.S. ACE on
the definition of "waters of the United States" under the Clean Water Act (CWA). Its purpose is to
summarize current scientific understanding about the connectivity and mechanisms by which streams
and wetlands, singly or in aggregate, affect the physical, chemical, and biological integrity of
downstream waters. Because this report is a technical review of peer-reviewed scientific literature, it
does not consider or set forth legal standards for CWA jurisdiction. Rather, the report evaluates,
summarizes, and synthesizes the available peer-reviewed scientific literature to address three
questions:
1. What are the physical, chemical, and biological connections to and effects of ephemeral,
intermittent, and perennial streams on downstream waters (e.g., rivers, lakes, reservoirs,
estuaries)?
2. What are the physical, chemical, and biological connections to and effects of riparian or
floodplain wetlands and open waters (e.g., riverine wetlands, oxbow lakes) on downstream
waters?
3. What are the physical, chemical, and biological connections to and effects of wetlands and open
waters in non-floodplain settings (e.g., most prairie potholes, vernal pools) on downstream
waters?
These questions were developed in collaboration with the U.S. EPA's Office of Water to translate
regulatory questions and terminology into more scientifically relevant questions and terms (Table 1-1).
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This report focuses on the physical, chemical, and biological connections (or lack thereof) by which
small or temporary streams, nontidal wetlands, and certain open waters can affect the integrity of
downstream waters.
In addition to a broad survey of literature responding to the three questions above, the U.S. EPA's Office
of Water asked the Office of Research and Development to create six case studies with more detailed
reviews of published literature on Carolina and Delmarva bays, oxbow lakes, prairie potholes, prairie
streams, southwestern streams, and vernal pools.
Table 1-1. Translating connectivity-related questions between policy and science. This table presents a
crosswalk of regulatory and scientific questions this report addresses. Policy questions use regulatory
terms (shown in quotation marks) that lack scientific definitions or are defined differently in scientific
usage. All terms used in this report reflect scientific definitions and usage.
Policy question
What tributaries have a "significant* nexus" to
"traditional navigable waters"?
What "adjacent" waters have a "significant* nexus" to
"traditional navigable waters"?
What categories of "other waters" have a "significant*
nexus" to "traditional navigable waters"?
Science question
What are the connections to and effects of ephemeral,
intermittent, and perennial streams on downstream
waters?
What are the connections to and effects of riparian or
floodplain wetlands and open waters on downstream
waters?
What are the connections to and effects of wetlands
and open waters in non-floodplain settings on
downstream waters?
* "Significant," as used here, is a policy determination informed by science; it does not refer to statistical significance.
1.2 Scientific Context
1.2.1 Concepts of Connectivity in Hydrology and Ecology
Streams, wetlands, and other surface waters interact with ground water and terrestrial environments
throughout the landscape, from the mountains to the oceans. Thus, an integrated perspective of the
landscape, described in this section, provides the appropriate scientific context for evaluating and
interpreting evidence about the physical, chemical, and biological connectivity of streams, wetlands, and
open waters to downstream waters.
Connectivity has long been a central tenet for the study of aquatic ecosystems. The River Continuum
Concept (Vannote et al., 1980) viewed the entire length of rivers, from source to mouth, as a complex
hydrologic gradient with predictable longitudinal patterns of ecological structure and function. The key
pattern is that downstream communities are organized, in large part, by upstream communities and
processes (Vannote et al., 1980; Battin et al., 2009). The Serial Discontinuity Concept (Ward and
Stanford, 1983) built on the River Continuum Concept to improve our understanding of how dams and
impoundments disrupt the longitudinal patterns of flowing waters with predictable downstream effects.
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The Spiraling Concept (Webster and Patten, 1979; Newbold et al., 1981; Elwood et al., 1983) described
how river network connectivity can be evaluated and quantified as materials cycle from dissolved forms
to transiently stored forms taken up by living organisms, then back to dissolved forms, as they are
transported downstream (Section 3.4.1). These three conceptual frameworks focused on the
longitudinal connections of river ecosystems, whereas the subsequent flood pulse concept (Junk et al.,
1989) examined the importance of lateral connectivity of river channels to floodplains, including
wetlands and open waters, through seasonal expansion and contraction of river networks. Ward (1989)
summarized the importance of connectivity to lotic ecosystems along four dimensions: longitudinal,
lateral, vertical (surface-subsurface), and temporal connections; he concluded that running water
ecosystems are open systems that are highly interactive with both contiguous habitats and other
ecosystems in the surrounding landscape. As these conceptual frameworks illustrate, scientists have
long recognized the hydrologic connectivity that the physical structure of river networks represents.
More recently, scientists have incorporated this connected network structure into conceptual
frameworks describing ecological patterns in river ecosystems and the processes linking them to other
watershed components, including wetlands and open waters (Power and Dietrich, 2002; Benda et al.,
2004; Nadeau and Rains, 2007; Rodriguez-Iturbe et al., 2009). The Network Dynamic Hypothesis (Benda
et al., 2004) is a physically based framework for predicting patterns of habitat heterogeneity observed
along a river, based on dynamics that generate potential biological "hotspots" at tributary confluences. It
essentially reexamines earlier, linearly driven frameworks given the patchy and stochastic nature of
lotic ecosystems (e.g., Resh etal., 1988; Townsend, 1989; Rice etal., 2001), and thus reflects a more
realistic river network perspective. Bunn and Arthington (2002) identified natural flow variability and
associated lateral and longitudinal connectivity of stream channels and floodplains as two principal
mechanisms linking hydrology to aquatic biodiversity of riverine species (also Leigh et al., 2010). In
addition, application of metapopulation theory and population genetic theory to natural populations has
greatly improved our understanding of the role of dispersal and migration in the demographic
persistence, community assembly, and evolution of aquatic species (Hastings and Harrison, 1994;
Moilanen and Hanski, 1998; Hanski, 1999; Pannell and Charlesworth, 2000; Pagan, 2002; Bohonak and
Jenkins, 2003; Waples, 2010; Fronhofer et al., 2012). Sheaves (2009) emphasized the key ecological
connections—which include process-based connections that maintain habitat function (e.g., nutrient
dynamics, trophic function) and movements of individual organisms—throughout a complex of
interlinked freshwater, tidal wetland, and estuarine habitats as critical for the persistence of aquatic
species, populations, and communities over the full range of time scales.
1.2.2 Connectivity Gradients and Descriptors
The landscape and flowpath perspectives illustrated in Figure 1-1 draw heavily from the connectivity
frameworks described in Section 1.2.1. These perspectives are essential to understanding connections
from streams, wetlands, and open waters that affect the integrity of downstream waters. Connectivity is
defined here as the degree to which components of a watershed are joined and interact by transport
mechanisms that function across multiple spatial and temporal scales (Section 2.3.2.1). The primary
transport mechanisms considered in this report are surface-water and shallow ground-water flows,
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Table 1-2. Dimensions of watershed connectivity.
Dimension
Examples and flowpaths in Figure 1-1 or Figure 1-2
Longitudinal
Streamflow and downstream transport of materials, organisms (1-1A); hyporheic flow
(1-1A); ground-water flow through local and larger scale aquifers (1-1A), aquatic or
overland movement of organisms in or along stream channels (1-1B); biogeochemical
transport and transformation (1-1B) (Alexander etal., 2007; Freeman etal., 2007)
Lateral
Overbank flow and transport from channels into banks, floodplains, and riparian areas
(1-1A); spillage and transport from wetlands and open waters into streams (1-1A);
overland flow and interflow (1-1A); ground-water recharge from streams and wetlands
(1-1A); bank storage (1-1A); transport or movement of organisms between streams and
wetlands or open waters (1-1B) (Ward, 1989; Stanford and Ward, 1993)
Vertical
Surface-subsurface exchange of water, materials, organisms (1-lAand 1-1B); ground-
water recharge from streams and wetlands (1-1A); atmospheric losses (1-1A) (Amoros
and Bornette, 2002; Banks etal., 2011)
Temporal
Variable source area (1-2); seasonal cycles of wetland inundation and outflow to
streams (1-1A); migration or diapause of aquatic organisms (1-1B) (Hewlett and
Hibbert, 1967; Bohonak and Jenkins, 2003; Zedler, 2003)
transport and transformation of physical and chemical materials, and movements of aquatic and
semiaquatic organisms, all of which connect watersheds in four dimensions (Table 1-2). Figure 1-1
illustrates the continuous hydrologic flowpaths (Figure 1-1A) and biological pathways (Figure 1-1B)
that connect watershed components spatially; Figure 1-2 illustrates the temporal dynamics of
hydrologic flowpaths (Sections 2.2.3 and 2.3.2.2).
Although all parts of a watershed are connected to some degree—by the hydrologic cycle or dispersal of
organisms, for example—the degree of connectivity among aquatic components varies along a
continuum from highly connected to highly isolated. This continuum can be described in terms of the
frequency, duration, magnitude, timing, and rate of change (Poff et al., 2007) of physical and
chemical fluxes to and biological exchanges with downstream waters. These terms, which we refer to
collectively as connectivity descriptors, characterize the range over which streams and wetlands vary
and shift along the connectivity gradient in response to changes in natural and anthropogenic factors
and, when considered in a watershed context, can be used to predict probable effects of different
degrees of connectivity over time. These and similar descriptors are used in hydrology and disturbance
ecology to characterize the variability and alteration of natural flow regimes (Resh et al., 1988; Poff,
1992; Poff etal., 1997; Lake, 2000; Leibowitz etal., 2008). For example, in hydrology, magnitude is the
amount of water moving past a fixed location per unit time, frequency is how often a particular flow
magnitude occurs, duration is a measure of how long a particular flow magnitude persists, and rate of
change is how quickly one type of flow changes to another. Because the presence of water determines
hydrologic connectivity, these descriptors also can be used to describe the timing and magnitude of
hydrologic connections. Further, they can describe other types of connections. The number of
individuals immigrating or emigrating during a dispersal event, for example, could be used to determine
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Figure 1-1A. Hydrologic f lowpaths. Arrows are representative of surface-water and ground-water flows occurring throughout the watershed.
Subsurface flows are shown within the cross section, and by faded arrows outside the cross section.
1-3
§ ffl
If
n
ro
o
Stream and wet/and recharge of
local and regional aquifers
Key
- Perennial stream
.... Ephemeral stream
Intermittent stream
Precipitation
Unsaturatedzone
Saturated zone
Local aquifer and hyporheic zone
Confining layer
Atmospheric losses
(e.g., evapotranspiration, volatilization, denitrification)
Wetland (dry period)
Open water (dry period)
Expansion and overbank flow into floodplain and overflow of
wetlands and open-waters during wetter periods
Water table level
V
Ground-water flow through local and larger scale aquifers
Overland and interflow
Streamflow and transport of materials, organisms
Overbank flow and transport or spillage of materials, organisms
Bank storage
Hyporheic flow and surface-subsurface exchange of water, materials, organisms
-------
Figure 1-1B. Biological flowpaths. Arrows are representative of biological pathways occurring throughout the watershed. This figure also
includes representative biogeochemical pathways occurring in streams and floodplains.
1-3
§ ffl
If
n
(/> o
Key
— Perennial stream
•••- Ephemeral stream
- Intermittent stream
Wetland (dry period)
Open water (dry period)
Unsaturatedzone
Saturated zone
Local aquifer and hyporheic zone
Confining layer
Expansion and overbank flow into floodplain and overflow of
wetlands and open-waters during wetter periods
Water table level
Aquatic transport or movement
Overland transport or movement (aerial or terrestrial)
Surface-subsurface exchange of water, materials, organisms
Biochemical transformation and transport (e.g., nutrient spiraling)
-------
Figure 1-2. Temporal dynamics of hydrologic flowpaths. (A) A riverscape at peak hydrologic expression.
(B) The same riverscape in a dry period. Intermittent and ephemeral streams, and some wetlands, are dry.
Key
Perennial stream
Unsaturated zone
Saturated zone
Local aquifer and
hyporheiczone
Confining layer
Onoint Ephemeral stream
^"^Y Intermittent stream
- Precipitation
„. ^ Atmospheric losses
(e.g., evapotranspiration,
volatilization, denitrification)
Wetland (dry period)
Open water (dry period)
Expansion and overbankflow intofloodplain and over-
flow of wetlands and open-waters during wetter periods
Dry wetland
Water table level
Ground-water flow through local and larger
scale aquifers
Overland and interflow
Streamflow and transport of
materials, organisms
Overbank flow and transport or spillage
of materials, organisms
Bank storage
Hyporheicflow and surface-subsurface
exchange of water, materials, organisms
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the magnitude of the event; the probability, length, and predictability of similar events could be
expressed in terms of their frequency, duration, and timing; and fluctuations in dispersal could be
described as the rate of change through time (e.g., across seasons or years).
Stream and wetland connections have particularly important consequences for downstream water
integrity. Longitudinal flows through ephemeral, intermittent, and perennial stream channels
(Figure 1-1A, blue lines and arrows) are much more efficient for transport of water, materials, and
organisms than diffuse overland flows and interflows (Figure 1-1A, green arrows). Over time, stream
transport path ways are reinforced by recurrent flows that maintain channel form. Areas that
concentrate water also provide mechanisms for storage, transformation, and transport of materials.
Differences in flow frequency, duration, magnitude, timing, and rate of change (e.g., rapid flow in
mountain streams, slow flow through glacial ice or bedrock, intermittent flow in seasonal streams,
ephemeral flow in arid rivers) create conditions needed for a range of ecosystem functions that affect
downstream waters. Such functions include short- and long-term storage of water and sediment,
transformation or sequestration of contaminants, recycling of excess nutrients, provision of habitat for
aquatic and semiaquatic species, recharge of river baseflow, and provision of drinking water for humans
and wildlife. For example, areas that are prone to wetting and drying cycles in response to seasonal
conditions (e.g., stream and wetland perimeters shown in Figure 1-1A) are "hotspots" for chemical
transformations (Vidon etal., 2010).
Ultimately, differences in the frequency, duration, magnitude, timing, and rate of change of physical,
chemical, and biological connections describe different positions along the connectivity gradient and
produce different types of downstream effects. For example, highly connected stream channels convey
water and channel-forming sediment to rivers, whereas highly isolated wetlands can reduce flooding
and store excess sediment. Connections with low values of one or more descriptors (e.g., low-frequency,
short-duration flooding) can have important downstream effects when values for other descriptors are
high (e.g., large-magnitude downstream transfer of floodwaters, sediment, large woody debris, and
organisms). At the other end of the frequency gradient, high-frequency, low-magnitude vertical and
lateral flows (Table 1-2) contribute to aquatic biogeochemical processes, including nutrient and
contaminant transformation and organic matter accumulation (e.g., Brunke and Gonser, 1997; Karwan
and Saiers, 2012; Lawrence et al., 2013).
In addition, timing is a key connectivity descriptor that can influence downstream waters. For example,
when soils are saturated by previous rainfall events, even low or moderate rainfall can cause streams or
wetlands to overflow, transporting water and materials to downstream waters. The same wetland or
wetland type can attenuate floods or generate floods, depending on hydrologic conditions (Acreman and
Holden, 2013). Predictable events also can profoundly influence the effects of connections. Wetlands and
river networks expand and contract in response to seasonal and decadal cycles and longer term changes
in environmental conditions. In wet conditions (Figure 1-2A), streams and rivers expand longitudinally
into headwaters and laterally into floodplains or riparian areas, wetlands inundate and connect via
surface water and ground water to other wetlands and the stream network, the water table rises, and
local aquifers are recharged. In dry conditions (Figure 1-2B), the river network is limited to perennial
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streams, wetlands dry down, and the water table level lowers. Seasonal flooding and drying events over
an annual cycle are formative processes of physical, chemical, and biological attributes of streams in the
western United States (Gasith and Resh, 1999). Large seasonal waterfowl migrations can move
nutrients, plants (seeds), and invertebrates between wetlands and downgradient waters (Figuerola and
Green, 2002; Green and Figuerola, 2005; Frisch et al., 2007). Fish that use nonperennial or perennial
headwater stream habitats to spawn or rear young, and invertebrates that move into seasonally
inundated floodplain wetlands prior to emergence, have life cycles that are synchronized with the timing
of flows and flood pulses, temperature thresholds, and food resource availability in those habitats (Junk
et al., 1989; Falke et al., 2010).
The surface-water and ground-water flowpaths along which water and materials are transported and
transformed (Sections 2.2.2, 3.3, 3.4, 4.3.2, 4.3.3, 4.4.2, and 4.4.3; Figure 1-1A) determine variations in
the degrees of physical and chemical connectivity. These flowpaths are controlled primarily by variation
in climate, geology, and terrain within and among watersheds and over time. These factors have been
used to group watersheds into hydrologic landscapes units that, although not necessarily spatially
contiguous, are predicted to exhibit similar hydrologic function (Wolock et al., 2004; Wigington et al.,
2013). Climate, geology, and terrain are reflected locally in factors such as rain and snowfall intensity,
soil infiltration rates, and the direction of ground-water flows. These local factors interact with stream
and wetland function and landscape position to influence degrees of connectivity through time and
across space. When considered together with these local factors, hydrologic landscapes could provide a
regional context for evaluating the physical and chemical connectivity of streams and wetlands in a
particular watershed or group of watersheds (Section 2.4.1).
Gradients of biological connectivity (i.e., the active or passive movements of organisms through water
and air and over land that connect populations of aquatic species; Sections 3.5, 4.3.4, and 4.4.4; Figure
1-1B) are determined primarily by species assemblages and by landscape features, including the factors
discussed above, that facilitate or impede the movement of organisms. Organisms move across the
landscape to colonize new habitats, avoid inbreeding, escape predation or competition, locate mates,
and acquire resources needed to survive and reproduce. The temporal and spatial scales at which
biological pathways connect aquatic habitats depend on characteristics of both the landscape and
species, and overland transport or movement can occur across watershed boundaries. Dispersal is
essential at higher levels of biological organization for population persistence, maintenance of genetic
diversity, and evolution of aquatic species (Labbe and Fausch, 2000; Pagan, 2002; Malmqvist, 2002;
Bohonak and Jenkins, 2003; Armsworth and Roughgarden, 2005). Consequently, dispersal strategies
reflect aquatic species' responses and adaptations to biotic and abiotic environments, including spatial
and temporal variation in resource availability and quality (e.g., Clobert et al., 2009). Dispersal-related
traits and behaviors (e.g., habitat specialization, dispersal mode, behavioral response to environmental
cues) therefore encompass species-environment relationships over time and provide an ecological and
evolutionary context for evaluating biological connectivity in a particular watershed or group of
watersheds.
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Pathways for chemical transport and transformation largely follow hydrologic flowpaths (Figure 1-1A),
but sometimes follow biological pathways (e.g., nutrient transport from wetlands to coastal waters by
migrating waterfowl, upstream transport of marine-derived nutrients by anadromous fish, uptake and
removal of nutrients by emerging stream insects; Figure 1-1B). The transport and transformation of
nutrients (e.g., sequential transformations, Section 2.3.2.1; and nutrient spiraling in streams, Section
3.4.1) and other chemicals associated with water integrate physical, chemical, and biological
connectivity of streams and wetlands to downstream waters (Figure 1-1B).
1.2.3 Cumulative Effects of Streams and Wetlands on Downstream
Waters
Stream and wetland connectivity to downstream waters, and the resulting effects on downstream water
integrity, must be considered cumulatively. First, when considering the effect of an individual stream or
wetland, including the cumulative effect of all the contributions and functions that a stream or wetland
provides is essential. For example, the same stream transports water, removes excess nutrients,
mitigates flooding, and provides refuge for fish when conditions downstream are unfavorable; ignoring
any of these functions would underestimate the overall effect of that stream.
Secondly, stream channel networks and the watersheds they drain are fundamentally cumulative in how
they are formed and maintained. Excess precipitation that is not evaporated, taken up by organisms, or
stored in soils and geologic layers moves downgradient as overland flow or through channels, which
concentrate flows and carry sediment, chemical constituents, and organisms (Sections 3.3, 3.4, and 3.5).
As flows from numerous headwater channels combine in larger channels, the volume and effects of
those flows accumulate as they move through the river network. As a result, the incremental
contributions of individual streams and wetlands accumulate in the downstream waters. Important
cumulative effects are exemplified by ephemeral flows in arid landscapes, which are key sources of
baseflow for downgradient waters (Sections 5.6 andB.5; Schlesinger and Jones, 1984; Baillie etal., 2007;
Izbicki, 2007), and by the high rates of denitrification in headwater streams (Section 3.4.1). The amount
of nutrients removed by any one stream over multiple years or by all headwater streams in a watershed
in a given year can have substantial consequences for downstream waters (Alexander et al., 2007;
Alexander et al., 2009; Bb'hlke et al., 2009; Helton et al., 2011). Similar cumulative effects on
downstream waters have been documented for other material contributions from headwater streams
(Chapter 3). For example, although the probability of a large-magnitude transfer of organisms from any
given headwater stream in a given year might be low (i.e., a low-frequency connection when each stream
is considered individually), headwater streams are the most abundant type of stream in most
watersheds (Section 3.2). Thus, the overall probability of a large-magnitude transfer of organisms is
higher when considered for all headwater streams in a watershed—that is, there is a high-frequency
connection when considered cumulatively at the watershed scale, compared with probabilities of
transport for streams individually. Similarly, a single pollutant discharge might be negligible but the
cumulative effect of multiple discharges could degrade the integrity of downstream waters.
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Evaluating cumulative contributions over time is critical in streams and wetlands with variable degrees
of connectivity. For example, denitrification in a single headwater stream in any given year might not
affect downstream waters; over multiple years, however, this effect could accumulate. Western vernal
pools provide another example of cumulative effects over time. These pools typically occur as complexes
in which the hydrology and ecology are tightly coupled with the local and regional geological processes
that formed them (Section B.6). When seasonal precipitation exceeds wetland storage capacity and
wetlands overflow into the river network and generate stream discharge, the vernal pool basins, swales,
and seasonal streams function as a single surface-water and shallow ground-water system connected to
the river network.
1.2.4 Effects of Human Activities on Connectivity
Human activities alter naturally occurring gradients of physical, chemical, and biological connectivity by
modifying the frequency, duration, magnitude, timing, and rate of change effluxes, exchanges, and
transformations. For example, all dimensions of connectivity (Table 1-2) can be reduced by dams and
levees (Ward and Stanford, 1983; Ligon et al., 1995; Collier et al., 1996; Wohl, 2005; Franklin et al.,
2009), water withdrawals (Haag and Pfeiffer, 2012), and habitat destruction. Alternatively, connectivity
can be increased by point source discharges (Brooks et al., 2006); channelization, drainage ditches, and
tiles (Randall et al., 1997; Min et al., 2010); and storm drains and impervious surfaces (Booth, 1990; Paul
and Meyer, 2001; Elmore and Kaushal, 2008; Walsh etal., 2012). The effects of human activities on
connectivity are often complex. For example, a levee will decrease connectivity between a river channel
and its floodplain at the levee site, but might increase connectivity of the channel and floodplain farther
downstream, due to increased flow. Similarly, drainage ditches that increase hydrologic connectivity
between isolated aquatic systems also can decrease biological connectivity through habitat loss and
fragmentation.
Human activities modify the natural biological processes, material fluxes, and energy fluxes that link
watershed components, resulting in a suite of stressors with measurable effects on downstream
ecosystems. Some of these activities are illustrated in a hypothetical watershed (Figure 1-3). In
Figure 1-3 (A), buried and ditched streams have eliminated aquatic habitat, increased downstream
export of runoff and contaminants, and eliminated stream functions that could benefit downstream
water quality. Figure 1-3 (B) shows a dam and reservoir that have constrained natural river expansion
and contraction cycles by increasing water storage, trapping sediment, and regulating the volume and
timing of river discharge. Dams and reservoirs also block upstream movement of migrating fish and
other organisms, alter riparian areas, and impair riparian and floodplain wetland functions. In Figure
1-3 (C), levees and channelization have disconnected the river from its floodplain; decreased exchange
of water, materials, and biota between the channel bed and hyporheic zone; and eliminated stream and
wetland habitats. In addition, levees decrease the volume of river discharge at the levee site, but
increase discharge downstream of the levee site. In Figure 1-3 (D), urban stormwater drainage has
increased export of runoff and contaminants from impervious surface areas, altered stream
temperature, and impaired instream habitats. In Figure 1-3 (E), drained and ditched wetlands have
impaired wetland habitat and functions; increased downstream export of excess nutrients and
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Figure 1-3. Effects of human alterations on watershed connectivity. See Section 1.2.4 for description of alterations illustrated in A-G.
f
I 3
a
~ s
<2. O
(/) o
Key
Perennial stream
•••• • Ephemeral stream
Unsaturatedzone
Saturated zone
' - i Effluent (
^^^ Ditch
Wetland
A
B
Si Confining layer
Buried and ditched streams
Dam and reservoir
c
D
E
Levees and channelization
Urban stormwater drainage
Drained and ditched wetlands
F
G
Ground-water withdrawal
Effluent dominated stream
ro
o
-------
Figure 1-4. The role of connectivity in maintaining the physical, chemical, and biological integrity of
water. Climate, landscape, and species' traits (Influencing Factors) interact to form Connections
(hydrologic, chemical, and biological) that control the frequency, duration, magnitude, timing, and rate
of change of material and energy fluxes, and biological dynamics (Processes) linking watershed
components. The Functions by which these connections affect downstream waters modify the timing
of transport and the quantity and quality of resources available to downstream communities.
Biomonitoring programs have developed structural metrics for assessing physical habitat, water
quality, and biological assemblages as indicators of the physical, chemical, and biological integrity of
downstream waters (Assessment Endpoints and Metrics).
FUNCTIONS
CLIMATE FACTORS
• Annual watersurplus
• Seasonality
• Rainfall intensity
• Temperature
LANDSCAPE FACTORS
• Topography
• Landform
• Soil type
• Aquifer permeability
• Spatial distribution
SPECIES' TRAITS
• Life cycle
• Dispersal capability
• Dispersal cues
• Dispersal behavior
HUMAN ACTIVITIES
INFLUENCING FACTORS -» CONNECTIONS -»
PHYSICAL & CHEMICAL
INTEGRITY
• Habitatquality
• Waterquality
• Toxicity
BIOLOGICAL INTEGRITY
• Community structure
• Indicatorspecies
• Functional groups
• Population attributes
ECOSYSTEM INTEGRITY,
SUSTAINABILITY, RESILIENCY
PROCESSES ->
EFFECTS
ASSESSMENT
ENDPOINTS & METRICS
other contaminants; and decreased recharge of local and regional aquifers. In Figure 1-3 (F), ground-
water withdrawal has lowered the water table, disconnecting surface water and ground water, thereby
causing local streams and wetlands to dry. Finally, in Figure 1-3 (G), pollutant discharges into effluent-
dominated streams have altered the volume and timing of streamflow, and increased the export of
contaminants into streams. Because watersheds typically experience multiple covarying stressors,
determining the cause of a specific downstream effect can be difficult. Relating observed effects to
probable causes requires not only reliable measures of candidate stressors and observed effects, but
also a clear understanding of the intermediate processes that link them mechanistically (U.S. EPA, 2010;
Farraretal., 2014).
Multiple indicators and measures have been proposed for detecting and quantifying changes in
connectivity associated with human activities (With etal., 1997; Tischendorf and Fahrig, 2000; Moilanen
and Nieminen, 2002; Calabrese and Pagan, 2004; Martin and Soranno, 2006; Fullerton et al., 2010;
Hermoso etal., 2012). Impairments that result from structural alteration of landscape attributes (e.g.,
dam construction, channel incision, loss of overland dispersal corridors) are relatively easier to detect
and quantify than impairments of functional processes (e.g., altered nutrient dynamics, reduced gene
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flow), but both have important consequences for the short- and long-term integrity of freshwater
ecosystems. Palmer and Febria (2012) proposed that ecosystem impairment can be better identified and
diagnosed by a combination of structural and functional metrics than by either type alone. Because
connectivity can be defined in both structural and functional terms and is an integral component of
aquatic ecosystem integrity, this approach is more appropriate for detecting and assessing effects of
altered connectivity. To this end, systematic approaches that are rooted in landscape analysis and which
incorporate hydroecological dynamics present in streams and wetland complexes (Section 2.4.6) are
likely to provide useful information for inferring when and where altered connectivity is a cause of
impairment to water resources.
1.3 Report Approach
In this report, we focus entirely on peer-reviewed, publicly accessible sources of information about
surface-water and ground-water (particularly shallow ground-water) connections and interactions from
streams, wetlands, and open waters that influence the function and condition of downstream surface
waters (Figure 1-5). Information about connections among water bodies of the same type (e.g., wetland-
to-wetland or headwater stream-to-headwater stream connections) and connections from terrestrial
systems to downstream waters are considered out of scope (Figure 1-5).
The topical scope of this report was chosen to consider waters that often fall under the purview of the
CWA. As a scientific review, however, this report does not consider or make judgments regarding legal
standards for CWA jurisdiction. Our review of subsurface flows emphasizes shallow (local) ground
water, because flows in this category have the greatest interchange with surface waters (Winter et al.,
1998) although relevant surface-subsurface exchanges occur at depths ranging from centimeters to tens
of meters, depending on geographic location, stream channel geometry, and other factors (Woessner,
2000). As with any literature review, readers should refer to the cited publications for quantitative
information, such as flow distance, depth, duration, timing, and magnitude, about specific surface-water
and ground-water connections, and for other details about the systems and studies discussed in this
report.
To identify connections and effects of streams, wetlands, and other water bodies on downstream waters,
we used two types of evidence from peer-reviewed, published literature: (1) direct evidence that
demonstrated a connection or effect (e.g., observed transport of materials or movement of organisms
from streams or wetlands to downstream waters) and (2) indirect evidence that suggested a connection
or effect (e.g., presence of environmental factors known to influence connectivity, a gradient of
impairment associated with cumulative loss of streams or wetlands). In some cases, an individual line of
evidence demonstrated connections along the entire river network (e.g., from headwaters to large
rivers). In most cases, multiple sources of evidence were gathered and conclusions drawn via logical
inference—for example, when one body of evidence shows that headwater streams are connected to
downstream segments, another body of evidence shows those downstream segments are linked to other
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Figure 1-5. Waters and connections considered to be within scope for this report.
fl
I 3
If
n
Chapter 3: Transport and transformation of water,
materials, and biota from ephemeral, intermittent, and
perennial streams
Chapters: Movement of biota between
tributaries and downstream waters
Key
Chapter4: Non-floodplain
wetlands and open-waters
with outlets to streams
Chapters 3 and 4: Lateral
exchange of water,
materials, and biota
between waters in
channels andf/oodp/ains
Chapter 4: Non-floodplain
wetland and open-water fill-
and-spill into streams
Chapter 4: Overland
movement of biota from
wetlands to streams
Chapter 3: Longitudinal hydrologic
flowpaths: surface and shallow
subsurface streamflow
—- Perennial stream
"•• Ephemeral stream
~ Intermittent stream
Wetland (dry period)
Open-water (dry period)
Unsaturatedzone
^^^, Saturated zone
^_ ^, Local aquifer and hyporheic zone
'ffl-y!* Confining layer
Chapter 3: Vertical exchange of water,
materials, biota between surface and
shallow subsurface waters
Expansion and overbank flow into floodplain and overflow of
wetlands and open-waters during wetter periods
Aquatic transport or movement of organisms
Overland transport or movement of organisms (aerial or terrestrial)
Overbank flow and transport or spillage of materials, organisms
^ Bank storage
t ^ Hyporheic flow and surface-subsurface exchange of water, materials, organisms
°
ui
-------
segments farther downstream, and so on. This approach, which borrows from weight-of-evidence
approaches in causal analysis (Suter et al., 2002; Suter and Cormier, 2011), is an effective way to
synthesize the diversity of evidence needed to address questions at larger spatial and longer temporal
scales than are often considered in individual scientific studies.
1.3.1 Selection and Screening of Scientific Materials
We searched the scientific literature for information on the types of waters, connections, and
downstream effects identified in the report objectives and scope (Section 1.1; Figure 1-5). We conducted
keyword searches using terms inclusive of the types of waters, connections, and downstream effects of
interest (e.g., [wetland* AND [river* OR stream*] AND [connect* OR isolat*]]). Because simple keyword
searches would have omitted relevant publications, we also searched for literature on related topics.
Topics included conceptual frameworks of watershed and landscape connectivity; hydrologic flowpaths
among watershed components; biogeochemical transformation and cycling in streams and wetlands;
natural or artificial tracers of difficult-to-observe flows (e.g., ground-water flow, gene flow); chemical
and biological processes associated with aquatic habitat fragmentation and spatial isolation; and climate
or landscape factors that influence connectivity or isolation. We also reviewed citations provided by
peer-review panels and in public comments on drafts of the report. We then screened those results and
selected the most relevant publications for review and synthesis in this report, based on the criteria in
Figure 1-6.
We used science citation databases and search engines available through Web of Science™ and Google
Scholar™ to search primary (original research) and secondary (review) literature. These searches
included examination of references citing or cited in relevant publications obtained through specific
searches.
Because the breadth and depth of topics covered in this report made an exhaustive literature review
impractical, we emphasized highly influential papers on relevant topics, review papers that summarized
multiple studies in narrative form, meta-analyses that used statistical methods to combine results from
multiple independent studies into a single evaluation of evidence, and superseding editions or versions
of published research. Publications that did not provide new information, an alternative perspective or
interpretation of evidence, or a technical improvement (e.g., improved accuracy or better study design)
were not summarized in the report to avoid redundancy and excessive length and detail.
We summarized the relevant literature in narrative form and organized each chapter into lines of
evidence pertaining to different types of connections (physical, chemical, biological) for different types
of systems (streams, riparian/floodplain wetlands, non-floodplain wetlands). Lines of evidence were
evaluated for strength, consistency, mechanistic plausibility, and relevance to the endpoints identified in
the report objectives. Finally, conclusions for each of the report's three questions were derived from the
key findings, and placed in context with concepts and evidence provided in each chapter.
Cited in this report are 1,353 references. Most were published in refereed scientific journals (86%), as
scientific reports by federal agencies that follow peer-review guidelines of the Office of Management and
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Figure 1-6. Flow chart for screening and compiling literature.
1. Search literature on streams,
wetlands, and open waters; review
citations provided by peer reviewers
and public comments.
2. For each publication under
consideration:
3. Does it
contain scientific
information about the physical,
chemical, or biological connectivity of
streams, wetlands, or open waters to
larger downstream waters (e.g., rivers,
lakes, coastal waters)?
4. Has it been
peer reviewed* or
verified for quality
assurance?
5. Does it present new
information on a relevant topic,
alternative interpretations of existing
information, or technical improvements
to studies already reviewed?
6. Add to compiled literature for
further evaluation.
* Peer review is the formal evaluation of scientific information by independent experts who were not involved in the work but
have equivalent scientific and technical expertise. Its purpose is to ensure that materials accepted for publication have been
critically reviewed and revised as needed to meet the documented standards of scientific integrity and qualityfor specific
journals or organizations. All reports published by the U.S. EPA Office of Research and Development meet or exceed peer-review
requirements established by the Office of Management and Budget (OMB, 2004).
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Budget (4%), or scientific books (~9%). The remaining citations refer to photographs, maps, non-federal
reports, or websites (<1%) that provide supplemental information.
1.3.2 Report Structure
The report is organized into six chapters. Chapter 1 outlines the purpose, scientific context, and
approach of the report. Chapter 2 describes the components of a river system and watershed; the types
of physical, chemical, and biological connections that link those components; the factors that influence
connectivity at various temporal and spatial scales; and methods for quantifying connectivity. Chapter 3
reviews literature on connectivity in stream networks in terms of physical, chemical, and biological
connections and their resulting effects on downstream waters. Chapter 4 reviews literature on the
connectivity and effects of nontidal wetlands and certain open waters on downstream waters. Chapter 5
applies concepts and evidence from previous chapters to the case studies detailed in Appendix B.
Chapter 6 presents the five major conclusions of this report, with a summary of key findings from the
literature synthesized to develop these conclusions. It also discusses the relative abundance of literature
on topics reviewed in this report, and briefly discusses emerging research that can close some current
data gaps identified in the report. A glossary of scientific terms used in the report and detailed case
studies of selected systems (summarized in Chapter 5) are included in Appendix A and Appendix B,
respectively.
1.4 Summary
This report evaluates, summarizes, and synthesizes available peer-reviewed scientific literature on the
connectivity and mechanisms by which streams, wetlands, and open waters, singly or in aggregate,
affect the physical, chemical, and biological integrity of downstream waters.
Connectivity has long been a central tenet for the study of aquatic ecosystems. Watersheds are
integrated at multiple spatial and temporal scales by flows of surface water and ground water, transport
and transformation of physical and chemical materials, and movements of organisms. Although all parts
of a watershed are connected, the degrees and downstream effects of those connections vary; the effects
also are influenced by characteristics of the physical environment, the biological environment, and by
human activities in the watershed.
Variation in the degree of connectivity is critical to the integrity and sustainability of downstream
waters, and can be described in terms of the frequency, duration, magnitude, timing, and rate of change
effluxes to and biological exchanges with downstream waters. These descriptors characterize the range
over which streams and wetlands vary and shift along connectivity gradients and the probable effects of
different types (hydrologic, chemical, biological) and degrees of connectivity over time. Gradients of
physical, chemical, and biological connectivity are controlled primarily by variation in climate, geology,
terrain, aquatic organisms, and human activities within and among watersheds, and over time.
Ultimately, differences in the frequency, duration, magnitude, timing, and rate of change of physical,
chemical, and biological connections describe different positions along the connectivity gradient and
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produce different types of downstream effects. Connections with low values of one or more descriptors
(e.g., low-frequency, short-duration floods) can have important downstream effects when values for
other descriptors are high (e.g., large-magnitude transfers of floodwaters, sediment, large woody debris,
and organisms downstream). Atthe other end of the frequency gradient, the effects of high-frequency,
low-magnitude vertical and lateral flows strongly contribute to biogeochemical functions, including
nutrient and contaminant transformation and organic matter accumulation.
Stream channel networks and the watersheds they drain are fundamentally cumulative in how they are
formed and maintained. The downstream consequences (e.g., the amount and quality of materials that
eventually reach a river) are determined by the aggregate effect of contributions and sequential
alterations that begin at the source waters and function along continuous flowpaths to the watershed
outlet. Cumulative effects across a watershed must therefore be considered when quantifying the
frequency, duration, and magnitude of connectivity, to evaluate the downstream effects of streams,
wetlands, and open waters.
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2.1 Introduction
A river is the time-integrated result of all waters contributing to it, and connectivity is the property that
spatially integrates the individual components of the watershed. In discussions of connectivity, the
watershed scale is the appropriate context for interpreting technical evidence about individual
watershed components (Newbold et al., 1982b; Stanford and Ward, 1993; Bunn and Arthington, 2002;
Power and Dietrich, 2002; Benda et al., 2004; Naiman et al., 2005; Nadeau and Rains, 2007; Rodriguez-
Iturbe et al., 2009). Such interpretation requires that freshwater resources be viewed within a
landscape—or systems—context (Baron et al., 2002). Addressing the questions asked in this report
(Section 1.1), therefore, requires an integrated systems perspective that considers both the components
contributing to the river and the connections between those components and the river. This chapter
describes this integrated systems perspective. Section 2.2 outlines the basic hydrologic foundation of
river systems. Section 2.3 provides a general overview of how streams and wetlands affect downstream
waters, focusing on functions within streams and wetlands and how they are connected to downstream
waters. Finally, Section 2.4 examines key factors that affect connectivity between streams and wetlands
and rivers. Although we focus our discussion here on interactions between streams, wetlands, and
rivers, similar exchanges of water, influenced by many of the same factors, also occur between rivers,
lakes, estuaries, and marine waters.
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2.2 An Introduction to River Systems
2.2.1 River System Components
In this report, the term river refers to a relatively large volume of flowing water within a visible
channel, including subsurface water moving in the same direction as the surface water and lateral flows
exchanged with associated floodplain and riparian areas (Naiman and Bilby, 1998). Channels are
natural or constructed passageways or depressions of perceptible linear extent that convey water and
associated materials downgradient. They are defined by the presence of continuous bed and bank
structures, or uninterrupted (but permeable) bottom and lateral boundaries. Although bed and bank
structures might in places appear to be disrupted (e.g., bedrock outcrops, braided channels, flow-
through wetlands), the continuation of the bed and bank downgradient from such disruptions is
evidence of the surface connection with the channel that is upgradientof the perceived disruption. Such
disruptions are associated with changes in the gradient and in the material over and through which the
water flows. If a disruption in the bed and bank structure prevented connection, the area downgradient
would lack a bed and bank, be colonized with terrestrial vegetation, and be indiscernible from the
nearby land. The concentrated longitudinal movement of water and sediment through these channels
lowers local elevation, prevents soil development, selectively transports and stores sediment, and
hampers the colonization and persistence of terrestrial vegetation. Streams are defined in a similar
manner as rivers: a relatively small volume of flowing water within a visible channel, including
subsurface water moving in the same direction as the surface water and lateral flows exchanged with
associated floodplain and riparian areas (Naiman and Bilby, 1998).
A river network is a hierarchical, interconnected population of channels that drains surface and
subsurface water (Sections 2.2.2 and 2.2.3) from a watershed to a river and includes the river itself.
Watershed boundaries traditionally are defined topographically, such as by ridges, but ground-water
sources and losses can occur outside of topographic boundaries (Winter et al., 2003). These channels
can convey water year-round, weekly to seasonally, or only in direct response to rainfall and snowmelt
(Frissell etal., 1986; Benda etal., 2004). The smallest of these channels, where streamflows begin, are
considered headwater streams. Headwater streams are first- to third-order streams (Vannote et al.,
1980; Meyer and Wallace, 2001; Gomi etal., 2002; Fritz etal., 2006b; Nadeau and Rains, 2007), where
stream order is a classification system based on the position of the stream in the river network (Figure
2-1; Strahler, 1957). The point at which stream or river channels intersect within a river network is
called a confluence (Figure 2-1). The confluence of two streams with the same order results in an
increase of stream order (i.e., two first-order streams join to form a second-order stream, two second-
order streams join to form a third-order stream, and so on); when streams of different order join, the
order of the larger stream is retained.
One weakness of classification based on stream order is that it disregards the contributions of lower
order streams where they join a higher order stream. Link magnitude, an alternative method for
classifying streams, resolves this issue. Link magnitude is the sum of all source streams draining into a
given stream segment (Scheidegger, 1965; Shreve, 1967). Therefore, unlike stream order, the link
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Figure 2-1. A generalized example of a river network within its watershed. Blue lines illustrate the
river network, within the light green area of its watershed. Numbers represent Strahler stream order,
with streams increasing in order when two streams of equal order join. Blue squares indicate channel
heads, and orange dots depict confluences.
magnitude of a segment accounts for all contributing lower order streams regardless of their position in
river networks. For some properties, link magnitude might better reflect the aggregate upstream
contributions to downstream waters.
Mock (1971) presented a classification of the streams comprising stream or river networks. He
designated first-order streams that intersect other first-order streams as sources. We refer to these as
terminal source streams. Mock defined first-order streams that flow into higher order streams as
tributary sources, and we refer to this class of streams as lateral source streams (Figure 2-1).
Terminal and lateral source streams typically originate at channel heads (Dietrich and Dunne, 1993),
which occur where surface-water runoff is sufficient to erode a definable channel. The channel head
denotes the upstream extent of a stream's continuous bed and bank structure (Figure 2-1). Channel
heads are relatively dynamic zones in river networks, as their position can advance upslope by overland
or subsurface flow-driven erosion, or retreat downslope by colluvial infilling. Source streams also can
originate at seeps or springs and associated wetlands.
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When two streams join at a confluence, the smaller stream (i.e., that with the smaller drainage area or
lower mean annual discharge) is called a tributary of the larger stream, which is referred to as the
mainstem. A basic way of classifying tributary contributions to a mainstem is the symmetry ratio,
which describes the size of a tributary relative to the mainstem at their confluence, in terms of their
respective discharges, drainage areas, or channel widths (Roy and Woldenberg, 1986; Rhoads, 1987;
Benda,2008).
Surface-water hydrologic connectivity within river network channels occurs, in part, through the
unidirectional movement of water from channels at higher elevations to ones at lower elevations—that
is, hydrologic connectivity exists because water flows downhill. In essence, the river network represents
the aboveground flow route and associated subsurface-water interactions, transporting water, energy,
and materials from the surrounding watershed to downstream rivers, lakes, estuaries, and oceans (the
River Continuum Concept; Vannote et al., 1980).
A river system (Figure 2-2) consists of a river network and its entire watershed. It includes all
connected or isolated surface-water bodies (e.g., lakes and wetlands), any ground-water flow systems
connecting the drainage basin with the river network and surface-water bodies, and terrestrial
ecosystems (Stanford and Ward, 1993; Naiman et al., 2005).
Streamflow and the quantity and character of sediment—interacting with watershed geology, terrain,
soils and vegetation—shape morphological changes in the stream channel that occur from river network
headwaters to lower rivers (Montgomery, 1999; Church, 2002). Headwater streams are typically
erosion zones in which sediment from the base of adjoining hillslopes moves directly into stream
channels and is transported downstream. As stream channels increase in size and decrease in slope, a
mixture of erosion and deposition processes usually is at work. At some point in the lower portions of
river networks, sediment deposition becomes the dominant process and floodplains form. Floodplains
are level areas bordering stream or river channels that are formed by sediment deposition from those
channels under present climatic conditions (Figure 2-3). These natural geomorphic features are
inundated during moderate to high water events (Leopold, 1994; Osterkamp, 2008). Floodplain and
associated river channel forms (e.g., meandering, braided, anastomosing) are determined by interacting
fluvial factors, including sediment size and supply, channel gradient, and streamflow (Church, 2002,
2006). Terraces are historical floodplains, formed under different climatic conditions, that are no longer
connected to the river or stream channel that formed them (Figure 2-3).
Both riparian areas and floodplains are important components of river systems (Figure 2-3). Riparian
areas are transition zones between terrestrial and aquatic ecosystems that are distinguished by
gradients in biophysical conditions, ecological processes, and biota. They are areas through which
surface and subsurface hydrology connect water bodies with their adjoining uplands, and they include
those portions of terrestrial ecosystems that significantly influence exchanges of energy and matter with
aquatic ecosystems (National Research Council, 2002). Riparian areas often have high biodiversity
(Naiman et al., 2005). They occur near lakes and estuarine-marine shorelines and along river networks,
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Figure 2-2. Elements of a river system. These elements include: the drainage basin (light green area),
river network (rivers and streams), and other water bodies (riparian/floodplain wetlands, lakes, and
wetlands in non-floodplain settings). Note that the non-floodplain wetland that lacks a stream outlet
also would be considered "geographically isolated" sensu Tiner (2003b).
Drainage boundary
Floodplain boundary
River
Perennial stream
Intermittent stream
——•- Ephemeral stream
Wetland with
surface outlet
Riparian wetland
Geographically
isolated wetland
where their width can vary from narrow bands along headwater streams (Figure 2-3A) to broad zones
that encompass the floodplains of large rivers (Figure 2-3B).
Floodplains are also considered riparian areas, but not all riparian areas have floodplains. All rivers and
streams within river networks have riparian areas, but small streams in constrained valleys are less
likely to have floodplains than larger streams and rivers in unconstrained valleys (Figures 2-2 and 2-3).
The Federal Emergency Management Agency defines the area that will be inundated by the flood event
having a 1% chance of being equaled or exceeded in any given year as the "Special Flood Hazard Area,"
also referred to as the "100-year floodplain" (https://www.fema.gov/floodplain-management/flood-
zones). The 100-year floodplain can but need not coincide with the geomorphic floodplain. Like riparian
areas, wetlands are transitional areas between terrestrial and aquatic ecosystems. According to
Cowardin etal. (1979), an area is classified as a wetland if it has one or more of the following three
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Figure 2-3. Hypothetical cross-sections of (A) a headwater stream and (B) a large river within a river
network. The headwater stream in (A) is a constrained reach with a narrow riparian area and no
floodplain; the river in (B) has both a riparian area and a floodplain with the same spatial extent.
Examples of other common natural floodplain features are shown in (B). The lateral extent of riparian
areas varies depending on the criteria used for delineation.
A. Headwater Stream with Riparian Area and Minimal
or No Floodplain
B. River with Riparian Area and Floodplain
I Floodplain and riparian area
Terrace
attributes: (1) the area supports predominantly hydrophytes (i.e., water-loving plants) at least
periodically; (2) the land has substrate that is predominantly undrained hydric soil; or (3) the land has
nonsoil substrate that is saturated with water or covered by shallow water at some time during the
growing season of each year. Note that the Cowardin etal. (1979) definition requires only one of these
characteristics, in contrast to the federal regulatory definition, which requires all three (33 Code of
Federal Regulations 328.3(b); see also USAGE, 1987). Thus, as used in this report, a wetland need not
meet the federal regulatory definition. Wetlands include areas such as swamps, bogs, fens, marshes,
ponds, and pools (Mitsch et al., 2009).
Many classification systems have been developed for wetlands (Mitsch and Gosselink, 2007). These
classifications can focus on vegetation, hydrology, hydrogeomorphic characteristics, or other factors
(Cowardin etal., 1979; Brinson, 1993; Tiner, 2003a; Comer etal., 2005). Because this report focuses on
downstream connectivity (Section 1.3), we consider two landscape settings in which wetlands occur
based on directionality of hydrologic flows. Directionality of flow also is included as a component of
hydrodynamic setting in the hydrogeomorphic approach (Brinson, 1993; Smith etal., 1995) and as an
element of water flowpath in an enhancement of National Wetlands Inventory data (Tiner, 2011). This
emphasis on directionality of flow is necessary because hydrologic connectivity plays a dominant role in
determining the types of effects wetlands have on downstream waters (Section 2.3.2).
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A non-floodplain wetland setting is a landscape setting where the potential exists for unidirectional,
lateral hydrologic flows from wetlands to the river network through surface water or ground water.
Such a setting would include upgradient areas such as hillslopes or upland areas outside of the
floodplain. Any wetland setting where water could only flow from the wetland toward a river network
would be considered a non-floodplain setting, regardless of the magnitude and duration of flows and of
travel times. In this document, we therefore refer to wetlands that occur in these settings as non-
floodplain wetlands.
A riparian or floodplain wetland setting is a landscape setting (e.g., floodplains, most riparian areas,
lake and estuarine fringes) that is subject to bidirectional, lateral hydrologic flows. Wetlands in
riparian/floodplain settings can have some of the same types of hydrologic connections as those in non-
floodplain settings. In addition, wetlands in these settings also have bidirectional flows. For example,
wetlands within a riparian area are connected to the river network through lateral movement of water
between the channel and riparian area (e.g., through overbank flooding, hyporheic flow). Given our
interest in addressing the effects of wetlands on downstream waters (Section 1.1), we have focused in
particular on the subset of these wetlands that occur in riparian areas with and without floodplains
(collectively referred to hereafter as riparian/floodplain wetlands); we generally do not address
wetlands at lake and estuarine fringes. Riparian wetlands are portions of riparian areas that meet the
Cowardin etal. (1979) three-attribute wetland criteria (i.e., having wetland hydrology, hydrophytic
vegetation, or hydric soils); floodplain wetlands are portions of the floodplain that meet these same
criteria.
Our use of landscape setting to define riparian/floodplain wetlands and non-floodplain wetlands is
similar to the use of landscape position by Tiner (2011) to supplement the Cowardin etal. (1979)
classification. Our use of riparian/floodplain wetland setting is generally consistent with Tiner's
estuarine, lotic, and lentic landscape positions, whereas our non-floodplain setting is similar to his
terrene category (Tiner, 2011). One important difference is that Tiner (2011) would consider a wetland
to be terrene if it were located along a river but not subject to frequent overflow. Given that even
infrequent flooding can have profound effects on wetland development and function, we would consider
such a wetland to be in a riparian/floodplain setting.
The terms "riparian/floodplain" and "non-floodplain" are meant to describe the landscape setting in
which wetlands occur and do not refer to wetland type or class. Many wetland types occur in both
settings. For example, a palustrine emergent wetland (Cowardin et al., 1979) could be located outside a
floodplain, or it could be located within a floodplain and subject to bidirectional flows. A wetland that is
classified as depressional in the hydrogeomorphic approach could have any combination of inlets and
outlets or none at all (Smith et al., 1995). The setting for such a wetland would be riparian/floodplain if
it had both an input and output channel because water from the stream flows into and affects the
wetland. A depressional wetland with a surface outlet and no inlet or with no outlets and inlets,
however, would be considered non-floodplain because water could flow downgradient only from the
wetland to the river network, and not from a stream to the wetland. Similarly, a riverine wetland (Smith
etal., 1995) that is the origin for a stream would be considered non-floodplain if it had no input channel,
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even though it occurs in a riparian area. In most cases, however, riverine wetlands would be considered
riparian/floodplain. Thus, directionality of hydrologic flow is a function of landscape setting and cannot
necessarily be determined from wetland class.
A major consequence of the two different landscape settings is that waterborne materials can be
transported only from the wetland to the river network for a non-floodplain wetland, whereas
waterborne materials can be transported from the wetland to the river network and from the river
network to the wetland for a riparian/floodplain wetland. In the latter case, there is a mutual,
interacting effect on the structure and function of both the wetland and river network. In contrast, a
non-floodplain wetland can affect a river through the transport of waterborne material, but the opposite
is not true. Note that we limit our use of riparian/floodplain and non-floodplain landscape settings to
describe the direction of hydrologic flow; the terms cannot be used to describe directionality of
geochemical or biological flows. For example, mobile organisms can move from a stream to a non-
floodplain wetland (e.g., Subalusky et al., 2009a; Subalusky et al., 2009b). In Alaska, transport of live
salmon or their carcasses from streams to riparian areas by brown bears (Ursus arctos) account for
more than 20% of riparian nitrogen budgets (Helfield and Naiman, 2006). Although this example is in a
riparian/floodplain setting, it shows how geochemical fluxes can be decoupled from hydrologic flows.
Both non-floodplain and riparian/floodplain wetlands can include geographically isolated wetlands,
or wetlands completely surrounded by uplands (Tiner, 2003b). These wetlands have no apparent
surface-water outlets, but can hydrologically connect to downstream waters through spillage or ground
water. We define an upland as any area not meeting the Cowardin et al. (1979) three-attribute wetland
criteria, meaning that uplands can occur in both terrestrial and riparian areas. Thus, a wetland that is
located on a floodplain but is surrounded by upland would be considered a geographically isolated,
riparian/floodplain wetland that is subject to periodic inundation from the river network. Although the
term "geographically isolated" could be misconstrued as implying functional isolation, the term has been
defined in the peer-reviewed literature to refer specifically to wetlands surrounded by uplands.
Furthermore, the literature explicitly notes that geographic isolation does not imply functional isolation
(Leibowitz, 2003; Tiner, 2003b). Discussion of geographically isolated wetlands is essential because
hydrologic connectivity (an element of connectivity, which is the focus of this document) is generally
difficult to characterize for these wetlands. The difficulty arises because hydrologic monitoring or
additional information and analyses would be necessary to determine whether surface or subsurface
hydrologic connections occur for such wetlands.
2.2.2 River System Hydrology
River system hydrology is controlled by hierarchical factors that result in a broad continuum of
belowground and aboveground hydrologic flowpaths connecting river basins and river networks
(Winter, 2001; Wolocketal., 2004; Devito etal., 2005; Poole etal., 2006; Wagener et al., 2007; Poole,
2010; Bencala etal., 2011; Jencso and McGlynn, 2011). At the broadest scale, regional climate interacts
with river-basin terrain and geology to shape inherent hydrologic infrastructure that bounds the nature
of basin hydrologic flowpaths. Different climate-basin combinations form identifiable hydrologic
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landscape units with distinct hydrologic characteristics (Winter, 2001; Wigington et al., 2013). Buttle
(2006) posited three first-order controls of watershed streamflow generated under specific
hydroclimatic conditions: (1) the ability of different landscape elements to generate runoff by surface or
subsurface lateral flow of water; (2) the degree of hydrologic linkage among landscapes by which
surface and subsurface runoff can reach river networks; and (3) the capacity of the river network itself
to convey runoff downstream to the river-basin outlet. River and stream waters are influenced by not
only basin-scale or larger ground-water systems, but also local-scale, vertical and lateral hydrologic
exchanges between water in channels and sediments beneath and contiguous with river network
channels (Ward, 1989; Woessner, 2000; Malard etal., 2002; Bencala, 2011). The magnitude and
importance of river-system hydrologic flowpaths at all spatial scales can radically change over time at
hourly to yearly temporal scales (Junk et al., 1989; Ward, 1989; Malard et al., 1999; Poole et al., 2006).
Because interactions between ground waters and surface waters are essential processes in rivers,
knowledge of basic ground-water hydrology is necessary to understand the interactions between
surface and subsurface water and their relationship to connectivity within river systems. Subsurface
water occurs in two principal zones: the unsaturated zone and the saturated zone (Figure 2-4; Winter et
al., 1998). In the unsaturated zone, the spaces between soil, gravel, and other particles contain both air
and water. In the saturated zone, these spaces are completely filled with water. Ground water refers to
any water that occurs and flows (saturated ground-water flow) in the saturated zone beneath a
watershed surface (Winter et al., 1998). Rapid flow (interflow) of water can occur through large pore
spaces in the unsaturated zone (Beven and Germann, 1982).
Traditionally, geologic formations in which ground water occurs are divided into two major categories:
(1) aquifers, which are saturated geologic units capable of transmitting significant amounts of water
under ordinary hydraulic gradients; and (2) aquicludes, which are saturated geologic units that are not
capable of transmitting significant quantities of water (aquicludes are also referred to as confining
layers or confining units; Freeze and Cherry, 1979). Water flow in an aquifer can take various forms:
Water can flow in small voids and pores between the aquifer strata (porous media aquifers), in large
voids (karst), or in fractures and cracks within the aquifer formation (fractured flow aquifers). Flow
differs in its characteristics between the various aquifer types mentioned, yet follows the same basic
rule, by which flow occurs from regions of high hydraulic pressure to regions of lower hydraulic
pressure, down the pressure gradient (Jones and Mulholland, 2000).
There are two main types of aquifers (Freeze and Cherry, 1979). Unconfined aquifers are underlain by
a confining unit but remain open to the atmosphere at their top and exchange gases with the
environment. The upper saturated horizon in unconfined aquifers is known as the water table (Figure
2-4). Complex geologic conditions can lead to more complex distributions of saturated and unsaturated
zones. Discontinuous saturated lenses creating perched water tables can occur where low
permeability layers (e.g., clay) are present in the midst of highly permeable materials such as sand
(Freeze, 1971). Confined aquifers are bounded by an underlying confining unit and an overlying
confining unit and typically lack a direct connection with current surface and atmospheric conditions
(Figure 2-5). Water in confined aquifers is often pressurized, and, consequently, water levels in wells
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Figure 2-4. Water below the land surface occurs in either the unsaturated or the saturated zone.
The upper surface of the saturated zone is the water table. Ground water and ground-water flow occur
in the saturated zone. If a surface-water body is connected to the ground-water system, the water
table intersects the water body at or near the surface of its shoreline. Modified from Winter et al.
(1998).
Unsaturated zone
-T Water table
Surf ace water
turated zone (around wat
penetrating confined aquifers occur at elevations above the upper confining unit The surface of the
water levels in wells penetrating a confined aquifer is called the potentiometric surface. Confined
aquifers typically occur deeper below the land surface than unconfined aquifers and generally have less
frequent influence on surface waters than unconfined aquifers.
Traditionally, aquifers were identified based solely on their ability to support wells for water
production, but in recent years hydrologists studying ground water-surface water interactions have
recognized the need for a broader definition that recognizes the importance of low-flow geologic
formations to aquatic ecosystems. Payne and Woessner (2010) highlighted the importance of aquifers
with varying flow rates on streams and proposed a classification of aquifer flow systems that ranged
from high flow to low flow, with low flow aquifers having limited ground-water discharge potential
except for small streams and wetlands. Winter et al. (1998) simply defined aquifers as the permeable
materials (e.g., soil, rock) through which ground water flows. In this report, we have adopted the Winter
et al. (1998) aquifer definition. Unless otherwise noted, our discussion of ground water and aquifers is
limited to unconfined systems.
Ground-water recharge areas occur where water from land surfaces or surface-water bodies infiltrates
and moves into saturated zones. Discharge areas occur where water flows from saturated zones into a
river network, other water bodies, or onto land surfaces. A gaining stream (or wetland; also referred
to as a discharge wetland) within a river network receives inflow of ground water. In this situation, the
water table elevation near the stream (or wetland) must be higher than the elevation of the stream
water surface. In a losing stream (or wetland; i.e., recharge wetland), water flows from the stream
(wetland) to ground water. In this situation, the water table elevation near the stream or wetland is
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Figure 2-5. Cross-section showing major hydrologic f lowpaths in a regional-scale stream-watershed
system. USF = unsaturated flow, GW = ground-water flowpath (saturated flow); GW1, GW2, and
GW3 = ground-water flowpaths of varying depth and length. GW1 represents local ground water and
GW3 represents regional ground water. GWCF = ground-water flowpath in confined aquifer.
Precipitation
GW3
Confining layer/formation
lower than the stream or wetland water surface. Conditions that determine whether streams and
wetlands are gaining or losing can change over short periods of time and over short distances within
river networks and river basins (Winter et al., 1998; Harrington et al., 2002; Wilson and Guan, 2004;
Goes and Pool, 2005; Scanlon etal., 2006; Vivoni etal., 2006; Larned etal., 2008). Overall, however, the
volume and sustainability of streamflow within river networks typically depend on contributions from
ground water (Winter, 2007), especially in areas with shallow ground-water tables and pervious
subsurfaces (de Vries, 1995; Kish et al., 2010).
Ground-water flow systems within river basins can be complex, of varying sizes and depths, and overlie
one another (Toth, 1963; Winter etal., 1998; Haitjema and Mitchell-Bruker, 2005). Although in reality
there is a continuum of flowpath lengths that occur within river basins (Bencala et al., 2011), they are
commonly grouped into three categories (Figure 2-5). In local ground-water flow systems (also
referred to as shallow ground-water flow systems), ground water flows from the highest elevations of
water tables (water table highs) to nearby lowlands or surface waters (Winter and LaBaugh, 2003).
Local ground-water flow is the most dynamic of ground-water flow systems, having the greatest
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interchange with surface waters. If the depth-to-width ratios of aquifers are sufficiently large, regional
flow systems (deepest ground-water flowpaths) also might be present. Regional ground water (also
referred to as deep ground water) originates from precipitation in distant upland recharge areas and
moves long distances, through deep regional-scale aquifers, to river networks (Figure 2-5). The contact
times between ground water and subsurface materials are longer for these deep and long flow systems
than for local systems. Eventually, deep regional flow systems also discharge to surface waters in the
lower portions of river networks where they influence surface-water conditions. An intermediate
ground-water flow system is one in which ground water flows from a water table high to a lowland that
is not immediately adjoining the water table high. Intermediate ground-water flow systems are
representative of the wide range of flowpath lengths and depths that occur between local and regional
ground-water systems.
Other hydrologic flowpaths are also significant in determining the characteristics of river systems. The
most obvious is the downstream water movement within stream or river channels, or open-channel
flow. River water in stream and river channels can reach riparian areas and floodplains via overbank
flow (Figure 2-6A), which occurs when floodwaters flow over stream and river channels (Mertes, 1997).
Overland flow is the portion of streamflow derived from net precipitation that flows over the land
surface to the nearest stream channel with (Figure 2-6A; Hewlett, 1982). Overland flow can be
generated by several mechanisms. Infiltration-excess overland flow occurs when the rainfall rates
exceed the infiltration rates of land surfaces (Horton, 1945). Saturation-excess overland flow occurs
when precipitation inputs cause water tables to rise to land surfaces so that precipitation inputs to the
land surfaces cannot infiltrate and flow overland (Dunne and Black, 1970). Return flow occurs when
water infiltrates, percolates through the unsaturated zones, enters saturated zones, and then returns to
and flows over watershed surfaces, commonly at hillslope-floodplain transitions (Dunne and Black,
1970).
Alluvium (Figure 2-3B) comprises deposits of clay, silt, sand, gravel, or other particulate materials that
running water has deposited in a streambed, on a floodplain, on a delta, or in a fan at the base of a
mountain. These deposits occur near active river systems but also can be found in buried river valleys—
the remnants of relict river systems (Lloyd and Lyke, 1995). In this report, we are concerned primarily
with alluvium deposited along active river networks. Commonly, alluvium is highly permeable, creating
an environment conducive to ground-water flow. Alluvial ground water (typically a mixture of river
water and local, intermediate, and regional ground water) moves through the alluvium. Together, the
alluvium and alluvial ground water comprise alluvial aquifers. Alluvial aquifers are closely associated
with floodplains and have high levels of hyporheic exchange (Stanford and Ward, 1993; Amoros and
Bornette, 2002; Poole et al., 2006). Hyporheic exchange occurs when water moves from stream or river
It occurs during both high- and low-flow periods, and typically has relatively horizontal flowpaths at
scales of meters to tens of meters (Bencala, 2005) and vertical flowpaths with depths ranging from
centimeters to tens of meters (Stanford and Ward, 1988; Woessner, 2000 and references therein).
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Figure 2-6. Hyporheic zone flows. (A) Common hydrologic flowpaths by which water flows between
watersheds and river networks. (B) and (C) The three-dimensional process of hyporheic flow, or the
movement of water from a river or stream to nearby alluvium and then back to the river or stream.
Modified from Winter et al. (1998).
A. Common River-Floodplain Hydrologic Flowpaths
Riparian
1 - overland flow 4 - regional ground water
2-overbankflow 5-hyporheic flow
3 - local ground water 6 - wetland overflow
B. Hyporheic Zone Cross-Section
Water
table
ground-water flow
hyporheic flow
C. Hyporheic Zone Longitudinal Profile
ground-water flow
hyporheic flow
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Riparian areas and floodplains can have a diverse array of hydrologic inputs and outputs, which, in turn
influence riparian/floodplain wetlands. Riparian areas and floodplains receive water from precipitation;
overland flow from upland areas; local, intermediate, regional ground water; and hyporheic flows
(Figure 2-6A; National Research Council, 2002; Richardson etal., 2005; Vidon etal., 2010). Water
flowing over the land surface in many situations can infiltrate soils in riparian areas. If low permeability
subsoils or impervious clay layers are present, water contact with the plant root zone is increased and
materials in the water are subject to ecological functions such as denitrification before it reaches the
stream channel (Section 4.3.2; National Research Council, 2002; Naiman et al., 2005; Vidon et al., 2010).
The relative importance of the continuum of hydrologic flowpaths among river systems varies, creating
streams and rivers with different flow duration (or hydrologic permanence) classes (Figures 2-2 and
2-7). Perennial streams or stream reaches (Figure 2-7A) typically flow year-round. They are
maintained by local or regional ground-water discharge or streamflow from higher in the stream or
river network. Intermittent streams or stream reaches (Figure 2-7B) flow continuously at certain
times of the year (e.g., during certain seasons such as spring snowmelt); drying occurs when the water
table falls below the channel bed elevation. Ephemeral streams or stream reaches (Figure 2-7C) flow
briefly (typically hours to days) during and immediately following precipitation; these channels are
above the water table at all times. Streams in these flow duration classes often transition longitudinally,
from ephemeral to intermittent to perennial, as drainage area increases and elevation decreases along
river networks. Many headwater streams, however, originate from permanent springs and flow into
intermittent downstream reaches. At low flows, intermittent streams can contain dry segments
alternating with flowing segments. Transitions between flow duration classes can coincide with
confluences or with geomorphic discontinuities within the network (May and Lee, 2004; Hunter et al.,
2005). Variation of streamflow within river systems occurs in response to hydrologic events resulting
from rainfall or snowmelt. Stormflow is streamflow that occurs in direct response to rainfall or
snowmelt (Figure 2-8A), which might stem from multiple ground-water and surface-water sources
(Dunne and Leopold, 1978). Baseflow is streamflow originating from ground-water discharge or
seepage (locally or from higher in the river network), which sustains water flow through the channel
between hydrologic events (Figure 2-8A). Perennial streams have baseflow year-round; intermittent
streams have baseflow seasonally; ephemeral streams have no baseflow. All three stream types convey
stormflow. Thus, perennial streams are more common in areas receiving high precipitation, whereas
intermittent and ephemeral streams are more common in the more arid portions of the United States
(Figure 2-9; NHD, 2008). The distribution of headwater streams (perennial, intermittent, or ephemeral)
as a proportion of total stream length is similar across geographic regions and climates (Figure 2-9C).
Similar to streams, the occurrence and persistence of riparian/floodplain wetland and non-floodplain
wetland hydrologic connections with river networks, via surface water (both channelized and
nonchannelized) or ground water, can be continuous, seasonal, or ephemeral, depending on the overall
hydrologic conditions in the watershed. For example, a non-floodplain wetland might have a direct
ground-water connection with a river network during wet conditions but an indirect regional ground-
water connection (via ground-water recharge) under dry conditions. Geographically isolated wetlands
can be hydrologically connected to the river network via nonchannelized surface flow (e.g., swales or
overland flow) or ground water.
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Figure 2-7. Hypothetical hydrographs illustrating maximum duration of flow (Dmax.q) for (A)
perennial, (B) intermittent, and (C) ephemeral streams. Source: Reprinted from Non-navigable
streams and adjacent wetlands: Addressing science needs following the Supreme Court's Rapanos
decision, (2008) by Leibowitz et al. with permission of Ecological Society of America.
TO
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Figure 2-8. (A) Hypothetical hydrograph showing stormflow and baseflow responses to a rainfall
event. (B) Expansion and contraction of flowing water in a stream network following a rainfall
event. Panel B Source: Reprinted from Subsurface stormflows in the highly permeable forested
watersheds of southwestern British Columbia, (1988) by Cheng et al. with permission of Elsevier.
Time (days)
.5
B.
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0.4-
0.3-
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Figure 2-9. Characteristics of U.S. streams by watershed, in terms of percent of total stream length
as (A) perennial, (B) intermittent, and (C) headwater streams. Data from the National Hydrography
Dataset (NHD) Reach Address Database (RAD) v2.0 at 1:100,000 scale using 8-digit HUG (Hydrologic
Unit Code) watersheds. Here, "intermittent" includes streams having intermittent or ephemeral flow.
Note that NHD data generally do not capture streams <1.6 km (1 mile) in length, and ranges of color
categories are not consistent across maps.
Perennial Stream Length
as a Percentage of
Total Stream Length
% Perennial Stream
1-15
16-35
36-58
59-81
82-100
State Boundary
B.
Intermittent Stream Length
as a Percentage of
Total Stream Length
% Intermittent Stream
I 10
I I 1-16
I I 17-38
I I 39-60
I 161-81
BB 82-100
I I State Boundary
Headwater Stream Length
as a Percentage of
Total Stream Length
% Headwater Stream
10
I 1-19
I 20-36
I 37-48
BB 49-57
BB 58-100
I I State Boundary
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2.2.3 River Network Expansion and Contraction
The portions of river networks with flowing water expand and contract longitudinally (in an upstream-
downstream direction) and laterally (in a stream channel-floodplain direction) in response to seasonal
environmental conditions and precipitation events (Hewlett and Hibbert, 1967; Gregory and Walling,
1968; Dunne and Black, 1970; Day, 1978; Junketal., 1989; Hunter etal., 2005; Wigington etal., 2005;
Rains et al., 2006; Rains et al., 2008). The longitudinal expansion of channels with flowing water in
response to major precipitation events represents a transient increase in the extent of headwater
streams. Figure 2-10 shows the expansion of the flowing portion of two stream networks in western
Oregon between dry, summer and wet, winter seasons. Intermittent and perennial streams flow during
wet seasons, whereas ephemeral streams flow only in response to rainfall or snowmelt. During dry
periods, flowing portions of river networks are limited to perennial streams; these perennial portions of
the river network can be discontinuous (Stanley et al., 1997; Hunter et al., 2005; Larned et al., 2010) or
interspersed with intermittently flowing stream reaches.
The dominant sources of water to a stream can shift during river network expansion and contraction
(Malard etal., 1999; McGlynn and McDonnell, 2003; McGlynn etal., 2004; Malard etal., 2006). Rainfall
and snowmelt cause a river network to expand in two ways. First, local aquifers expand and water
moves into dry channels, which increases the total length of the wet channel (Winter et al., 1998); the
resulting intermittent streams will contain water during the entire wet season. Second, stoonflow can
cause water to enter ephemeral and intermittent streams (Figure 2-8). The larger the rainfall or
snowmelt event, the greater the number of ephemeral streams and total length of flowing channels that
occur within the river network. Ephemeral flows cease within days after rainfall or snowmelt ends
(Figure 2-8B), causing the length of wet channels to decrease and river networks to contract. The
flowing portion of river networks further shrinks as the spatial extent of aquifers with ground water in
contact with streams contract and intermittent streams dry. In many river systems across the United
States, stoonflow comprises a major portion of annual streamflow (Hewlett et al., 1977; Miller et al.,
1988; Turton et al., 1992; Goodrich et al., 1997; Vivoni et al., 2006). In these systems, intermittent and
ephemeral streams are major sources of river water (Section B.5). When rainfall or snowmelt induces
stormflow in headwater streams or other portions of the river network, water flows downgradient
through the network to its lower reaches. As water moves downstream through a river network, the
hydrograph for a typical event broadens with a lower peak (Figure 2-11). This broadening of the
hydrograph shape (Figure 2-11A) results from transient storage of water in river network channels and
nearby alluvial aquifers (Fernald etal., 2001).
Floodplains and riparian areas can be locations with significant ground-water recharge and discharge
(National Research Council, 2002; Naiman etal., 2005). During very large hydrologic events, aggregate
flows from headwaters and other tributary streams can result in overbank flooding in river reaches with
floodplains; this occurrence represents lateral expansion (Figure 2-12) of the river network (Mertes,
1997). Water from overbank flows can recharge alluvial aquifers, supply water to floodplain wetlands,
surficially connect floodplain wetlands to rivers, and shape the geomorphic features of the floodplain
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Figure 2-10. Extent and connectivity of streams with flowing water, wetlands, and other water
bodies in (A) Spring Valley Creek, OR and (B) Spoon Creek, OR during dry summer (left) and wet
winter (right) conditions. Source: Reprinted from Stream network expansion: A riparian water quality
factor, (2005) by Wigington et al. with permission of John Wiley & Sons.
A. Spring Valley Creek
Summer
Winter
0 1 2
Kilometers
B. Spoon Creek
Summer
Winter
Kilometers
Perennial stream
Intermittent stream
Ephemeral stream
Surface field drainage
Swale
Road ditch
Ephemeral water bodies (depressions)
(Wolman and Miller, 1960; Hammersmark et al., 2008). Depending on the nature of the hydraulic
gradients, ground water within floodplain alluvium can move both parallel and perpendicularly to
streams or rivers (National Research Council, 2002) and enter river networks at various discharge
points. Bidirectional exchanges of water between ground water and river networks, including hyporheic
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Figure 2-11. Stormflow moves downstream through the river network and interacts with lower
stream reaches, f loodplains, and alluvial aquifers. (A) Hydrographs for three nested rivers in the
Potomac River watershed (drainage area Potomac > Shenandoah > South). (B) Hydrographs for the
same three rivers with streamflow normalized by drainage area. Source: Reprinted from Elements of
physical hydrology, (1998) by Hornberger et al., with permission of Johns Hopkins University Press.
5000
A.
4000-
3000-
2000-
r
1000-
0
Potomac
4
I
6
8
10
12
2468
Time (days from April 13,1987)
10
12
flow, can occur under a wide range of streamflows, from flood flows to low flows (National Research
Council, 2002; Naiman etal., 2005; Vivoni etal., 2006).
The hydrologic connections with river networks fundamentally differ for riparian/floodplain wetlands
and non-floodplain wetlands. Riparian/floodplain wetlands can have bidirectional, lateral hydrologic
connections to the river network, either through overbank flooding (i.e., lateral expansion of the
network) or hyporheic flow, in addition to unidirectional flows from upland and ground-water sources
(Figure 2-6A). In contrast, hydrologic connections between non-floodplain wetlands and river networks
originate via surface-water spillage or ground-water flow when water inputs exceed evapotranspiration
and available storage. Although wetlands that serve as origins for streams are riparian, we group them
with non-floodplain wetlands because they also have unidirectional flow through their outlet streams. In
both cases, the degree of hydrologic connectivity between riparian/floodplain and non-floodplain
wetlands and the river network varies with lateral expansion and subsequent contraction.
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Figure 2-12. Landsat 5 satellite images of the Mississippi River along the borders of Tennessee,
Kentucky, Missouri, and Arkansas on (A) May 12, 2006 and (B) May 10, 2011. Images courtesy of
U.S. Geological Survey/National Aeronautics Space Administration.
One factor affecting the lateral distance that overbank flow spreads is preexisting moisture conditions
on the floodplain (Mertes, 1997; Naiman et al., 2005). River overbank flow that enters a dry floodplain
will spread and then infiltrate the soil (Naiman et al., 2005). If inflows from streams, rainfall, or ground
water have water tables elevated to the floodplain surface, water entering the riparian area from
overbank flow cannot infiltrate soils. The result is standing water on the floodplain and subsequent
movement of water to lower elevations of the floodplain. This water can alter the geomorphology of the
floodplain (Hupp and Osterkamp, 1996), be biogeochemically transformed (Section 4.3.2; Naiman et al.,
2005), be lost by evaporation, or be transpired by vegetation (Meyboom, 1964). As the river and
floodplain water table elevations decrease, surface water on the floodplain can flow back into the river,
infiltrate floodplain soils, or evapotranspire.
Many studies have documented the fact that riparian/floodplain wetlands can attenuate flood pulses of
streams and rivers by storing excess water from streams and rivers. Bullock and Acreman (2003)
reviewed wetland studies and reported that wetlands reduced or delayed floods in 23 of 28 studies. For
example, Walton et al. (1996) found that peak discharges between upstream and downstream gaging
stations on the Cache River in Arkansas were reduced 10-20% primarily due to floodplain water
storage. Locations within floodplains and riparian areas with higher elevations likely provide flood
storage less frequently than lower elevation areas.
The interactions of high flows with floodplains and associated alluvial aquifers of river networks are
important determinants of hydrologic and biogeochemical conditions of rivers (Ward, 1989; Stanford
and Ward, 1993; Boulton etal., 1998; Burkart et al., 1999; Malard etal., 1999; Amoros and Bornette,
2002; Malard etal., 2006; Poole, 2010). Bencala (1993; 2011) noted that streams and rivers are not
pipes: They interact with the alluvium and geologic materials adjoining and under channels. In streams
or river reaches constrained by topography, significant floodplain and near-channel alluvial aquifer
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interactions are limited (Figure 2-3A). In reaches with floodplains, however, stormflow commonly
supplies water to alluvial aquifers during high-flow periods through the process of bank storage
(Figure 2-13; Whiting and Pomeranets, 1997; Winter et al., 1998; Chen and Chen, 2003). As streamflow
decreases after hydrologic events, the water stored in these alluvial aquifers can serve as another source
of baseflowin rivers (Figure 2-13C).
In summary, the extent of wetted channels is dynamic because interactions between surface water in the
channel and alluvial ground water, via hyporheic exchange, determine open-channel flow. The flowing
portion of river networks expands and contracts in two primary dimensions: (1) longitudinally, as
intermittent and ephemeral streams wet up and dry; and (2) laterally, as floodplains and associated
alluvial aquifers gain (via overbank flooding, bank storage, and hyporheic exchange) and lose (via
draining of alluvial aquifers and evapotranspiration) water. Vertical ground-water exchanges between
streams and rivers and underlying alluvium are also key connections, and variations in these vertical
exchanges contribute to the expansion and contraction of the portions of river networks with open-
channel flow. Numerous studies have documented expansion and contraction of river systems (e.g.,
Gregory and Walling, 1968); the temporal and spatial pattern of this expansion and contraction varies in
response to many factors, including interannual and long-term dry cycles, climatic conditions, and
watershed characteristics (Cayan and Peterson, 1989; Fleming et al., 2007).
2.3 Influence of Streams and Wetlands on Downstream
Waters
The previous section provided background on river system hydrology. In this section, we present a
general overview of how streams and wetlands affect downstream waters, focusing on functions within
streams and wetlands and their connectivity to rivers.
The structure and function of rivers are highly dependent on the constituent materials stored in and
transported through them. Most of these materials, broadly defined here as any physical, chemical, or
biological entity, including water, heat energy, sediment, wood, organic matter, nutrients, chemical
contaminants, and organisms, originate outside of the river: They originate from either the upstream
river network or other components of the river system, and then are transported to the river by water
movement or other mechanisms. Thus, the fundamental way in which streams and wetlands affect river
structure and function is by altering fluxes of materials to the river. This alteration of material fluxes
depends on two key factors: (1) functions within streams and wetlands that affect material fluxes, and
(2) connectivity (or isolation) between streams and wetlands and rivers that allows (or prevents)
transport of materials between the systems.
2.3.1 Effects of Streams and Wetlands on Material Fluxes
Streams and wetlands affect the amounts and types of materials that are or are not delivered to
downstream waters, ultimately contributing to the structure and function of those waters. Leibowitz et
al. (2008) identify three functions, or general mechanisms of action, by which streams and wetlands
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Figure 2-13. The direction and magnitude of interactions between surface water and ground water
can dramatically change during large hydrologic events, including floods. (A) In a hypothetical
stream-floodplain cross-section, ground water flows from the alluvial aquifer to the stream before a
major hydrologic event. (B) During the bank-full hydrologic event, surface water moves from the
stream and becomes ground water in the alluvial aquifer. (C) After recession of the event, ground
water that was stored in the alluvial aquifer during the hydrologic event flows back to the stream. This
process is called bank storage, which can sustain baseflow in streams and rivers after the hydrologic
event has ended. Modified from Winter et al. (1998).
HUnsaturated
zone
• Saturated
zone
^ Elevation of
water in stream
^ Groundwater
flow
A.
influence material fluxes into downstream waters: source, sink, and refuge. We have expanded on this
framework to include two additional functions: lag and transformation. These five functions
(summarized in Table 2-1) provide a framework for understanding how physical, chemical, and
biological connections between streams and wetlands and downstream waters influence river systems.
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These five functions (Table 2-1) are neither static nor mutually exclusive, and often the distinctions
between them are not sharp. A stream or wetland can provide different functions at the same time.
These functions can vary with the material considered (e.g., acting as a source of organic matter and a
sink for nitrogen) and can change over time (e.g., acting as a water sink when evapotranspiration is high
and a water source when evapotranspiration is low). The magnitude of a given function also is likely to
vary temporally: For example, streams generally are greater sources of organic matter and
contaminants during high flows.
Leibowitz et al. (2008) explicitly focused on functions that benefit downstream waters, but these
functions also can have negative effects—for example, when streams and wetlands serve as sources of
chemical contamination (Sections 3.4.4, 4.3.3.5, 4.3.3.6; Table 2-1). In fact, benefits need not be linear
with respect to concentration; a beneficial material could be harmful at higher concentrations due to
nonlinear and threshold effects. For example, nitrogen can be beneficial at lower concentrations but can
reduce water quality at higher concentrations. Although here we focus primarily on the effects of
streams and wetlands on downstream waters, these same functions can describe effects of downstream
waters on streams and wetlands (e.g., downstream rivers can serve as sources of colonists for upstream
tributaries).
Because many of these functions depend on import of materials and energy into streams and wetlands,
distinguishing between actual function and potential function is instructive. For example, a wetland with
appropriate conditions (e.g., a reducing environment and denitrifying bacteria) is a potential sink for
nitrogen (Sections 4.3.3.2 and 4.4.3.2): If nitrogen is imported into the wetland, the wetland can remove
it by denitrification. The wetland will not serve this function, however, if nitrogen is not imported. Thus,
even if a stream or wetland does not currently serve a function, it has the potential to provide that
function under appropriate conditions (e.g., when material imports or environmental conditions
change). Although potential functions do not actively affect downstream waters, they can be
instrumental in protecting those waters from future impacts. Ignoring potential function also can lead to
the paradox that degraded streams and wetlands (e.g., those receiving nonpoint-source nitrogen inputs)
receive more protection than less degraded systems (Leibowitz etal., 2008).
Three factors influence the effect that material and energy fluxes from streams and wetlands have on
downstream waters: (1) proportion of the material originating from (or reduced by) streams and
wetlands relative to the importance of other system components, such as the river itself; (2) residence
time of the material in the downstream water; and (3) relative importance of the material. In many
cases, the effects on downstream waters need to be considered in aggregate. For example, the
contribution of material by a particular stream or wetland (e.g., a specific ephemeral stream) might be
small, but the aggregate contribution by an entire class of streams or wetlands (e.g., all ephemeral
streams in the river network) might be substantial. Integrating contributions over time also might be
necessary, taking into account the frequency, duration, and timing of material export and delivery.
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1-3
a
" 3
<2. O
(/) o
energy imports to and exports from a stream or wetland, in terms of mass or energy; arrow widths represent relative material mass or energy
and differences in arrow shades represent timing (lag) or composition (transformation) changes. Imports to streams and wetlands can come
from upland terrestrial areas, other streams and wetlands, or from the river itself. Arrows are meant to be illustrative, and do not necessarily
represent upstream/downstream relationships. For example, materials and energy can move downstream, upstream, or laterally into streams
and wetlands. Examples of commonly exchanged materials and energy include water, heat energy, nutrients, contaminants, sediment,
participate organic matter, organisms, and reproductive propagules; note that exchange of materials and energy between streams and wetlands
and downstream systems can result in positive or negative effects on downstream waters.
Function
Definition
Examples
Net increase in a material or energy flux
(exports > imports)
Streams: invertebrate production (Wipfli and Gregovich, 2002)
Wetlands', phytoplankton production from floodplain (Schemel etal., 2004;
Lehman eta I., 2008)
Net decrease in a material or energy flux
(exports < imports)
Streams: upstream fish populations that are not sustainable without net
immigration from downstream areas (Woodford and Mclntosh, 2010)
Wetlands', sediment deposition, denitrification (Johnston, 1991)
ro
to
en
Avoidance of a nearby sink function, thereby
preventing a net decrease in material or energy
flux (exports = imports)
Streams: headwaters as summer coldwater refuges (Curry et al., 1997)
Wetlands', riparian wetlands as aquatic refuges in dryland rivers (Leigh etal.
2010)
River
Temporary storage and subsequent release of
materials or energy without affecting
cumulative flux (exports = imports); delivery is
delayed and can be prolonged
Streams: delay of downstream peak flows due to bank storage (Burt, 1997);
temporary heat storage within the alluvial aquifer (Arrigoni etal., 2008)
Wetlands', flood attenuation (Bullock and Acreman, 2003)
River
Conversion of a material or energy into a
different form; the amount of the base material
or energy is unchanged (base exports = base
imports), but its composition (e.g., mass of the
different forms) can vary
Streams: conversion of coarse to fine pa rticulate organic matter (Wallace etal.,
1995)
Wetlands', mercury methylation (Galloway and Branfireun, 2004; Selvendiran et
a I., 2008)
ro
o
-------
Considering the cumulative material fluxes that originate from a specific stream or wetland, rather than
the individual materials separately, is essential in understanding the effects of material fluxes on
downstream waters (Section 1.2.3).
In general, the more frequently a material is delivered to a river, the greater its effect. The effect of an
infrequently supplied material, however, can be large if the material has a long residence time in the
river (Leibowitz et al., 2008). For example, woody debris might be exported to downstream waters
infrequently but it can persist in downstream channels. In addition, some materials are more important
in defining the structure and function of a river. Using the same example, woody debris can have a large
effect on river structure and function because it affects water flow, sediment and organic matter
transport, and habitat (Harmon et al., 1986; Gurnell et al., 1995). Another example is salmon migrating
to a river: They can serve as a keystone species to regulate other populations and as a source of marine-
derived nutrients (Schindler et al., 2005).
2.3.2 Connectivity and Transport of Materials to and from Streams and
Wetlands
2.3.2.1 Connectivity and Isolation
The functions discussed above represent general mechanisms by which streams and wetlands influence
downstream waters. For these altered material and energy fluxes to affect a river, however, transport
mechanisms that deliver (or could deliver) these materials to the river are necessary. Connectivity
describes the degree to which components of a system are connected and interact through various
transport mechanisms; connectivity is determined by the characteristics of both the physical landscape
and the biota of the specific system. This definition is related to, but is distinct from, definitions of
connectivity based on the actual flow of materials between system components (e.g., Pringle, 2001).
That connectivity among river-system components, including streams and wetlands, plays a significant
role in the structure and function of these systems is not a new concept. In fact, much of the theory
developed to explain how these systems work focuses on connectivity and linkages between system
components (e.g., Section 1.2; Vannote etal., 1980; Newbold etal., 1982a; Newbold etal., 1982b; Junk et
al., 1989; Ward, 1989; Benda etal., 2004; Thorp etal., 2006).
In addition to its central role in defining river systems (Section 2.2.1), water movement through the
river system (Figure 2-6) is the primary mechanism providing physical connectivity both within river
networks and between those networks and the surrounding landscape (Fullerton et al., 2010).
Hydrologic connectivity results from the flow of water, which provides a "hydraulic highway" (Fausch et
al., 2002) along which physical, chemical, and biological materials associated with the water are
transported (e.g., sediment, woody debris, contaminants, organisms).
Ecosystem function within a river system is driven by interactions between the river system's physical
environment and the diverse biological communities living within it (Wiens, 2002; Schroder, 2006).
Thus, river system structure and function also depend on biological connectivity among the system's
populations of aquatic and semiaquatic organisms. Biological connectivity refers to the movement of
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organisms, including transport of reproductive materials (e.g., seeds, eggs, genes) and dormant stages,
through river systems. These movements link aquatic habitats and populations in different locations
through several processes important for the survival of individuals, populations, and species (Sections
3.5, 4.3.4, and 4.4.4). Movements include dispersal, or movement away from an existing population or
parent organism; migration, or long-distance movements occurring seasonally; localized movement
over an organism's home range to find food, mates, or refuge from predators or adverse conditions; and
movement to different habitats to complete life-cycle requirements. At the population and species levels,
dispersal and migration contribute to persistence at local and regional scales via colonization of new
habitats (e.g., Hecnar and McLoskey, 1996; Tronstad etal., 2007); location of mates and breeding
habitats (Semlitsch, 2008); rescue of small populations threatened with local extinction (Brown and
Kodric-Brown, 1977); and maintenance of genetic diversity (e.g., Waples, 2010). These movements can
result from passive transport by water, wind, or other organisms (e.g., birds, terrestrial mammals);
active movement with or against water flow (e.g., upstream fish migration); or active movement over
land (for organisms capable of terrestrial dispersal) or through the air (for birds or insects capable of
flight; Figure 1-1B). Thus, biological connectivity can occur within aquatic ecosystems or across
ecosystem or watershed boundaries, and it can be multidirectional. For example, organisms can move
downstream from perennial, intermittent, and ephemeral headwaters to rivers; upstream from
estuaries to rivers to headwaters; and laterally between floodplain wetlands, geographically isolated
wetlands, rivers, lakes, or other water bodies. Significant biological connectivity can also exist between
aquatic and terrestrial habitats (Nakano et al., 1999; Gibbons, 2003; Baxter et al., 2004), but our focus is
on connections among components of aquatic systems (Section 1.3).
As noted in Section 2.2.3, streams and rivers are not pipes (Bencala, 1993; Bencala et al., 2011); they
provide opportunities for water to interact with internal components (e.g., alluvium, organisms)
through the five functions by which streams and wetlands alter material fluxes (Table 2-1). Connectivity
between streams and wetlands provides opportunities for material and energy fluxes to be altered
sequentially by multiple streams and wetlands as the materials are transported downstream. The
aggregate effect of these sequential fluxes determines the proportion of material that ultimately reaches
the river. The form of the exported material can change as it moves down the river network (Figure 2-
14), however, making quantitative assessments of the importance of individual stream and wetland
resources within the entire river system difficult. For example, organic matter can be exported from
headwater streams and consumed by downstream macroinvertebrates (Figure 2-14). Those
invertebrates can drift farther downstream and be eaten by juvenile fish that eventually move into the
mainstem of the river, where they continue to feed and grow.
The assessment of stream and wetland influence on rivers also is complicated by the cumulative time lag
resulting from these sequential transformations and transportations. For example, removal of nutrients
by streambed algal and microbial populations, subsequent feeding by fish and insects, and release by
excretion or decomposition delays the export of nutrients downstream (Figure 2-14).
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Figure 2-14. Illustration of the sequential transformation of materials as they move through the
river network, via either downstream transport with water flow (solid black arrows) or via aerial or
terrestrial movements (dashed black arrows). Here, an ephemeral headwater stream exports organic
matter (at left) and an intermittent headwater stream exports ammonium, which is incorporated into
algal biomass (at right). Macroinvertebrates consume these basal food resources and transform them
reaches.
The opposite of connectivity is isolation, or the degree to which transport mechanisms (i.e., pathways
between system components) are lacking; isolation acts to reduce material fluxes between system
components. Although here we primarily focus on the benefits that connectivity can have on
downstream systems, isolation also can have important positive effects on the condition and function of
downstream waters. For example, waterborne contaminants that enter a wetland cannot be transported
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to a river if the wetland is hydrologically isolated from the river, except by nonhydrologic pathways.
Increased isolation can decrease the spread of pathogens (Hess, 1996) and invasive species (e.g.,
Bodamer and Bossenbroek, 2008), and increase the rate of local adaptation (e.g., Fraser et al., 2011).
Thus, both connectivity and isolation should be considered when examining material fluxes from
streams and wetlands, and biological interactions should be viewed in light of the natural balance
between these two factors.
When assessing the effects of connectivity or isolation and the five general functions (sources, sinks,
refuges, lags, and transformations; Table 2-1) on downstream waters, dimensions of time and space
must be considered. Water or organisms transported from distant headwater streams or wetlands
generally will take longer to travel to a larger river than materials transported from streams or wetlands
near the river (Section 2.4.2). This can introduce a lag between the time the function occurs and the time
the material arrives at the river. In addition, the distribution of streams and wetlands can be a function
of their distance from the mainstem channel. For example, in a classic dendritic network, there is an
inverse geometric relationship between number of streams and stream order. In such a case, the
aggregate level of function could be greater for terminal source streams, compared to higher order or
lateral source streams. This is one reason why watersheds of terminal source streams often provide the
greatest proportion of water for major rivers. Connectivity, however, results from many interacting
factors (Section 2.4.5). For example, the relationship between stream number and order can vary with
the shape of the watershed and the configuration of the network (Section 2.4.2). Thus, caution must be
exercised when generalizing about these spatial and temporal relationships. Spatial and temporal
variability of connectivity is discussed below, and the factors influencing them are considered in
Section 2.4.
2.3.2.2 Spatial and Temporal Variability of Connectivity
Connectivity is not a fixed characteristic of a system, but varies over space and time (Ward, 1989;
Leibowitz, 2003; Leibowitz and Vining, 2003). Variability in hydrologic connectivity results primarily
from the longitudinal (Figures 2-8 and 2-10) and lateral (Figure 2-12) expansion and contraction of the
river network and transient connection with other components of the river system (Section 2.2.3). The
variability of connectivity can be described in terms of frequency, duration, magnitude, timing, and rate
of change (Section 1.2.2).
The expansion and contraction of river networks affect the extent, magnitude, timing, and type of
hydrologic connectivity. For example, intermittent and ephemeral streams (Figure 2-7) flow only during
wetter seasons (Section 2.4) or during and immediately following precipitation events. Thus, the spatial
extent of connectivity between streams and wetlands and rivers increases greatly during these high-
flow events because intermittent and ephemeral streams are estimated to account for 59% of the total
length of streams in the contiguous United States (Nadeau and Rains, 2007). Changes in the spatial
extent of connectivity due to expansion and contraction are even more pronounced in the arid and
semiarid Southwest, where more than 80% of all streams are intermittent or ephemeral (Figures 2-9B
Stream and Wetland Connectivity: January 2015
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and B-5; Levicketal., 2008). Expansion and contraction also affect the magnitude of connectivity
because larger flows provide greater potential for material transport (e.g., Section 3.3.2).
Besides affecting the spatial extent and magnitude of hydrologic connectivity, expansion and contraction
of the stream network also affect the duration and timing of flow in different portions of the network.
Perennial streams have year-round connectivity with a downstream river, whereas intermittent streams
have seasonal connectivity. The temporal characteristics of connectivity for ephemeral streams depend
on the duration and timing of storm events (Figure B-10). Similarly, connectivity between wetlands and
downstream waters can range from permanent to seasonal to episodic.
The expansion and contraction of river systems also affect the type of connectivity. For example, during
wet periods when input from precipitation can exceed evapotranspiration and available storage, non-
floodplain wetlands could have connectivity with other wetlands or streams through surface spillage
(Leibowitz and Vining, 2003; Rains et al., 2008). When spillage ceases due to drier conditions,
hydrologic connectivity could only occur through ground water (Rains et al., 2006; Rains et al., 2008).
When the flow of water mediates dispersal, migration, and other forms of biotic movement, biological
and hydrologic connectivity can be tightly coupled. For example, seasonal flooding of
riparian/floodplain wetlands creates temporary habitat that fish, aquatic insects, and other organisms
use (Junk et al., 1989; Smock, 1994; Tockner et al., 2000; Robinson et al., 2002; Tronstad et al., 2007).
Factors other than hydrologic dynamics also can affect the temporal and spatial dynamics of biological
connectivity. Such factors include movement associated with seasonal habitat use (Moll, 1990;
Lamoureux and Madison, 1999) and shifts in habitat use due to life-history changes (Huryn and Gibbs,
1999; Gibbons et al., 2006; Subalusky et al., 2009a), quality or quantity of food resources (Smock, 1994),
presence or absence of favorable dispersal conditions (Schalk and Luhring, 2010), physical differences
in aquatic habitat structure (Grant et al., 2007), or the number and sizes of nearby populations (Gamble
et al., 2007). For a specific river system with a given spatial configuration, variability in biological
connectivity also occurs due to variation in the dispersal distance of organisms and reproductive
propagules (Section 2.4.4; Semlitsch and Bodie, 2003).
Finally, just as connectivity from temporary or seasonal wetting of channels can affect downstream
waters, temporary or seasonal drying also can affect river networks. Riverbeds or streambeds that
temporarily dry up are used by aquatic organisms that are specially adapted to wet and dry conditions,
and can serve as egg and seed banks for several organisms, including aquatic invertebrates and plants
(Steward et al., 2012). These temporary dry areas also can affect nutrient dynamics due to reduced
microbial activity, increased oxygen availability, and inputs of terrestrial sources of organic matter and
nutrients (Steward etal., 2012).
2.4 Factors Influencing Connectivity
Numerous factors affect physical, chemical, and biological connectivity within river systems. These
factors operate at multiple spatial and temporal scales, and interact with each other in complex ways to
Stream and Wetland Connectivity: January 2015
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determine where components of a system fall on the connectivity-isolation gradient at a given time. In
this section, we focus on four key factors—climate-watershed characteristics, spatial distribution
patterns, biota, human activities and alterations. These are by no means the only factors influencing
connectivity, but they illustrate how many different variables shape physical, chemical, and biological
connectivity. We also examine how interactions among different factors influence connectivity, using as
an example wetlands in the prairie pothole region.
2.4.1 Climate-Watershed Characteristics
The movement and storage of water in watersheds varies with climatic, geologic, physiographic, and
edaphic characteristics of river systems (Winter, 2001; Wigington etal., 2013). Atthe largest spatial
scale, climate determines the amount, timing, and duration of water available to watersheds and river
basins. Key characteristics of water availability that influence connectivity include annual water surplus
(precipitation minus evapotranspiration), timing (seasonality) of water surplus during the year that is
heavily influenced by precipitation timing and form (e.g., rain, snow), and rainfall intensity.
Annual runoff generally reflects water surplus and varies widely across the United States (Figure 2-15).
Seasonality of water surplus during the year determines when and for how long runoff and ground-
water recharge occur. Precipitation and water surplus in the eastern United States is less seasonal than
in the West (Finkelstein and Truppi, 1991). The Southwest experiences summer monsoonal rains
(Section B.5), whereas the West Coast and Pacific Northwest receive most precipitation during the
winter season (Wigington et al., 2013). Throughout the West, winter precipitation in the mountains
occurs as snowfall, where it accumulates in seasonal snowpack and is released during the spring and
summer melt seasons to sustain streamflow during late spring and summer months (Brooks et al.,
2012). The flowing portions of river networks tend to have their maximum extent during seasons with
the highest water surplus (Section 2.2.3; Figure 2-10), when conditions for flooding are most likely.
Typically, the occurrence of ephemeral and intermittent streams is greatest in watersheds with low
annual runoff and high water surplus seasonality but also is influenced by watershed geologic and
edaphic features (Gleeson et al., 2011).
Rainfall intensity can affect hydrologic connectivity in localities where watershed surfaces have low
infiltration capacities relative to rainfall intensities. Infiltration-excess overland flow occurs when
rainfall intensity exceeds watershed surface infiltration, and it can be an important mechanism in
providing water to wetlands and river networks (Goodrich et al., 1997; Levick et al., 2008). Overland
flow is common at low elevations in the Southwest, due to the presence of desert soils with low
infiltration capacities combined with relatively high rainfall intensities (Section B.5). The Pacific
North west has low rainfall intensities, whereas many locations in the Mid-Atlantic, Southeast, and Great
Plains have higher rainfall intensities. The prevalence of impermeable surfaces in urban areas can
generate overland flow in virtually any setting (Booth et al., 2002).
River system topography and landscape form can profoundly influence river network drainage patterns,
distribution of wetlands, and ground-water and surface-water flowpaths. Winter (2001) described six
generalized hydrologic landscape forms (Figure 2-16) common throughout the United States. Mountain
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I 3
Figure 2-15. Map of annual runoff in contiguous United States showing locations of five example streams that illustrate daily runoff patterns
and total annual runoff depths. (A) Rapidan River, VA; (B) Noyo River, CA; (C) Crystal River, CO; (D) San Pedro River, AZ; and (E) Metolius River,
OR. All data are from http://waterdata.usgs.gov/usa/nwis/sw (downloaded June 27, 2011). Runoff can be conceived as the difference between
precipitation and evapotranspiration at the watershed scale. The varied runoff patterns in the five rivers result from divergent climate, geology,
and topography.
<£. O
(/) o
<25.4mm
i 25.4 254mm
1254-508mtn
I -508mm
C. Crystal River, CO |2009 annual runoff = 679 mm)
100.00
: 10.00
; 1.00
0.10
0.01
c*
•< < <
!O
(JO
ro
A. Rapidan River, VA |2005 annual runoff = 450 mm|
100.00
• 10.00
'. 1.00
0.10
0.01
f 9" Jf
* *i <
D. San Pedro River, AZ [2005 annual runoff = 9 mm]
100.00
10.00
0.10
0.01
.«- A O
*° oa
to
o
B. Noyo River, CA |2006 annual runoff = 1338 runoff]
100.00
I 10.00
\ 1.00
! 0.10
I
: 0.01
E. Metolius River, OR |2008 annual runoff = 1744 mm|
100.00
10.00
t 1.00
c 0.10
3
tt 0.01
/
S; ^? (
* i? J(
< < •>.
-------
Figure 2-16. Generalized hydrologic landscape forms. (A) Mountain Valley: narrow uplands and
lowlands separated by large steep valley sides; (B) Playa: large broad lowland separated from narrow
uplands by steeper valley sides (playas and basins of interior drainage); (C) Plateau and High Plains:
small narrow lowlands separated from broad uplands by steeper valley sides; (D) Riverine Valley:
small fundamental landscape units nested inside broader fundamental landscape unit; (E) Coastal
Terrain: small fundamental landscape units nested inside broader fundamental landscape unit
(coastal plain with terraces and scarps); and (F) Hummocky Terrain: small fundamental landscape
units superimposed randomly on larger fundamental landscape unit. A fundamental hydrologic
landscape unit is defined by land-surface form, geology, and climate. Modified from Winter (2001).
A. Mountain Valley
IUnsaturated
zone
I Saturated Surface
zone water
Ground-water
flow
Water
table
B. Playa
C. Plateau and High Plains
D. Riverine Valley
E. Coastal Terrain
F. Hummocky Terrain
Valleys (Figure 2-16A) and Plateaus and High Plains (Figure 2-16C) have constrained valleys through
which streams and rivers flow. The Mountain Valleys form has proportionally long, steep sides with
narrow to nonexistent floodplains resulting in the rapid movement of water downslope. In contrast,
Riverine Valleys (Figure 2-16D) have extensive floodplains that promote strong surface-water,
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hyporheic water, and alluvial ground-water connections between wetlands and rivers. Small changes in
water table elevations can influence the water levels and hydrologic connectivity of wetlands over
extensive areas in this landscape form (Figure 2-16D). Local ground-water flowpaths are especially
important in Hummocky Terrain (Figure 2-16F). Constrained valleys, such as the Mountain Valley
landform (Figure 2-16A), have limited opportunities for the development of floodplains and alluvial
aquifers, whereas unconstrained valleys, such as the Riverine Valley landform (Figure 2-16D), provide
opportunities for the establishment of floodplains. Some river basins can be contained within a single
hydrologic landscape form, but larger river basins commonly comprise complexes of hydrologic
landscape forms. For example, the James River in Virginia, which flows from mountains through the
Piedmont to the Coastal Plain, is an example of a Mountain Valley-High Plateaus and Plains-Coastal
Terrain-Riverine Valley complex.
Floodplain hydrologic connectivity to rivers and streams occurs primarily through overbank flooding,
shallow ground-water flow, and hyporheic flow (Section 2.2). Water-table depth can influence
connectivity across a range of hydrologic landscape forms, but especially in floodplains. Rivers and
wetlands can shift from losing reaches (or recharge wetlands) during dry conditions to gaining reaches
(or discharge wetlands) during wet conditions. Wet, high water-table conditions influence both ground-
water and surface-water connectivity. When water tables are near the watershed surface, they create
conditions in which swales and small stream channels fill with water and flow to nearby water bodies
(Wigington et al., 2003; Wigington et al., 2005). Nanson and Croke (1992) noted that a complex
interaction of fluvial processes forms floodplains, but their character and evolution are essentially a
product of stream power (the rate of energy dissipation against the bed and banks of a river or stream)
and sediment characteristics. They proposed three floodplain classes based on the stream power-
sediment characteristic paradigm: (1) high-energy noncohesive, (2) medium-energy noncohesive, and
(3) low-energy cohesive. The energy term describes stream power during floodplain formation, and the
cohesiveness term depicts the nature of material deposited in the floodplain. The cohesiveness term is
also related to the hydraulic properties of alluvial aquifers. Alluvium for Class 1 and 2 floodplains tends
to have higher hydraulic conductivity, or a higher rate at which water moves through a saturated,
permeable soil or rock layer, than Class 3 floodplains. The higher the hydraulic conductivity of an
alluvial aquifer, the greater the exchange rate between the alluvial aquifer and river waters (Whiting
and Pomeranets, 1997). In addition, hyporheic and alluvial aquifer exchanges are more responsive to
seasonal discharge changes in floodplains with complex topography (Poole et al., 2006).
Within hydrologic landscape forms, soil and geologic formation permeabilities are important
determinants of hydrologic flowpaths (Figure 2-17). Permeable soils promote infiltration that results in
ground-water hydrologic flowpaths (Figures 2-17A and 2-17B), whereas the presence of impermeable
soils with low infiltration capacities is conducive to overland flow (Figures 2-17C and 2-17D). In
situations in which ground-water outflows from watersheds or landscapes dominate, the fate of water
depends in part on the permeability of deeper geologic strata. The presence of an aquiclude near the
watershed surface leads to shallow subsurface flows through soil or geologic materials (Figure 2-17A).
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Figure 2-17. Major hydrologic flowpaths for hillslopes with combinations of permeable and
impermeable soils and geologic formations. (A) Permeable soil and impermeable underlying
geologic formation; (B) permeable soil and permeable underlying geologic formation; (C)
impermeable soil and impermeable underlying geologic formation; and (D) impermeable soil and
permeable underlying geologic formation. Width of arrow indicates relative magnitude of flow. Note
that pavement can be another source of impermeable surfaces and subsequent overland flow in
anthropogenically influenced settings.
Precipitation
A.
Pre
cipitati
on
Precipitation
Precipitation
1 Permeable soil
• Permeable geologic
formation
8sl3 Impermeable soil
ggSl Impermeable geologic
^^ formation
> Ground-water
recharge
Overland
. / fl«w
Local
. ground-water
W flow
These local ground-water flowpaths connect portions of watersheds to nearby wetlands or streams
(Figure 2-3). Alternatively, if a deep permeable geologic material (an aquifer) is present, water is likely
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to move farther downward within watersheds and recharge deeper aquifers (Figure 2-17B). The
permeability of soils and geologic formations both can influence the range of hydrologic connectivity
between non-floodplain wetlands and river networks. For example, a wetland that is the origin of a
stream can have a permanent or temporary surface-water connection with downstream waters through
a channelized outlet (Figure 2-18A); a wetland can be connected to downstream waters by transient
surface-water flows through swales (Figure 2-18B) or by shallow ground-water flows (Figure 2-18C); or
a wetland can be hydrologically isolated from downstream waters (Figure 2-18D) because it recharges a
deep ground-water aquifer that does not feed surface waters, or it is located in a basin where
evapotranspiration is the dominant form of water loss.
The importance of climate-watershed interactions in determining the amount and seasonality of water
surpluses, the timing and duration of streamflow, and thus the timing and extent of hydrologic
connectivity, is illustrated by annual hydrographs for five rivers in different regions of the United States
(Figure 2-15).
The hydrograph for the Rapidan River in Virginia (Figure 2-15A) illustrates the uniform annual
precipitation pattern of the East (with small variations due to increased evapotranspiration in the
summer months) interacting with a steep Blue Ridge Mountain watershed comprising metamorphic
bedrock with alluvial and colluvial fill in the lower riparian areas (Castro and Hornberger, 1991).
Hydrologic events driven by rainfall can occur anytime during the year, but are especially common in
winter and spring months; these events result in expansion of the river network as ephemeral streams
flow. Baseflow sustains perennial flow over a large part of the network.
Located in a region of steep slopes and impermeable bedrock (Mayer and Naman, 2011), the Noyo River
watershed in California (Figure 2-15B) has highly seasonal water surplus because rainfall occurs
primarily from November through May and the impermeable bedrock prevents precipitation water from
moving to deep ground water. Consequently, runoff timing is similar to precipitation temporal patterns.
Total runoff for the basin is high, and baseflow levels are high during the winter and low during the dry
summer season. These low baseflow periods create conditions favorable for intermittent flows in
streams with significant channel alluvium (Wigington et al., 2006).
The Crystal River of Colorado (Figure 2-15C) drains a glaciated landscape in the upper portion of the
Gunnison River in the Colorado Rocky Mountains. It has protracted high flow during the spring that is
controlled by the accumulation and melting of snow in the basin's higher elevations during the winter
and subsequent melting during spring and summer. This streamflow pattern also promotes the
occurrence of intermittently flowing streams due to large water surplus differences between the high-
flow and low-flow periods.
Total runoff in the San Pedro River, Arizona is low (Figure 2-15D), and short, intense rainstorms during
the summer monsoons commonly drive hydrologic events (Levick et al., 2008). Because a major
proportion of water reaching the San Pedro River originates as overland flow to ephemeral streams that
ultimately flow to the mainstem river, baseflow is limited (Section B-5). In other San Pedro River
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Figure 2-18. Types of hydrologic connections between non-floodplain wetlands and streams or
rivers. (A) A wetland connected to a river by surface flow through a headwater stream channel. (B) A
wetland connected to a river by surface flow through a nonchannelized swale. Such a wetland would
be considered geographically isolated if the swale did not meet the Cowardin et al. (1979)
three-attribute wetland criteria. (C) A geographically isolated wetland connected to a river by ground-
water flow (flowpath can be local, intermediate, or regional). (D) A geographically isolated wetland
that is hydrologically isolated from a river. Note that in A-C, flows connecting the wetland and river
may be perennial, intermittent, or ephemeral.
A.
B.
D.
Wetland with
surface outlet
Geographically
isolated wetland
Headwater stream
^^™ Stream/river
- - • • Ground-water flow
LX~'J Swale
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mainstem reaches, ground-water flow from regional and alluvial aquifers supports baseflow (Dickinson
et al., 2010). Like the Crystal River, the Metolius River in Oregon (Figure 2-15E) also has snowpack in its
higher elevations, but geologic conditions in the watershed alter the climate signal. Meltwaters in the
Metolius River flow through long flowpaths in porous bedrock to springs in or near the river (James et
al., 2000; Gannett et al., 2001). Although intermittent and ephemeral streams occur in the Metolius
basin, most streams are spring-fed and perennial.
2.4.2 Spatial Distribution Patterns
Climate and watershed characteristics directly affect spatial and temporal patterns of connectivity
between streams and wetlands and rivers by influencing the timing and extent of river network
expansion and contraction. They also influence the spatial distribution of water bodies within a
watershed (e.g., Tihansky, 1999), and in particular, the spatial relationship between those water bodies
and the river.
Hydrologic connectivity between streams and rivers can be a function of the distance between the two
water bodies (Bracken and Croke, 2007; Peterson etal., 2007). If channels functioned as pipes, this
would not be the case, and any water and its constituent materials exported from a stream eventually
would reach the river. Because streams and rivers are not pipes (Section 2.2.3; Bencala, 1993), water
can be lost from the channel through evapotranspiration and bank storage and diluted through
downstream inputs. Thus, material from a headwater stream that flowed directly into the river would be
subject to less transformation or dilution. On the other hand, the greater the distance a material travels
between a particular stream reach and the river, the greater the opportunity for that material to be
altered (e.g., taken up, transformed, or assimilated) in intervening stream reaches; this alteration could
reduce the material's direct effect on the river, but it could also allow for beneficial transformations. For
example, organic matter exported from a headwater stream located high in a drainage network might
never reach the river in its original form, instead becoming reworked and incorporated into the food
web (Figure 2-14). Similarly, higher order streams generally are located closer to rivers and, therefore,
can have higher connectivity than upstream reaches of lower order. Note that although an individual
low-order stream can have less connectivity than a high-order stream, a river network has many more
low-order streams, which can represent a large portion of the watershed (Section 3.2); thus, the
magnitude of the cumulative effect of these low-order streams can be significant.
The relationship between streams and the river network is a function of basin shape and network
configuration. Elongated basins tend to have trellis networks where relatively small streams join a
larger mainstem (Figure 2-19A); compact basins tend to have dendritic networks with tree-like
branching, where streams gradually increase in size before joining the mainstem (Figure 2-19B). This
network configuration describes the incremental accumulation of drainage area along rivers, and
therefore provides information about the relative contributions of streams to downstream waters.
Streams in a trellis network are more likely to connect directly to a mainstem, compared with a
dendritic network. The relationship between basin shape, network configuration, and connectivity,
however, is complex. A mainstem in a trellis network also is more likely to have a lower stream order
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Figure 2-19. Major types of basin shapes and network configurations. (A) A rectangular basin with
trellis network, and (B) a compact basin with a dendritic network.
A.
than one in a dendritic network. For example, the lowest reach in the trellis network in Figure 2-19A is a
third-order stream, while that of the dendritic network (Figure 2-19B) is a fourth-order stream.
Distance also affects connectivity between non-floodplain and riparian/floodplain wetlands and
downstream waters. Riverine wetlands that serve as origins for lateral source streams that connect
directly to a mainstem river have a more direct connection to that river than wetlands that serve as
origins for terminal source streams high in a drainage network. This also applies to riparian/floodplain
wetlands that have direct surface-water connections to streams or rivers. If geographically isolated non-
floodplain wetlands have surface-water outputs (e.g., depressions that experience surface-water spillage
or ground-water seeps; Figure 2-18B), the probability that surface water will infiltrate or be lost
through evapotranspiration increases with distance. For non-floodplain wetlands connected through
ground-water flows, less distant areas are generally connected through shallower flowpaths
(Figure 2-5), assuming similar soil and geologic properties. These shallower ground-water flows have
the greatest interchange with surface waters (Section 2.2.2) and travel between points in the shortest
amount of time. Although elevation is the primary factor determining areas that are inundated through
overbank flooding, connectivity with the river generally will be higher for riparian/floodplain wetlands
located near the river's edge compared with riparian/floodplain wetlands occurring near the floodplain
edge.
Distance from the river network also influences biological connectivity among streams and wetlands.
For example, mortality of an organism due to predators and natural hazards generally increases with
the distance it has to travel to reach the river network. The likelihood that organisms or propagules
traveling randomly or by diffusive mechanisms such as wind will arrive at the river network generally
decreases as distance increases.
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The distribution of distances between wetlands and river networks depends on both the drainage
density of the river network (the total length of stream channels per unit area) and the density of
wetlands. Climate and watershed characteristics influence these spatial patterns, which can vary widely.
For example, a subset of fens in New York State was located closer to each other, on average, than a
subset of Carolina bays at the Savannah River Site: The proportion of wetlands located at distances of 0-
100,100-500, and >500 m was 27, 39, and 35%, respectively, for the fens and 12, 44, and 44% for the
Carolina bays, respectively (Bedford and Godwin, 2003; Sharitz, 2003). When interpreting such
distributions, however, other factors that affect connectivity (e.g., differences in soils or slope) should be
considered.
Figure 2-20 compares the spatial distribution of wetlands and streams to the river network in six
different landscape settings. The figure shows landscape settings ranging from no nearby streams and
dense small wetlands (Figure 2-20A), to a few nearby streams with high wetland density (Figures 2-20B
and 2-20C), to less spatially uniform wetlands (Figure 2-20D), to areas with higher drainage densities
and riparian (Figure 2-20E) or larger, more extensive (Figure 2-20F) wetlands. The maps in Figure 2-20
are single examples of these various settings, so they might not be representative. They are useful,
however, for illustrating the degree to which landscape setting can affect the interspersion—and thus
average distance—between wetlands and the river network, and the large variability that can result. In
settings having many wetlands and relatively low drainage density (Figures 2-20B, 2-20C, and 2-20D),
the distances between individual wetlands and the stream can vary greatly. In contrast, the distances in
areas having a higher drainage density (Figure 2-20E and 2-20F) are shorter and vary less. All factors
being equal, wetlands closer to the stream network will have greater hydrologic and biological
connectivity than wetlands located farther from the same network.
2.4.3 Biota
Biological connectivity results from the interaction of physical characteristics of the
environment—especially those facilitating or restricting dispersal—and species' traits or behaviors, such
as life-cycle requirements, dispersal ability, or responses to environmental cues (Section 1.2.2). Thus,
the types of biota within a river system are integral in determining the river system's connectivity, and
landscape features or species traits that necessitate or facilitate movement of organisms tend to
increase biological connectivity among water bodies.
Diadromous fauna (e.g., Pacific and Atlantic salmon, certain freshwater shrimps and snails, American
eels), which require both freshwater and marine habitats over their life cycles and therefore migrate
along river networks, provide one of the clearest illustrations of biological connectivity. Many of these
taxa are either obligate or facultative users of headwater streams (Erman and Hawthorne, 1976;
Wigington et al., 2006), meaning that they either require (obligate) or can take advantage of (facultative)
these habitats; these taxa thereby create a biological connection along the entire length of the river
network. For example, many Pacific salmon species spawn in headwater streams, where their young
grow for a year or more before migrating downstream, living their adult life stages in the ocean, and
then migrating back upstream to spawn. Many taxa also can exploit temporary hydrologic connections
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Figure 2-20. Examples of different landscapes showing interspersion of wetlands and streams or
rivers. (A) Prairie potholes within the Missouri Coteau in North Dakota; (B) prairie potholes within the
Drift Prairie in North Dakota; (C) playas in Texas; (D) vernal pools in California; (E) bottomland
hardwood wetlands in Illinois; and (F) Carolina bays in North Carolina. Note all maps are at the same
scale. Wetlands smaller than the minimum mapping unit (currently 0.4 ha) might not appear on
maps. Source: National Wetlands Inventory Wetlands Mapper
(http://www.fws.gov/wetlands/Data/Mapper.html).
A. Prairie potholes (Missouri Coteau)
••->»!
' *
'
U. Prairie potholes (Drift Prairie)
I
^Wfe-JT^^X'l
•^'^j&is^V
, v. ' * *W ._•_»•'• •»,
"*^V-'.>?:1^?^«
|>... ;^^^
5' ,».,-•: V
-------
Figure 2-20. Examples of different landscapes showing interspersion of wetlands and streams or
rivers (continued).
C. Playa
D. Vernal pools
Wetlands
Riparian
d] Freshwater Emergent CZ1 Estuarine and Marine CZI Riverine Herbaceous
^•Freshwater Forested/Shrub d] Freshwater Pond CD Other CZI Forested/Shrub
• Estuarine and Marine Deepwater • Lake
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January 2015
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Figure 2-20. Examples of different landscapes showing interspersion of wetlands and streams or
rivers (continued).
E. Bottomland hardwood wetlands
••'..«
w
,
-.-W' :,v';> •••:*&<' :
• •j •
•
• .""••a.irv • .> •
:&. • •:»• .-•••• *-I
V-, ' ' - AV • • /•
'
F. Carolina bay$
Wetlands Riparian
EH Freshwater Emergent CU Estuarine and Marine •Riverine I Herbaceous
^•Freshwater Forested/Shrub I ] Freshwater Pond L J Other I ) Forested/Shrub
^B Estuarine and Marine Deepwater ^B Lake
between rivers and floodplain wetland habitats caused by flood pulses (Section 1.2.1; Junk et al., 1989;
Tockner et al., 2000), moving into these wetlands to feed, reproduce, or avoid harsh environmental
conditions and then returning to the river network (Copp, 1989; Smock, 1994; Richardson et al., 2005).
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Biological connectivity does not solely depend on diadromy, however, as many nondiadromous
organisms are capable of significant movement within river networks (Section 1.2.2). For example,
organisms such as pelagic-spawning fish and mussels release eggs or larvae that disperse downstream
with water flow (e.g., Platania and Altenbach, 1998; Schwalb et al., 2010); many fish swim significant
distances both upstream and downstream (e.g., Gorman, 1986; Hitt and Angermeier, 2008); and many
aquatic macroinvertebrates move or drift downstream (e.g., Elliott, 1971; Miiller, 1982; Brittain and
Eikeland, 1988; Elliott, 2003). Taxa capable of movement over land, via either passive transport (e.g.,
wind dispersal or attachment to animals capable of terrestrial dispersal) or active movement (e.g.,
terrestrial dispersal or aerial dispersal of winged adult stages), can establish biotic linkages between
river networks and wetlands, as well as linkages across neighboring river systems (Hughes et al., 2009).
The fundamental influence that biological connectivity has on species distributions can last long after a
system is disconnected. In a global analysis of freshwater fish diversity, Bias et al. (2014) found that
paleoconnected drainage basins (basins that had hydrologic connections during the most recent glacial
maximum) currently have greater species richness and lower endemism and beta diversity than
paleodisconnected basins. This study indicates that hydrologic connectivity, by allowing dispersal of
aquatic organisms, can have a long-lasting legacy effect on the geographic distribution of species.
The examples discussed above illustrate how environmental characteristics provide the physical
structure through which biological connectivity occurs, as mediated by biological traits and behavior.
The physical structure of the environment is not static, however, and also can be altered by biological
behavior. The beaver (Castor canadensis) is a keystone species that builds dams that can alter
connectivity in several ways. Most obviously, beaver dams reduce hydrologic connectivity by
impounding streams and modifying conditions above the dam from lotic to lentic. The area impounded
by beaver dams can be large: In the Kabetogama Peninsula of Minnesota, impounded area accounted for
up to 13% of the landscape, with an average pond area of about 4 ha (Johnston and Naiman, 1990a, b).
In a review of the effects of beaver on stream ecosystems, Collen and Gibson (2001) noted that, although
the hydrologic effects of a single beaver dam can be small, the impact of a series of dams on streams can
be significant; for example, up to 30% of the water in an Oregon catchment was impounded by beaver
dams. Such dams can directly affect material transport (e.g., the ability of the stream to carry sediment is
reduced) and alter biogeochemical characteristics (Naiman et al., 1994; Collen and Gibson, 2001).
Beaver dams also can affect biological connectivity, for example, by obstructing upstream migration, and
cause changes in fish distributions (Collen and Gibson, 2001).
2.4.4 Human Activities and Alterations
Human activities frequently alter connectivity between headwater streams, riparian/floodplain
wetlands, non-floodplain wetlands, and downgradient river networks (Sections 1.2.4, 3.2, 4.3, and 4.4).
In doing so, they alter the transfer and movement of materials and energy between river system
components. In fact, the individual or cumulative effects of headwater streams and wetlands on river
networks often become discernible only following human-mediated changes in degree of connectivity.
These human-mediated changes can increase or decrease hydrologic and biological connectivity (or,
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alternatively, decrease or increase hydrologic and biological isolation). For example, activities and
alterations such as dams, levees, water abstraction, piping, channelization, and burial can reduce
hydrologic connectivity between streams and wetlands and rivers, whereas activities and alterations
such as wetland drainage, irrigation, impervious surfaces, interbasin transfers, and channelization can
increase hydrologic connections. Biological connectivity can be affected similarly: For example, dams
and impoundments might impede biotic movement, whereas nonnative species introductions artificially
increase biotic movement. Further complicating the issue is that a given activity or alteration might
simultaneously increase and decrease connectivity, depending on which part of the river network is
considered. For example, channelization and levee construction reduce lateral expansion of the river
network (thereby reducing hydrologic connections with floodplains), but might increase this
connectivity downstream due to increased frequency and magnitude of high flows.
To illustrate, we describe two notable alterations that affect river system connectivity: dams (and their
associated impoundments) and wetland drainage. The United States has more than 80,000 dams, over
6,000 of which exceed 15 m in height (USAGE, 2009). Numerous studies have shown that dams impede
biotic movements, reduce biological connectivity between upstream and downstream locations (e.g.,
Greathouse et al., 2006; Hall et al., 2011), and form a discontinuity in the normal stream-order-related
progression in stream ecosystem structure and function (Stanford and Ward, 1984). Dams, however, can
have the opposite effect with respect to natural lakes: increasing their biological connectivity with
respect to invasive species by adding impoundments that decrease average distances between lakes and
serving as stepping stone habitat (Johnson et al., 2008). Upstream of large dams, riparian areas are
permanently inundated, increasing lateral hydrologic connectivity. Downstream, dams decrease peak
stream volumes during the normal high-runoff seasons, while increasing minimum flows during normal
low-flow seasons—an overall dampening of stream-flow variability (Poff etal., 2007). Because many
riverine organisms are adapted (via life history, behavioral, and morphological characteristics) to the
seasonality of natural flow regimes, dampening flow variability can have deleterious effects on species
persistence where dams have been built (Lytle and Poff, 2004). This reduction in high flows also
decreases the connectivity of riparian wetlands with the stream by reducing the potential for overbank
lateral flow. Reducing overbank lateral flow can affect downstream water quality, because overbank
flow deposits sediment and nutrients that otherwise remain entrained in the river (Hupp et al., 2009).
The greatest human impact on riparian/floodplain wetlands and non-floodplain wetlands has been
through wetland drainage (Figure 2-21), primarily for agricultural purposes. Estimates show that, in the
conterminous United States, states have lost more than half their original wetlands, with some losing
more than 90%; wetland surface areas also have declined significantly (Dahl, 1990).
Drainage causes a direct loss of function and connectivity in cases where wetland characteristics are
completely lost. In the Des Moines lobe of the prairie pothole region, where more than 90% of the
wetlands have been drained, a disproportionate loss of smaller and larger wetlands has
occurred. Accompanying this loss have been significant decreases in perimeter-area ratios—which are
associated with greater biogeochemical processing and ground-water recharge rates—and increased
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Figure 2-21. Comparison of percent wetland loss between (A) the 1780s and mid-1980s with (B)
the distribution of artificially drained agricultural land in 1985. One dot equals 8100 ha. From Blann
et al. (2009), as modified from Dahl (1990).
A. Percent of Wetlands Lost, 1780s-1980s
CH Less than 50
CH 50-95 (16 States)
CD More than 95 (6 States)
The
Everglades
B. Artificially Drained Agricultural Land, 1985 (1 dot = 8100 ha)
mean distances between wetlands, which reduces biological connectivity (Van Meter and Basu, In
press). Wetland drainage also increases hydrologic connectivity between the landscape—including
drained areas that retain wetland characteristics—and downstream waters. Effects of this enhanced
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hydrologic connectivity include (1) reduced water storage and more rapid conveyance of water to the
network, with subsequent increases in total runoff, baseflows, stormflows, and flooding risk (Wiskow
and van der Ploeg, 2003; Blann etal., 2009); (2) increased delivery of sediment and pollutants to
downstream waters; and (3) increased transport of water-dispersing organisms (Babbitt and Tanner,
2000; Baber et al., 2002; Mulhouse and Galatowitsch, 2003). Biological connectivity, however, also can
decrease with drainage and ditching, as average distances between wetlands increase and limit the
ability of organisms to disperse between systems aerially or terrestrially (Leibowitz, 2003). Ground-
water withdrawal can have an effect similar to drainage on some wetlands, which can affect wetland
connectivity by reducing the number of wetlands. Of particular concern in the arid Southwest is that
ground-water withdrawal can decrease regional and local water tables, reducing or altogether
eliminating ground-water-dependent wetlands (Patten etal., 2008). Ground-water withdrawal,
however, also can increase connectivity in areas where that ground water is applied or consumed.
Particularly noteworthy is that restoration of hydrologic connectivity, especially in systems with
widespread human alterations, also might adversely affect downstream waters (Jackson and Pringle,
2010). For example, dam removal can result in the downstream transport of previously sequestered
pollutants (Jackson and Pringle, 2010); dam releases to restore flows, without simultaneous restoration
of sediment supplies, can result in downstream channel degradation (Germanoski and Ritter, 1988;
Schmidt and Wilcock, 2008). Hammersmark et al. (2008) used a modeling study to show how the
restoration of incised stream channels can improve connectivity between streams and floodplains and
thus restore predisturbance hydrology (i.e., increased floodplain water storage, reduced peak
stoonflow, and reduced baseflow).
2.4.5 Interactions Among Factors
Interactions among the factors discussed above can be complex. Here we provide an example of
temporary surface-water connections between wetlands in the prairie pothole region (PPR) to illustrate
these complex interactions (Leibowitz and Vining, 2003). Further details on wetlands in the PPR are
provided in Sections 5.4 and B.3.
During high-water conditions in 1995, a temporary surface-water connection was observed between
two geographically isolated prairie potholes in the region's Drift Prairie. Based on a spatial analysis
during similarly wet conditions in 1996, 28% of the wetlands in a 40 km2 area containing the sites had a
temporary surface-water connection to at least one other wetland, including a complex (defined in the
study as a group of wetlands interconnected through temporary surface-water connections) of 14
wetlands.
In considering these findings, Leibowitz and Vining (2003) suggested that precipitation and local relief
are the primary factors controlling the spatial distribution of these temporary surface connections.
Precipitation is the ultimate source of water that fills these wetlands, whereas relief controls how much
the water level in a wetland must rise before spillage occurs (water level is also influenced by
evapotranspiration and ground water, but ground-water dynamics are difficult to predict for individual
wetlands). Relief also controls mixing—which could occur in flatter areas when the boundaries of
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expanding wetlands overlap—by determining the change in surface area per change in water level. Thus,
for a given level of precipitation, the number of surface connections occurring between wetlands should
be inversely proportional to local relief. Within the PPR, precipitation generally decreases from east to
west, while relief generally increases. The easternmost physiographic region in the PPR is the Red River
Valley, a relatively flat ancient lakebed (Lake Agassiz) having deep deposits of silt and clay. Water can
pond easily on these deposits, producing shallow wetlands and integrated drainage (i.e., the presence of
stream networks). The Missouri Coteau, which forms the western boundary of the PPR, consists of dead-
ice glacial moraine. This area has hummocky terrain, and local relief can be as great as 15-45 m in
steeper areas (Winter et al., 1998). As a result, the Coteau has deeper wetlands and little to no integrated
drainage. The Drift Prairie, located between the Red River Valley and the Missouri Coteau, is an
undulating plain formed on ground moraine. Relief, wetland depth, and the level of integrated drainage
in the Drift Prairie are intermediate in comparison with the other two regions.
Leibowitz and Vining (2003) hypothesized that the combined effect of these patterns in precipitation
and relief should produce a strong east-west gradient across the PPR in the occurrence of intermittent
surface-water connections. Both the absolute number of connections and complex size (the number of
wetlands contained in a complex) should be highest in the Red River Valley. Given the relative flatness of
this area, mixing should be the more common mechanism for temporary connections. The number of
temporary connections and complex size should be lower in the Drift Prairie, and spillage might
dominate in this hillier terrain. In the Missouri Coteau, where relief is greatest, the occurrence of these
temporary connections should be rare and limited to small complex sizes. Human impacts, however,
could affect these trends (Section 2.4.4).
Beyond these regional trends in relief and precipitation, local variation in the occurrence of intermittent
surface-water connections should be influenced strongly by ground-water dynamics. The ground-water
hydrology of prairie potholes has been well investigated at several sites (e.g., Winter et al., 1998; Winter
and Rosenberry, 1998). The specific ground-water interactions—and hence the effects of ground-water
movement on spillage or mixing, however, are unknown for most prairie potholes. All else being equal,
ground-water discharge wetlands should receive more water, and so should have a higher probability of
spillage, than ground-water recharge wetlands because recharge should reduce the amount of water
available for spillage.
A major factor influencing the temporal distribution of intermittent connections within the PPR is wet-
dry cycles. Climatic changes that have occurred throughout the Holocene drive these cycles. Evidence,
for example, exists for 20-, 22-, 50-, 100-, and 200-year climatic cycles (Ashworth, 1999). Wetland
hydrology responds dramatically to these wet-dry cycles as ground-water levels and precipitation
patterns fluctuate. In 1996, the average monthly Palmer Hydrological Drought Index for central North
Dakota was 4.02 (88th percentile), compared with a median of 1.00 for annually calculated monthly
averages between 1895 and 2001. Moisture levels of this magnitude—and consequently the degree of
connectivity observed (Leibowitz and Vining, 2003)—would be expected to occur during wetter portions
of wet-dry cycles.
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2.4.6 Quantifying Connectivity
As previously discussed, watershed connectivity is a dynamic and scalable quantity that occurs along a
gradient from highly connected to highly isolated (Ward, 1989; Euliss etal., 2004). Connectivity can be
quantified using structural metrics of physical landscape features (e.g., watershed topography, the
spatial arrangement of habitat patches), or functional metrics of system dynamics, which integrate
information about processes and interactions that influence hydrologic flows or biological dispersal.
Selection of specific metrics for quantifying connectivity depends on the purpose of the assessment, the
environmental context (e.g., humid versus arid), type of connection (e.g., hydrologic, chemical,
biological), spatial and temporal scale of interest, and available data (Calabrese and Pagan, 2004;
Lexartza-Artza and Wainwright, 2009).
2.4.6.1 Hydrologic and Chemical Connectivity
In hydrology, connectivity research has aimed to understand how and when water volume inputs (e.g.,
precipitation minus water loss through infiltration, evaporation and transpiration) and moisture
thresholds trigger surface and subsurface flow, thereby influencing streamflows in a given watershed
(Western etal., 2001; Ali and Roy, 2010; Bracken etal., 2013). Because movement of water is the
primary mechanism by which chemical substances are transported downstream, quantifying chemical
connectivity is closely related to quantifying hydrologic connectivity (Michalzik et al., 2001; Borselli et
al., 2008). Hydrologic connectivity research has focused on relating patterns of soil moisture following
precipitation events to stream discharge (Western etal., 2001; James and Roulet, 2007; Ali and Roy,
2010) or measuring flow-process connectivity at the hillslope scale (Knudby and Carrera, 2005; Reaney,
2008; Smith et al., 2010). Although this research provides a critical understanding of how water moves
through a watershed, it is only indirectly related to connectivity between small streams and rivers, or
between wetlands and streams. Metrics for quantifying hydrologic connections between upstream and
downstream waters have started to be explored through research characterizing the hydrologic
permanence of streams (Fritz etal., 2008; Fritz etal., 2009) or mapping temporal variation in surface
connections between wetlands and streams using field (McDonough et al., 2015) or remotely sensed
data (Sass and Creed, 2008; Lang et al., 2012; Huang et al., 2014). More commonly, research efforts have
focused on data collection methods that could inform measurement of connectivity (e.g., deriving
relationships between connectivity and topography or water quality; hydrologic tracers; geostatistical
modeling; and watershed, ground-water, or coupled surface water-ground water modeling).
Structural indices derived from topography can be used to predict patterns of watershed wetness.
Examples include the Topographic Wetness Index (Quinn et al., 1995), which is quantified using the
upslope contributing area and local slope, as well as quasi-dynamic indices that calculate the effective
contributing area (variable source area) in a watershed (e.g., Barling et al., 1994; Tarboton, 1997;
Creed and Beall, 2009). These indices could be used to predict the location of hydrologic flowpaths and
areas of a watershed that might be efficient exporters of nutrients, sediment, or pollutants following
heavy rainfall or snowmelt periods (Creed and Beall, 2009; Lane etal., 2009). In flatter landscapes, a
more dynamic contributing area model is typically required (Shaw et al., 2013). One example is the fill-
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and-spill model in which the watershed contributing area expands when wetland storage reaches
capacity (fill) and wetlands overflow (spill) onto the land surface and into other surface-water bodies
(Tromp-van Meerveld and McDonnell, 2006; Shaw etal., 2013; McCauley and Anteau, 2014). Other
researchers have quantified basin-scale hydrologic connectivity as the ratio of flowing stream reaches
connected to the outlet to the total potentially flowing stream reaches (Phillips etal., 2011; Spence and
Phillips, 2015), or as transport potential in a given direction quantified by a directional connectivity
index (Larsen et al., 2012). Similar to this, the volume-to-breakthrough concept quantifies connectivity
as actual runoff relative to water inputs, where connectivity decreases with increased infiltration,
depression storage, slope length, barriers, or other factors (Bracken and Croke, 2007).
Several other lines of research are contributing to a general understanding of connectivity between
water features. Water quality indicators have been used to identify connectivity between wetlands and
streams (Johnston et al., 1990; Leibowitz and Vining, 2003). Tracer experiments using 15N, bromide, salt
solutions, fluorescing particles, or other conservative compounds have been conducted that can inform
flowpaths in aquatic systems (Mulholland et al., 2004; Bencala et al., 2011; O'Brien et al., 2012).
Modeling and measuring the mass transfer efficiency of a watershed using a parameter such as the
sediment delivery ratio, which describes and predicts the relationship between erosion and sediment
yield in a watershed, can indicate the degree of connectivity within a watershed (Atkinson, 1995; Hooke,
2003; Bracken and Croke, 2007). Geostatistical approaches are being developed to consider how
connectivity would be quantified within a branched stream network (Pagan, 2002; Ganio et al., 2005;
Peterson et al., 2007). In addition, numerous mechanistic modeling and simulation tools can be modified
and applied to investigate connectivity dynamics from geographically isolated wetland systems (Golden
et al., 2014) and headwaters (e.g., TOPMODEL; Beven and Kirkby, 1979) to downstream surface-water
systems.
Although the research community has not reached a consensus regarding the best methods or metrics to
quantify or predict hydrologic or chemical connectivity (Lexartza-Artza and Wainwright, 2009; Ali and
Roy, 2010; Bracken et al., 2013), future efforts to quantify connectivity using the descriptors discussed
in Chapter 1 (frequency, magnitude, duration, timing, and rate of change) or other connectivity metrics
will help to further refine and quantify the lines of research summarized above.
2.4.6.2 Biological Connectivity
In the quantification of biological connectivity, species traits (e.g., dispersal mode, habitat requirements,
behavior) also must be considered (Calabrese and Pagan, 2004). Structural connectivity can be
quantified from the physical landscape (e.g., the size, shape, and arrangement of habitat patches)
assuming that the spatial configuration of habitats reflects species' ability to move between them.
Functional connectivity directly incorporates information about species' movement obtained from field
studies or from models to inform estimates of connectivity (Calabrese and Pagan, 2004; Wainwright et
al.,2011).
Indices based on graph theory calculate connectivity using a graph to represent the landscape as a
network of nodes (e.g., habitat patches) connected by edges (pathways of movement; Urban and Keitt,
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2001). Such connectivity indices include the Minimum Spanning Tree (Urban and Keitt, 2001),
Correlation Length (Keitt et al., 1997; Rae et al., 2007), the Integral Index of Connectivity (Pascual-Hortal
and Saura, 2006), and the Probability of Connectivity (Saura and Pascual-Hortal, 2007). Graph-theory
approaches can be used to assess structural or functional connectivity at multiple spatial scales (Eros et
al., 2012). Specific information about habitats and focal species is incorporated by applying node
weights (e.g., habitat area or quality, population abundance), edge weights (e.g., Euclidean distance,
landscape resistance), or edge characteristics (e.g., direction of movement; Galpern et al., 2011). Indices
derived from such graphs seek to characterize connectivity in terms of habitat (e.g., total connected
habitat area), dispersal pathways (e.g., relative abundance of individuals using a pathway, path
redundancy or vulnerability), or both (Rayfield et al., 2010). The Integral Index of Connectivity, for
instance, incorporates patch area, the topological distance between patches and the proportion of
connected patches (Pascual-Hortal and Saura, 2006), and has been successfully used to quantify
connectivity within a river network at varying spatial scales for otters (Van Looy et al., 2013).
The dendritic nature of stream networks also can be explicitly integrated when considering the
biological connectivity for obligate aquatic species, such as fish. The branching structure of a dendritic
network (Figure 2-19B), which has a single pathway (the stream channel) between habitat patches (e.g.,
stream reaches), influences individual movement and population distribution and abundance, and thus
the impact of disturbances and fragmentation on connectivity (Pagan, 2002; Grantetal., 2007); this can
be reflected in graph-theoretic connectivity indices (Malvadkar et al., 2015). An example of a dendritic
metric is the Dendritic Connectivity Index, which uses the number of barriers (e.g., culverts) and the
passability of these barriers to quantify the probability that fish can move between two points in a river
network (Cote etal., 2009).
2.4.6.3 Summary
This section briefly reviews the growing body of research into testable indices and metrics that
represent hydrologic and biological connectivity of functional importance to downstream waters. Data
availability is a critical issue, as the information content that connectivity indices provide often is related
directly to their data requirements (Calabrese and Pagan, 2004; Bergsten and Zetterberg, 2013).
Additionally, the many proposed connectivity indices and approaches discussed in the literature suggest
that different metrics are needed to quantify different types of connectivity across diverse
environments, scales, and ecosystem functionalities (Rayfield et al., 2010; Galpern et al., 2011; Bracken
et al., 2013). With further development and refinement, the utilization of connectivity indices can
provide graphical, quantitative assessments of connectivity.
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3.1 Abstract
The physical structure of a river network inherently demonstrates cumulative connectivity
(Section 1.2.3) between all streams and their downstream rivers. Substantial evidence supports
physical, chemical, and biological connections from headwater streams—including those with
ephemeral, intermittent, and perennial flows—to waters immediately downstream through transport of
water and associated materials, movement of organisms and their products, and bidirectional
geomorphic adjustments. Among the most compelling evidence for the effects of headwater streams on
rivers is as sources of water, nitrogen, organic carbon, and sediment (clean and contaminated); as
transformers of and sinks for nitrogen, carbon, and contaminants; and as providers of essential habitat
for migratory animals such as anadromous salmon. Headwater streams as a class provide substantial
quantities of water to larger water bodies. For example, first-order streams cumulatively contribute
approximately 60% of the total mean annual flow to all northeastern U.S. streams and rivers. Infrequent,
high-magnitude events are especially important for transmitting materials from headwater streams in
most river networks. The strongest lines of evidence supporting the effects of headwater streams are
from watersheds where headwater streams drain a unique portion of the basin (e.g., hydrology, geology,
human alteration). Investigation of connections among river network components continues to be an
active area of scientific research, with progress occurring in the development of river network models
and connectivity metrics for quantifying connections and their downstream effects. Physical, chemical,
and biological connections between headwater streams and downstream waters are fundamental to the
structure and function of river networks, and additional empirical data and further breakthroughs that
quantify linkages across large spatiotemporal scales will continue to enhance our understanding of river
network complexity.
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3.2 Introduction
In this chapter, we describe the state of knowledge of stream connectivity and its effects on the physical,
chemical, and biological condition of downstream waters. Although we recognize that streams also are
important sources of water and other materials to nearby terrestrial systems and deep ground-water
systems via lateral and vertical connections (e.g., Gray, 1993; Shentsis and Rosenthal, 2003; Walters et
al., 2008), we focus here on longitudinal surface-water connections between streams and rivers, and on
shallow subsurface-water interactions integral to surface-water connections and downstream water
condition. The evidence primarily focuses on the connections between headwater streams and
downstream waters, but we draw some evidence from connections of larger streams to rivers,
reservoirs, lakes, and coastal waters. We consider the peer-reviewed evidence for connectivity and its
consequent effects on downstream rivers in terms of physical (Section 3.3), chemical (Section 3.4), and
biological (Section 3.5) connections between upstream and downstream habitats. Although we
recognize that many linkages between streams and downstream waters cross physical, chemical, and
biological boundaries, we have chosen this format for ease of presentation. In each section we also
consider how human alteration of streams affects their connectivity and resulting effects on
downstream rivers (Sections 1.2.4 and 2.4.4). In some cases, connectivity and its effects on downstream
waters become more discernable with human alteration (e.g., Chin and Gregory, 2001; Wigmosta and
Perkins, 2001); however, when human alterations are widespread and relatively uniform (e.g., Blann et
al., 2009) attributing downstream effects to particular tributaries or parts of the river network can be
more complex. Coupled human-natural systems are an area of active research (Box 3-1). Section 3.6
closes this general section with a discussion on stream-river connections by synthesizing evidence in
terms of the conceptual framework (Chapter 2) and viewing streams in a connectivity gradient context
(Section 1.2.2). In addition, two case studies on specific types of stream systems are in Appendix B:
prairie streams (Section B.4) and southwestern intermittent and ephemeral streams (Section B.5).
Streams range greatly in size in terms of both drainage area and discharge. In general, their abundance
is inversely related to their size. First-order streams typically are most abundant, although individually
they have the smallest drainage areas and shortest average stream lengths (Horton, 1945; Schumm,
1956; Ijjasz-Vasquez etal., 1993). When considering drainage area and stream length of headwater
streams together, however, they can represent most of the river watershed and network. Thus, despite
their small individual size, these headwater streams cumulatively can have a large influence on
downstream waters (Section 1.2.3).
Some headwater streams lack channel connections to large downstream water bodies because they
drain closed or endorheic basins. Endorheic basins have no surface outflows to oceans, but terminate as
inland lakes, seas, playas, or pans (Shaw and Bryant, 2011). Although endorheic streams are common in
some areas (Section B.5), endorheic basins represent only approximately 2% of the North American
continent (Vorb'smarty et al., 2000) and generate 0.15% (9 of 5,892 km3 yr1) of its annual discharge
(Fekete etal., 2002).
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Box 3-1. Urban Streams
Urban development alters the structure and function of stream ecosystems in numerous ways (Paul and
Meyer, 2001; Walsh et al., 2005). Although the specific symptoms of what Walsh etal. (2005) referred to as
the "urban stream syndrome" depend on numerous factors, including the location, density, type, and age of
urban development, common patterns have emerged. For example, urban streams typically experience
increased stormflows (from direct runoff to channels), flashier hydrographs, altered baseflows, increased
nutrient and contaminant concentrations, and decreased organic matter retention. Many of these attributes
are related to changes in connectivity—that is, alteration of the longitudinal, lateral, and vertical connections
between the landscape, headwater streams, and downstream waters.
Connectivity and consequences on downstream waters. One pervasive effect of urban development is the
alteration of hydrologic connectivity along river networks. The frequency, duration, magnitude, timing, and rate
of change of both stormflows and baseflows are altered via multiple pathways. The highly connected,
engineered network of impervious surfaces, pipes, and storm drains increases the volume and rapidity of
stormwater runoff into urban streams, resulting in increased frequency, magnitude, and rate of change of
stormflows within the river network. This quick delivery of stormwater runoff to streams also means that
stormflows tend to recede quickly, shorteningstormflow duration. Because impervious surfaces reduce
infiltration and watershed storage of water, urban development also can reduce baseflow magnitudes.
Together, these patterns result in the typical flashy hydrographs of urban streams and altered hydrologic
connections throughout urban river networks. Higher stormflow magnitudes and frequencies also can scour
sediments from urban channels, which, in combination with engineered channel straightening, can cause
urban channels to enlarge via incision and expansion. Direct
wastewater discharges to urban streams (e.g., from wastewater
treatment plants, industrial facilities, or combined sewer
overflows) and water withdrawals for municipal and industrial uses
also can affect the frequency, duration, and magnitude of
hydrologic connections in urban streams. Vertical hydrologic
connections can be augmented by leaky subsurface sewer and
water pipes, or diminished by reduced infiltration due to increased
impervious surface area and channel incision, straightening,
hardening, and simplification.
Stream burial, or the diversion of streams into pipes, culverts, and
other conveyances, is common in urban watersheds, and provides
another illustration of how urban development alters connections.
For example, more than 60% of all streams in Baltimore City,
particularly small headwater streams, have been buried (Elmore
and Kaushal, 2008). As a result, most lateral and longitudinal
connections along urban river networks have been replaced by
urban infrastructure, resulting in greatly expanded headwater
drainage areas (Kaushal and Belt, 2012).
These changes in hydrologic connectivity have significant
consequences for downstream waters in urban areas. Between
rain events, urban landscapes accumulate materials such as organic material, nutrients, and contaminants,
which then are delivered quickly to urban streams with surface runoff. As natural stream channels are
converted to simplified engineered structures, they lose their ability to retain and transform these materials,
resulting in reduced storage and lagtime before transport to downstream waters (Nedeau etal., 2003; Carey
and Migliaccio, 2009; Kaushal and Belt, 2012).
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Box 3-1. Urban Streams (continued)
Longitudinal connectivity in urban streams also influences the
movement and distribution of organisms in these systems.
Urban stream habitats frequently become fragmented and
homogenized, as connectivity is disrupted by road crossings,
channel incision, and other impacts associated with urban
development. Habitat homogenization reduces complexity,
which limits the availability of habitats needed throughout
species' life cycles (for reproduction, rearing, refuge from
disturbance and predation). Fragmentation can result in
reduced movement of mobile organisms, most notably fish,
through the river network (Perkin and Gido, 2012). Urban
streams also can benefit, however, from connectivity with
intact, upstream habitats. For example, Waits etal. (2008)
found that immigration from less disturbed upstream areas
serving as source habitats maintained central stoneroller
populations in an urban stream.
Connectivity and restoration of urban streams. Because so
many of the adverse effects associated with urban
development are related to changes in lateral, longitudinal,
and vertical connections along urban riverscapes, restoration
of these systems often involves re-establishing connections
that existed before urbanization. For example, detention ponds
and green infrastructure (rain gardens, bioswales, permeable
pavements, green roofs) are designed to slow stormwater runoff into urban streams, thereby increasing
retention and processing of water, nutrients, sediment, and contaminants. Ultimately, the slowing of
stormwater runoff can re-establish lateral and longitudinal connections as retention and transformation
pathways, rather than the primarily export pathway these connections traditionally served in urban river
networks.
The contribution of headwater streams to river networks in terms of stream number, length, or drainage
area over large geographic regions has been difficult to determine, even with advances in remote
sensing and geographic information systems. The small size of headwater streams makes distinguishing
them from surrounding areas and overlying tree canopies difficult in most regions (Gilvear and Bryant,
2003). Numerous studies have shown that existing U.S. hydrographic databases and topographic maps
underestimate the extent of headwater streams (Morisawa, 1957; Gregory, 1976; Hansen, 2001; Heine
etal., 2004; Stoddard etal., 2005; Roy etal., 2009). Therefore, most streams portrayed on databases and
maps as first-order streams are, when ground-truthed, second- or third-order streams. For example,
more than 80% of mapped (l:25,000-scale topographic maps) stream terminuses in a Massachusetts
watershed underestimated the upstream extent of the channels (Brooks and Colburn, 2011). On
average, these unmapped upstream segments were nearly 0.5 km in length, and 40% had one or more
upstream tributaries (Brooks and Colburn, 2011). Even with this widely known underestimation by
databases and maps, first-order streams recognized in the U.S. Geological Survey (USGS) medium-
resolution (l:100,000-scale) National Hydrography Dataset (NHD) represent 53% (2,900,000 km) of
total stream length (Nadeau and Rains, 2007). Moreover, approximately 50% of these first-order
streams were classified as not having year-round flow (i.e., nonperennial; Section 2.2.2; Nadeau and
Rains, 2007). Southwestern and prairie streams are predominantly ephemeral and intermittent
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(Sections 5.5, 5.6, B.4, and B.5). Thus, despite the shortcomings of existing national maps and
hydrographic databases, it is still clear that headwater streams—including ephemeral and intermittent
streams—represent a large fraction of river networks in the United States. Combining their
overwhelming extent with their high biogeochemical activity (Section 3.4) means that headwater
streams, including ephemeral and intermittent channels, have a large cumulative or aggregate effect on
the river network (Benstead and Leigh, 2012).
In the following sections, we consider longitudinal connectivity between streams and downstream
waters in terms of the physical, chemical, and biological connections between them.
3.3 Physical Connections
Physical connections result from the transport of nonliving materials that do not chemically change (or
change slowly) enroute from streams to downstream rivers. In this section, we discuss factors
controlling water, temperature (heat energy), sediment, and wood in streams; how these materials are
transported downstream; and evidence that these connections affect the condition of downstream
rivers.
3.3.1 Water
The recurrent, concentrated surface flow of water from surface runoff and ground water develops and
maintains river networks, and water is the primary medium carrying other materials from streams to
rivers (Section 2.3). The temporal dynamics of flow (its frequency, duration, magnitude, timing, and rate
of change) within and among river networks vary in space and time and influence the physical, chemical,
and biological connectivity between streams and downstream waters (Sections 2.2.2, 2.2.3, and 2.3.2.2).
Thus, the physical connection of water flow through river networks largely forms the foundation for
chemical and biological connections and where along the dynamic connectivity gradient streams are
positioned (Section 1.2.2).
Most rivers receive the majority of their water from tributaries rather than from direct precipitation on
or ground-water input to river segments (Winter, 2007; Bukaveckas, 2009). Alexander etal. (2007)
modeled flow through river networks in the northeastern United States and estimated that first-order
streams (designated on the l:100,000-scale NHD river network) provide approximately 70% of the
mean annual water volume in second-order streams and about 55% and 40% of the mean water volume
in fourth- and higher order rivers, respectively. Overall, first-order streams cumulatively contribute
about 60% of the total volume of mean annual flow to all northeastern streams (Alexander et al., 2007).
Headwater stream contributions to downstream baseflow vary among river networks, based on several
large-scale factors (Section 2.4). For example, headwater streams that have stronger connections to
ground water or that consistently receive more precipitation relative to downstream reaches have a
larger effect on downstream river baseflows. Hydrologic data from 11 nested gages distributed
throughout a watershed (176 km2) in the Catskill Mountains, NY were used to assess the extent of
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spatial correlation in baseflow discharge (Shaman et al., 2004). Baseflow discharge in smaller streams
(i.e., with watersheds <8 km2) was more weakly correlated with mainstem discharge than discharge in
larger streams; the authors concluded that this pattern reflected greater contributions by deep ground
water as drainage area increased (Shaman et al., 2004). Using geochemical tracers and hydrologic data
from 32 nested stations in a watershed (1,849 km2) of the River Dee in Scotland, Tetzlaff and Soulsby
(2008) determined that streams draining the upper 54% of the watershed contributed 71% of baseflow.
However, the upper watershed received only 58% of the total annual precipitation, indicating that long
residence time ground-water flowpaths from the headwater watersheds were also important in
maintaining downstream baseflows (Tetzlaff and Soulsby, 2008). In contrast, headwater streams
(0.11-3.5 km2) making up 33% of the total area in a northern Sweden watershed (78 km2) contributed
only 18% of the summer baseflow at the basin outlet (Temnerud et al., 2007). The specific discharge
contribution (L s-1 knv2) for headwater streams, however, varied by an order of magnitude (~0.5-8.0),
reflecting the heterogeneity (i.e., mires, lakes, forest) of the study watershed (Temnerud et al., 2007).
Jencso etal. (2009) monitored 24 transects with a total of 84 wells along lower hillslopes, toe-slope, and
riparian areas in a northern Rocky Mountains watershed (22.8 km2) and found that the duration of
connectivity from hillslopes to streams was positively correlated (r2 = 0.95) with the duration of higher
than normal downgradient watershed streamflow. This finding demonstrates the strong link between
downstream flow conditions and the connectivity of ephemeral and intermittent streamflow from
nearby hillslopes, and that the cumulative downstream effect of the hydrologic connections between the
hillslope and stream channel is time varying. Hydrologic connections to downstream rivers are often
complex, involving longitudinal, lateral, and vertical exchanges that vary over space and time. This
means that the flowpath by which headwater streams contribute to downstream waters will vary
according to climatic, topographic, and geologic context.
We can also infer the importance of headwater streams from variation in river hydrologic responses
over space. Discharge increases with drainage area, and the general assumption, particularly for mesic
environments, is that drainage area can be used as a proxy for discharge. The relationship can be written
as Q = kAc, where Q is discharge (m3 s'1), k is a constant representing hydrologic factors such as
antecedent moisture and precipitation, A is drainage area (km2), and c is the scaling power constant.
This scaling power reflects how the rate of discharge increases with drainage area, and can be useful for
qualitatively assessing headwater contributions to downstream discharge. Where c « 1, discharge is
generated proportionally with increasing drainage area. Where c < 1, upstream portions of the
watershed (where headwater streams tend to be most abundant) generate more discharge per unit area
than downstream portions, suggesting that rivers with c < 1 derive a higher proportion of their flow
from headwater streams. Where c> 1, downstream portions generate more discharge per area than
upstream reaches, suggesting that rivers with c> 1 might store more water per unit area in upstream vs.
downstream areas. Alternatively, urbanization in the lower portions of the watershed can lead to a
similar relationship (Galster et al., 2006). Data from multiple USGS gages along large, unregulated rivers
showed that mean and peak annual discharge do not always increase proportionally with drainage area
(Galster, 2007, 2009). Of the 40 rivers examined, only 16 had linear peak annual discharge-area
relationships (c« 1) throughout their period of record (Galster, 2009). Eleven rivers had relationships
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where c < 1, three rivers had relationships where c> 1, and ten showed changes in the relationship over
their period of record.
Despite variability in area-discharge relationships, most mesic watersheds have a value of c between 0.8
and 1 (Galster, 2007), suggesting that drainage area can be used to roughly estimate the proportion of
flow that arises from headwater streams. For example, Alexander et al. (2007) found that the
watersheds of first-order streams cumulatively accounted for 57% of the total drainage area and 55% of
the total annual river flow of the New England states. In more xeric arid and semiarid watersheds where
the ground-water table can be below the stream channel and thunderstorms of limited spatial extent
dominate runoff, however, c is generally < 1. For instance, in the highly instrumented Walnut Gulch
Experimental Watershed (operated by the U.S. Department of Agriculture, Agricultural Research
Services [USDA-ARS]) in southeastern Arizona, discharge becomes more nonlinear (c decreases) with
increasing watershed area, and a critical transition threshold area occurs roughly within 37-60 ha
(Goodrich etal., 1997). The primary causes of increasingly nonlinear response are (1) the increasing
role of ephemeral channel infiltration losses to the subsurface, unconsolidated alluvium, and (2) the
continual decline of fractional storm area coverage as watershed area increases. Caruso and Haynes
(2011) reported that first-order watersheds made up 61% of total drainage area of the Upper Colorado
River basin. In this case, the first-order streams produced a lower proportion (41%) of the total annual
river flow than suggested by their total drainage area, in part because 84% of the streams were
intermittent. Both studies used the l:100,000-scale NHD, in which first-order watersheds generally
correspond to second-order watersheds at the 1:24,000 scale (Alexander etal., 2007). These results,
representing two very different parts of the United States, strongly suggest that headwater streams,
even where seasonally dry, cumulatively generate a large fraction of the nation's stream and river flows.
The propagation of stormflow through river networks provides clear evidence of hydrologic
connectivity between headwater streams and rivers, particularly when an intense storm occurs over
only the headwater portions of a river network. In these cases, the hydrograph peaks sharply in the
headwater streams, indicating a quick response to precipitation (Figures 2-8 and 2-11). Timing of the
storm and onset of the peak are increasingly delayed with increasing distance down the network (Figure
2-11; see below for further discussion of hydrologic dispersion). Typically, discharge magnitude
increases as stormflow accumulates incrementally over the river network (Allan, 1995). The
contribution of tributaries to rivers during widespread floods manifests as stepped increases in
discharge immediately below confluences, as water flows accumulate through a river network
(Figure 3-1).
Such propagation was recorded following a monsoonal storm event through an arid network of
ephemeral channels in the Rio Grande, NM (Figure 3-2). The high-intensity storm dropped
approximately 18-25% of annual rainfall on the stream's approximately 16,000 km2 drainage area over
a 2-day period. Discharge recorded at two gages on the stream and three gages on the Rio Grande
downstream of the confluence illustrated lag (residence) time and peak hydrograph broadening at least
127 km downstream (Vivoni etal., 2006). Stormflow contributions from the ephemeral stream
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Figure 3-1. Longitudinal pattern of flow along (A) River Derwent and (B) River Trent, illustrating
stepped increases in flow associated with tributary inflows. Small arrows indicate location of
tributary confluences along the mainstem; bold arrow in (B) indicates the confluence of the two rivers.
Source: Reprinted from Fluvial Forms and Processes: A New Perspective, (1998) by Knighton with
permission of Routledge.
ISO
120
90
60
30
0
o
o
0 10 20 30 40 50 60 70 80 90 100 110
c
ro
-------
Figure 3-2. Time series of rainfall and streamflow observations in the Rio Puerco and Rio Grande,
6-18 September 2003. Source: Reprinted from Analysis of a monsoon flood event in an ephemeral
tributary and its downstream hydrologic effects, (2006) by Vivoni et al. with permission of John Wiley
&Sons.
IIT)
Mean Radar
Rainfall
Rio Puerco at
Bernardo
Rio Puerco at
Guadalupe
ario ovii an 2
ana 9/1 s ane 9ri7 ana
Rto Grande at
San Acacia
Rio Puerco at
Bernardo
Rio Grande at
San Marcial .
Rio Grande at
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-------
These areas, which mix with the main channel flows at relatively slow rates, are collectively part of the
stream's transient storage. As streamflow decreases after a storm, water that was temporarily stored in
the banks, the floodplain, and other off-channel habitats flows back into the channel and supports
stream baseflow (Sections 2.2.3 and 4.3.2.1; Whiting and Pomeranets, 1997; Chen and Chen, 2003;
Baillie etal., 2007). Hydrodynamic dispersion is also readily apparent when flow resumes in ephemeral
channels. The velocity at the front of flow moving down the dry channel is much slower than upgradient
of the front because of higher turbulence and infiltration losses at the front Flow in these situations,
piles up at the front and is reflected as rapid rises in the hydrographs (Figures B-6 and B-10).
Hyporheic flowpaths have been characterized for a variety of situations that affect streambed
topography and impede flow across a range of spatial scales (e.g., gravel bars, channel meanders, pool-
riffle sequences, and large woody debris; Buffington and Tonina, 2009; Stonedahl et al., 2010; Sawyer et
al., 2011) and in varying flow conditions that shift streambed topography (Harvey et al., 2012). The
residence time that water spends in the subsurface alluvium before upwelling into streams—that is, the
hyporheic residence time—is defined locally by the pressure head, alluvial volume, hydraulic
conductivity, bed stability, and near-bed turbulence. For example, because 90% of the stream length in
mountainous drainage basins is composed of steep channels with associated bed-form sequencing and
limited alluvial volumes, most hyporheic exchange in these systems is expected to be rapid, shallow, and
occur over small spatial scales (Buffington and Tonina, 2009). Slower, deeper, and longer hyporheic
flowpaths occur in streams in unconfined valleys, with moderate hydraulic gradients and extensive
alluvial volumes. In streams of both regions, hydrologic connections exist between shallow ground-
water sources and stream channels, but the characteristics of these connections differ. These differences
in hydrologic residence time are important, given that residence time reduces downstream flooding,
controls various biogeochemical processes, and influences the distribution of stream organisms
(Sections 3.4 and 3.5).
Geomorphologic dispersion is the cumulative effect of different travel distances over the larger spatial
scale of entire river networks (Rodriguez-Iturbe and Valdes, 1979; Gupta etal., 1980; Rinaldo etal.,
1991; Snell and Sivapalan, 1994). Not all points along the river network (or even headwater streams)
are equidistant from the network outlet, so water simultaneously entering different parts of the network
will not simultaneously arrive at the outlet.
Geomorphologic dispersion assumes water flowing through the network moves at a constant velocity or
has varying resistance to downgradient transport. Within river networks, however, water velocity and
related hydrodynamics change over space and time (e.g., channel slope and dimensions are not uniform
across all pathways through the river network; Saco and Kumar, 2002; Paik and Kumar, 2004).
Kinematic dispersion is the cumulative effect of spatially variable water velocity as it moves through
river networks (Saco and Kumar, 2002). The physical configuration and variable channel form of
streams within a river network, which influence components of hydrologic dispersion at varying scales,
are the primary controls dispersing flow from streams to rivers over time and thereby cumulatively
mediate the arrival time of stormwater pulses in rivers following rainstorms (Saco and Kumar, 2008).
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Another factor that influences hydrologic response is channel transmission loss, or the loss of surface-
flow volume due to infiltration into unconsolidated alluvium (Section 2.2). Transmission is another
process by which streams, particularly in arid and semiarid regions, can slow or divert from the
longitudinal flow of water to downstream rivers and thus minimize downstream flooding. Channel
transmission losses are readily apparent from a series of hydrographs recorded in the USDA-ARS
Walnut Gulch Experimental Watershed (Figure B-10). These hydrographs are the result of several high-
intensity thunderstorms in the upper and lower portions of the watershed. As little or no lateral
overland or tributary inflow occurs between the two upstream flumes, the decrease in both peak runoff
rate and runoff volume is the result of channel transmission losses and dispersion only. As illustrated in
this figure, however, even though runoff transmission losses are large there is sufficient runoff to
increase flow in the San Pedro River at the downstream Tombstone USGS gaging station. Over relatively
short time frames, infiltration or seepage through channel bed and banks typically dominates
transmission losses, although evapotranspiration losses can be significant in stream reaches with
prolonged surface flows (Hamilton et al., 2005; Costelloe et al., 2007). In many arid areas, precipitation
and the potential for runoff are highest in mountainous regions, where small, ephemeral streams are
most abundant (Section B.5). Because streams represent the topographic low points in watersheds that
collect and concentrate surface water, they tend to have more water available for infiltration, be more
permeable (have coarser sediment) than upland soils, have fewer plants, have higher antecedent
moisture, and be closer to shallow ground water—all of which are factors that increase the potential for
infiltration. In fact, evidence is mounting that ground-water recharge in hot arid and semiarid areas will
occur only where water is concentrated and focused, such as in channels, depressions, or areas of high
infiltration (e.g., karst; Brahana and Hollyday, 1988; Hughes and Sami, 1992; Sharma and Murthy, 1995;
Scanlon etal., 1997; Scott etal., 2000; Constantz etal., 2002; Goes and Pool, 2005). Infiltrated
precipitation in upland portions of alluvial drainage basins rarely reaches the ground-water table as
recharge due to high potential evapotranspiration, the adaptation of xeric plants to use available soil
moisture efficiently, and upward temperature gradients that transport water vapor upward in thick
vadose zones. Relative to their cumulative surface area, an inordinate amount of ground-water recharge
occurs in headwater ephemeral and intermittent channels within arid drainage basins (Osterkamp et al.,
1994; Goodrich etal., 2004).
Channel bed and bank permeability also governs the degree to which infiltration is an important
pathway between streams and ground-water aquifers. Fine bed and bank sediments slow infiltration. In
many semiarid and arid streams, bed sediments become finer in the downstream direction because flow
competence declines (Dunkerley, 1992), suggesting that lateral and vertical hydrologic connections
might be especially important in headwater streams. Sand and gravel mining in ephemeral and
intermittent channels and other human alterations that increase fine sediment loading and deposition
can further slow percolation (Bull and Scott, 1974). Because fine sediments can concentrate in channels
following moderate flows, higher flows that scour fine sediments or submerge more permeable
floodplains have higher infiltration rates (Lange, 2005). In the Walnut Gulch Experimental Watershed,
cumulative transmission losses over 54 km of channel resulted in a 57% decrease in flow volume
associated with a storm (Renard and Keppel, 1966). Infiltration losses accounted for up to half the flow
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volume along three ephemeral channels in the southwestern United States (Constantz et al., 2002).
Chemical and isotopic tracers have confirmed that ephemeral streams are cumulatively important areas
for floodwaters to recharge ground-water aquifers in desert regions (Tang et al., 2001). Although
transmission losses represent disruptions of surface connectivity between streams and downstream
waters, such losses indicate vertical hydrologic connections that reduce downstream flooding and
recharge the ground-water aquifers that eventually contribute to flow in downstream waters (Izbicki,
2007).
Human alterations designed to control the spatial and temporal distribution of water have affected the
longitudinal, lateral, vertical, and temporal dimensions of hydrologic connectivity in river networks.
Structures such as dams, weirs, levees, culverts, and pipes alter longitudinal transport, restrict lateral
expansion, and alter vertical exchange (e.g., Gregory, 2006; Hester and Doyle, 2008; Park et al., 2008).
Surface-water and ground-water abstraction and diversion can cause tributary segments to dry, thereby
severing longitudinal and vertical connectivity and reducing or eliminating lateral connectivity during
low-flow periods (e.g., Colvin and Moffitt, 2009; Scanlon et al., 2012). Human alterations that increase
fine sediment deposition or microbial biofilm in streambeds also can hamper vertical exchange (Battin
and Sengschmitt, 1999; Rehg et al., 2005), causing conditions that can become chronic without periodic
floods to flush out deposited sediments and biofilms (Box 3-1).
Human alterations also can affect the temporal dynamics of hydrologic connectivity in river networks. In
a predominantly rural river network in central Illinois, the total dispersion of the flow was controlled
primarily by geomorphological (~60%) and kinematic dispersion (~35%; Saco and Kumar, 2002). In
contrast, hydrodynamic dispersion cumulatively contributed to 72-86% of the total dispersion in highly
urbanized watersheds in the Chicago metropolitan area (Cantone and Schmidt, 2011). The rapid
hydrologic travel times associated with impervious surface runoff and rapid flow through the sewer and
storm drain networks contributed to the predominant influence of hydrodynamic dispersion (Cantone
and Schmidt, 2011).
Interbasin water transfer also affects the temporal and spatial dynamics of flow in human-dominated
river networks (Meador, 1996). Water is fundamental to human societies for drinking, food production,
industry, waste transport and processing, recreation, and aesthetics. Engineered infrastructure moves
water (and associated waste products) where and when it is needed (or removes it from where it is
unwanted). Many streams in human-dominated watersheds, particularly streams that historically have
ephemeral and intermittent flows, receive a significant proportion of their baseflow from municipal and
industrial waste water effluent discharges (Box 3-1). Streams that would be dry in the absence of these
discharges are called effluent-dependent streams, whereas those that receive most, but not all, of their
flow from effluent are called effluent-dominated streams (Brooks etal., 2006). About 25% of permitted
effluent discharges in the United States enter streams with mean annual flows incapable of diluting
effluents by more than 10-fold. This percentage of permitted effluent discharges entering streams
incapable of diluting effluents by more than 10-fold increases to 60% when low-flow discharge is
considered (Brooks et al., 2006). Streams draining human-dominated areas also can derive baseflow
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from ground water recharged by over-irrigation and leaky infrastructure (Lerner, 1986; Roach et al.,
2008; Townsend-Small et al., 2013).
Ultimately, these alterations can increase the frequency, duration, magnitude, and predictability of
baseflows when tributaries might otherwise contain little or no water. Because dry periods in
intermittent and ephemeral streams contribute to the key transformation, lag, and refuge functions
these systems perform (Sections 3.4 and 3.5), loss of these dry periods has consequences for
downstream waters. In addition, when water is stored or imported for human use, it is essentially being
"borrowed" from another period or location, which then must contend with reduced water availability.
Without careful water management and reuse (e.g., Bischel et al., 2013), any benefits of baseflow
augmentation can be overshadowed by potential risks, such as increased contaminant and pathogen
exposures (Section 3.4.4) and increased success of introduced species (Jackson and Pringle, 2010).
3.3.2 Sediment
Sediment carried with water flow from streams to downstream waters is critical for maintaining the
river network. Fluvial sediments scour channels, deposit to form channel features, and influence channel
hydrodynamics (Church, 2006). Although sediment is essential to river systems, excess sediment can
impair ecological integrity by filling interstitial spaces, reducing channel capacity, blocking sunlight
transmission through the water column, and increasing contaminant and nutrient concentrations (Wood
andArmitage, 1997).
Sediment in headwater streams originates from nearby hillslopes and enters these streams via overland
flow, bank erosion (Grimshaw and Lewin, 1980), and infrequent disturbances such as landslides and
debris flows (e.g., Benda and Dunne, 1987; Swanson etal., 1998; Eaton etal., 2003). Sediment
transported within river networks can be divided into two major categories: suspended and bedload.
Suspended sediment is fine sediment (clay, silt, fine sand) that requires slow velocities and little
turbulence to remain entrained in the water column; bedload sediment is coarser particles that slide,
roll, and bounce along the streambed during faster, more turbulent flows (Church, 2006; Wilcock et al.,
2009).
The dynamic balance between sediment supply and transport capacity (Lane, 1955; Bull, 1991; Trimble,
2010)—with the variables of sediment flux and sediment grain size on one side, and discharge and
channel slope on the other—is a principal paradigm of fluvial geomorphology. If one of these variables
changes, a compensatory change occurs in at least one of the other variables. For example, if discharge
increases, a lower channel slope is needed to transport the same amount of similarly sized sediment;
alternatively, less discharge or lower channel slope is needed to move a load of fine sediment than the
same load of coarse sediment. Associated with this balance is the relationship between channel
geometry (width and depth) and discharge (Leopold and Maddock, 1953), and adjustments to maintain
a dynamic balance also can include changes in channel dimensions.
The sediment supply-transport capacity balance is particularly relevant to geomorphologic connectivity
in river networks, because these variables typically differ as one moves from headwater streams to
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downstream rivers (Ferguson etal., 2006; Ferguson and Hoey, 2008). For example, slope and grain size
typically decrease, whereas discharge and channel size typically increase, in downstream reaches
(Church, 2002). Thus, streams cumulatively and aggregatively affect rivers in part by changing sediment
supply or transport capacity locally at confluences over time. Relatively small, local contributions in
sediment and discharge from a tributary stream might elicit no detectable change or only a short-lived
spike in downstream sediment characteristics, discharge, or channel geometry. In contrast, tributary
streams making large relative contributions at mainstem confluences elicit strong, stepped changes in
mainstem characteristics. Because headwater streams can make large contributions during infrequent
disturbances (e.g., floods, debris flows), the influence of headwater streams on downstream waters can
vary significantly over time, and even headwater streams can have long-lasting effects on rivers.
Human alterations can exert considerable influence on the structure and distribution of a watershed's
river network, thereby affecting sediment-based connections between headwater streams and
downstream waters. For example, road building in steep forested areas in the U.S. Pacific Northwest can
cause soil erosion, create concentrated discharge, and increase stream channel network lengths, all of
which affect the spatial distribution, intensity, and timing of erosional processes and cumulative
sediment delivery to downstream waters (Montgomery, 1994; Wemple et al., 1996; Wemple et al.,
2001).
Dams also modify sediment dynamics within river networks. Sediment concentrations and suspended
loads can be reduced for hundreds of kilometers downstream of dams, as is especially apparent in the
semiarid and arid western U.S. river networks (Williams and Wolman, 1984). The disruption of
downstream sediment supply by dams alters the balance between sediment supply and transport
capacity (Williams and Wolman, 1984; Kondolf, 1997). Water released from dams lacks sediment load
and thus has excess energy. This energy often downcuts channels downstream of dams, causing channel
incision and streambed coarsening as finer gravels and sands are transported downstream over time
(Williams and Wolman, 1984; Kondolf, 1997). The elimination of floods enables the encroachment of
terrestrial vegetation, resulting in channel narrowing and the conversion of complex, multithreaded
channels into simple, single-thread channels.
Other human activities also can affect sediment dynamics. Gravel and sand mining locally removes bed
sediment and lowers streambed elevation, creating a steep gradient change. Erosion of the streambed
can occur both upstream and downstream of the mine. The steep gradient change increases stream
power locally, which increases sediment demand and causes the streambed to erode in the upstream
direction via headcutting, which often extends far up into tributary channels (e.g., Florsheim et al., 2001;
Rinaldi et al., 2005; Rieke-Zapp and Nichols, 2011). Erosion in the downstream direction occurs because
most of the sediment being carried by water is deposited in the mining pit, leaving the water that passes
over the pit with excess energy that subsequently leads to downstream channel downcutting (Bull and
Scott, 1974; Kondolf, 1997). These examples show that the dynamic balance between sediment supply
and transport capacity represents a fundamental longitudinal connection along the river network that
must be considered to determine the potential repercussions of human alterations.
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Streams transport and store sediment. Headwater streams tend to have low competence to transport
sediment during baseflow (Gooderham etal., 2007), but they have structures (boulders, woody debris)
that entrain and store colluvial sediments between infrequent disturbances (e.g., stoonflows) that are
the dominant means for downstream sediment transport (e.g., Gomi and Sidle, 2003). Because of their
abundance and distribution, headwater streams can have a substantial cumulative effect on downstream
waters via sediment storage and transport. Poor soil conservation, drainage of wetlands, deforestation,
and tributary channelization associated with the development of agricultural land has long been
recognized as being detrimental to downstream waters via their connections with headwater streams
(Person et al., 1936). To stem further degradation, government agencies encouraged and funded various
soil conservation practices and the construction of small impoundments on headwater streams to trap
sediment and provide stable water supplies for livestock, irrigation, and recreation (Person et al., 1936;
Renwick et al., 2005). Although most such ponds are small (<1 ha or 2.5 acre) and represent only ~20%
of the total impounded area (or 0.4% of the total watershed area), they can cumulatively have a
significant effect. For example, Smith and Kraft (2005) estimated that the approximately 2.3 million
ponds distributed primarily on headwater streams of the Mississippi River network cumulatively
captured 25-50% of the eroded soil from the landscape.
Ephemeral desert streams are another example of sediment connections between headwater streams
and downstream waters. These ephemeral streams can exhibit high sediment export efficiency by
having higher bedload per unit stream power than that of forested perennial streams (Laronne and
Reid, 1993). Despite infrequent flows of short duration, flood waves (bores) in ephemeral desert
streams can carry substantial amounts of sediment downstream (Hassan, 1990). The transport distance
associated with these floods, however, often is insufficient to link them directly to perennial rivers. For
example, a reach-scale study in the Walnut Gulch Experimental Watershed in Arizona estimated sand
transport distances of only 401 and 734 m in nine floods over two consecutive years (Powell et al.,
2007). Over longer times spans the episodic nature of flow in ephemeral and intermittent channels
transfers sediment in a stepwise manner, depositing sediment some distance downstream and then
moving it farther downstream by subsequent events. The frequency, timing, and predictability of stream
runoff and therefore sediment transport vary widely with significant seasonal, annual, and interannual
variations that depend on elevation, climate, channel substrate, geology and the presence of shallow
ground water. Over longer time spans, however, sediment will continue to move downstream and affect
downstream waters (Brooks and Lemon, 2007).
Despite increasing bank erosion rates with increasing channel size and discharge, sediment yield from
watersheds typically decreases with increasing drainage area, due to increased sediment deposition
within channels and on nearby floodplains (Walling, 1983). This storage of sediment contributes to the
temporal attenuation or lag in the sediment delivery to downstream waters; it also illustrates that
headwater streams are important sediment sources for maintaining channels and floodplains.
Streams also can store substantial amounts of sediment that are released only during rare export events.
A series of experimental sediment introductions into steep, ephemeral, second-order streams in
southwestern Washington showed that between 30 and 45% of the added sediment (ranging from clay
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to coarse sand) was exported to the mainstem 95-125 m downstream, during stormflows of 66-69% of
bank full discharge (Duncan et al., 1987). Virtually all the added fine clay particles were exported from
the ephemeral streams to the mainstem, presumably because this fraction remained suspended at even
moderate flows (Duncan etal., 1987). Headwater streams within an Oregon Coastal Range watershed
(2.5 km2 area) stored 23% of total stored sediment within the watershed's river and valley network,
compared with only 9% storage within the mainstem channel (May and Gresswell, 2003). Trimble
(1999) constructed a long-term sediment budget for the Coon Creek watershed (360 km2), a Wisconsin
stream in the Mississippi River drainage, over periods coinciding with major land-use changes. When
agricultural practices caused major soil erosion (1853-1938), streams acted as net sources of sediment
(42 x 103 Mgyr-1); after erosion control, streambank stabilization, and revegetation (1975-1993),
streams became net sediment sinks (9 x 103 Mgyr-1) (Trimble, 1999).
Several studies identify abrupt changes in sediment size and channel morphology that coincide with
stream confluences having sufficiently high symmetry ratios (Knighton, 1980; Rhoads, 1987; Rice and
Church, 1998; Rice et al., 2001). Reviews of tributary confluence data have identified that symmetry
ratios ranging from 0.2 to 0.7 are needed to create a discernible sediment or channel morphology
discontinuity along a mainstem (Rhoads, 1987; Benda, 2008). Suspended particulate matter (inorganic +
organic) and bed particle size were measured above and below eight confluences on the Acheron River
in Australia to determine stream contributions (Wallis et al., 2008; Wallis et al., 2009). Suspended
particulate matter downstream of confluences approximated the sum of mainstem and stream exports
during high flows, but stream contributions were negligible during low flows (Wallis et al., 2009). Four
of the eight confluences showed expected changes in bed particle size below confluences with streams,
but bed particle sizes were similar in the mainstem and stream for the remaining confluences (Wallis et
al., 2008).
Streams, through their connections to rivers at confluences, can disrupt longitudinal trends in discharge
of water and sediment in rivers (Best, 1988; Benda etal., 2004; Ribeiro etal., 2012). For example, dams
often remove much of the sediment from transport, whereas most streams naturally are sediment
sources. The objective of a study on the Agigawa River in Japan was to examine contrasting disruptions
associated with a dam (sediment removal) and a stream confluence (sediment discharge) downstream
of the dam (Katano et al., 2009). Stream sediment contributions to the river reversed many of the dam-
related changes to downstream waters, including restoration of turbidity levels and the proportion of
sand and gravel substrate in the river bed (Katano et al., 2009). Other upstream land uses can also have
an effect on downstream sediment transport. Numerous modeling studies have shown how land use can
affect sediment export from headwater streams to downstream waters. For example, Howarth et al.
(1991) used the Generalized Watershed Loading Function model in the Hudson River estuary and its
associated watershed and demonstrated that urban, suburban, and agricultural land uses in headwater
watersheds produced the highest proportion of downstream sediment and organic carbon delivery to
the estuary. More recently, Wilson and Weng (2011) applied the Soil and Water Assessment Tool in the
Des Plaines River watershed in Illinois to simulate the cumulative effects of headwater streams on
downstream total suspended solids concentrations. Their calibrated model projected that expansion of
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medium- and high-density residential development in place of low-density residential development in
headwater subwatersheds would decrease downstream total suspended solid concentrations.
3.3.3 Wood
Large woody debris (typically considered >10 cm diameter and >1 m long) has a strong influence on
hydrodynamics, sediment transport and storage, and channel morphology (e.g., Harmon et al., 1986;
Nakamura and Swanson, 1993; Abbe and Montgomery, 1996; Naiman and Decamps, 1997; Montgomery
et al., 2003). Woody debris dissipates energy, traps moving material, and forms habitat for aquatic
plants and animals (Anderson and Sedell, 1979; Harmon etal., 1986; Abbe and Montgomery, 1996;
Naiman and Decamps, 1997; Gurnell et al., 2002). In-channel wood can redirect water movements,
create pools, and slow water movement through a channel (Nakamura and Swanson, 1993; Abbe and
Montgomery, 1996; Naiman and Decamps, 1997). Wood recruitment to forested streams occurs because
of chronic tree mortality; episodic disturbances such as fire, debris flows, landslides, and windthrow;
and bank erosion. The steeper topography associated with hillslopes along many headwater streams
increases the likelihood that trees will fall toward the channel (Sobota et al., 2006), relative to streams in
flatter terrain. Environmental setting, including valley slope, influences the supply of wood to streams
and therefore the degree of connectivity between streams and downstream waters.
Wood tends to accumulate in, rather than be exported from, most forested headwater streams, due to
their low discharge and relatively narrow channel widths (Keller and Swanson, 1979; Bilby and Ward,
1989; Gurnell, 2003). For example, wood was determined to have entered the channel more than 60
years earlier in a North Carolina headwater stream (Wallace et al., 2001); in some Pacific Northwest
streams, wood entered the channel more than a century earlier (Swanson and Bachmann, 1976; Keller
et al., 1981). Because of the large occurrence of wood and small size of streams, wood has a stronger
influence on hydrologic and geomorphic processes in headwater streams than in most larger rivers
(Bilby and Bisson, 1998).
Large, infrequent disturbance events are the primary drivers for wood movement from headwater
streams (Benda and Cundy, 1990; Benda et al., 2005; Bigelow et al., 2007). Reeves et al. (2003)
determined that 65% of the wood pieces and 46% of the wood volume in a fourth-order stream in
Oregon's Coastal Range were delivered downstream from headwater streams by debris flows, rather
than originating from the riparian zone next to the fourth-order channel. Using data from 131 reservoirs
in Japan, investigators identified a curvilinear relationship between watershed area and large woody
debris export (Seo et al., 2008); wood export per unit area increased with stream size for headwater
streams (6-20 km2), peaked at intermediate-sized streams (20-100 km2), and then decreased with
stream size for large streams (100-2,370 km2). The amount of wood in low-gradient midwestern
streams was determined to be supply limited mainly because human alteration both depletes large
wood sources and results in altered hydrology and channel structure enhancing downstream transport
of small wood (Johnson etal., 2006). Topography and topology also govern wood delivery from
headwater streams. Downstream segments draining steep, finely dendritic networks receive a greater
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proportion of wood from headwater streams than networks that are low gradient and weakly dissected
(Benda and Cundy, 1990; Reeves etal., 2003).
Additional evidence on wood-mediated connections along the river network comes from studies of
wood upstream and downstream of tributary confluences. Several studies have assessed the
distribution of wood associated with confluences. Wood volumes were measured upstream and
downstream of 13 confluences (symmetry ratios ranged from 0.05 to 0.49) in the Cascade Range of
western Washington (Kiffney et al., 2006). Wood volumes tended to peak at or immediately downstream
of stream confluences (Kiffney et al., 2006), suggesting that streams are either important sources of
wood to mainstems or alter channel form to enhance wood storage at confluences. Elevated wood
density, however, was not associated with confluences of eight streams to the Acheron River in Australia
(Wallis etal., 2009). The authors concluded that the study streams had insufficient capacity to transport
wood to the mainstem, because streams had similar slope to the mainstem but lower discharges (Wallis
etal., 2009).
Large wood can shorten sediment transport distances and debris flow runout by entrainment
(Lancaster et al., 2003). Woody debris in 13 Coastal Range streams in Oregon had accumulation rates
ranging from 0.003 to 0.03 m3 nr1 yr-1, largely based on time since the last debris flow (May and
Gresswell, 2003). The volume of instream wood was strongly related to the volume of sediment stored.
On average, 73% of stream sediment, prone to debris flow transport, was stored behind instream wood
(May and Gresswell, 2003). Unlike most human-built dams, natural logjams and beaver dams are
temporary structures that do not completely restrict transport of water, sediment, and biology across all
discharge levels. Although natural wood accumulations act to restrict longitudinal connectivity by
slowing the downstream transport, these features enhance lateral and vertical connectivity with the
floodplain andhyporheic zone, respectively (Burchsted etal., 2010; Sawyer etal., 2011). The importance
of wood in decreasing longitudinal connectivity, while enhancing lateral connectivity, temporary
storage, and habitat diversity has been documented not only locally at unit and reach spatial scales (1-
100m stream length) but along entire networks where valley confinement is an important predictor for
wood storage (Wohl and Beckman, 2014). Past and ongoing human activities (timber harvest, beaver
trapping, road building along streams, placer mining, log floating, desnagging) have so completely
removed in-channel wood and availability of near-channel old-growth wood recruitment, that retention
of new wood in channels is unlikely (Wohl and Beckman, 2014). Wood (and associated sediment)
movement from headwater streams to downstream segments occurs through infrequent, high-
magnitude events (e.g., debris flows, fire). Once in larger streams, wood and sediment can be stored in
alluvial fans and floodplains between stormflows that trigger additional downstream movement
through the network (Benda et al., 2005). Because of the long distances and infrequent triggers
associated with wood transport from most headwater streams to rivers, the relevant periods for
governing transport aggregate over decades to centuries (Benda et al., 1998). Wood entering headwater
streams can affect the downstream transport of water and materials in headwater streams, but also can
be transported downstream from headwater streams where it is important habitat for aquatic life, a
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source of dissolved and particulate organic matter, and influential in controlling hydrodynamics and
channel morphology of rivers.
3.3.4 Temperature (Heat Energy)
Connections between streams and downstream waters can affect heat transfer, and thus water
temperature, throughout river networks (Knispel and Castella, 2003; Rice etal., 2008). Heat is thermal
energy transferred across a boundary, whereas temperature is the amount of thermal energy per unit
volume (Coutant, 1999; Poole and Berman, 2001). Therefore, the amount of heat and the size of the
water body (i.e., volume, discharge) are fundamental controls of water temperature. Because water
temperature is such a fundamental property that drives physical (e.g., viscosity and density of water),
biological (e.g., organism behavior and physiology), and biogeochemical (e.g., nutrient assimilation and
mineralization) characteristics of stream ecosystems, it can cumulatively have significant indirect effects
on downstream waters via its effects on other forms of connectivity. This influence can occur over even
relatively small spatial scales or patches (Sections 3.4 and 4.5; Allan, 1995). For example, water
temperature strongly regulates stream ecosystem respiration, which then drives nutrient uptake
(Section 3.4.1; Demars et al., 2011). Warmer temperatures exacerbates eutrophication problems such as
fish kills, and heat stress can interact with chemicals synergistically or antagonistically making them
more or less toxic to organisms, respectively (e.g., Holmstrup et al., 2010).
The total net heat exchange for a stream has several components, including heat flux from solar
radiation, evaporation, convection with air, conduction with the streambed sediments, and advection
with direct inputs from precipitation, ground water, tributaries, and effluents (Webb, 1996; Coutant,
1999). Given these diverse thermal energy fluxes, numerous direct and indirect factors can change
stream temperature. For instance, riparian vegetation directly affects stream temperature by insulation
(shading incoming solar radiation and trapping air, reducing wind; Moore et al., 2005) and indirectly
affects stream temperature via its influence on channel morphology (e.g., Trimble, 1997) and degree of
hyporheic exchange through input of woody debris (e.g., Sawyer et al., 2012). Channel morphology can
directly influence stream temperature by affecting bank shading and altering channel width-to-depth
ratio, and indirectly influence stream temperature by affecting hyporheic exchange. Hyporheic exchange
influences stream temperature via buffering (reducing the diel temperature range) and lagging
(offsetting daily temperature patterns relative to surface-water patterns) effects, due to the extended
alluvial flowpath and by the advection or conduction of thermal energy or both (Arrigoni et al., 2008).
Over coarse spatial scales, a nonlinear increase in mean daily water temperature typically occurs from
headwater streams to large rivers (Caissie, 2006). A unimodal trend occurs in daily variation (i.e., daily
maximum-minimum) of water temperature, as stable ground-water temperatures (in headwater
streams) and greater depth and volume of water (in large rivers) buffer water temperatures from the
daily changes typical in intermediate-sized streams (Caissie, 2006). The steep increase in water
temperature immediately downstream of headwater streams is associated with more rapid flux of heat
into headwater streams, as shallow water contacts the surrounding air and receives direct radiation
(Caissie, 2006). This longitudinal pattern, however, does not hold for all river networks: Some river
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networks receive substantial deep ground-water contributions at lower reaches or advective inputs
from tributaries along the mainstem. Channel network configurations can influence the length,
dominant aspect relative to the sun, and distribution of tributaries, which influence the thermal
heterogeneity along a stream that might be associated with inflowing surface and hyporheic water.
Callahan etal. (2015) illustrated how topographic, geomorphic, riparian, and hyporheic exchange can
interact to influence stream temperature in the Kenai Peninsula, AK. Ground-water inputs played
important moderating roles in determining stream temperatures in both low-gradient versus steep
headwater streams, despite these streams having different channel morphologies, draining contrasting
topographies, and having different riparian canopies (Callahan et al., 2015). Although low-gradient
headwater streams had fewer channel margin seeps and lower hyporheic exchange than the steep
headwater streams, the subsurface-water temperature entering the low-gradient streams was lower
during summer than that entering the steep streams (Callahan et al., 2015).
Although many studies have determined that several direct and indirect factors can alter stream
temperature, including those listed above, these effects typically have been documented to carry for only
short distances downstream. This is in part because most studies measuring stream temperature
changes are conducted over reach or subreach scales (<100 m) and because stream-water temperature
equilibrates rapidly (~4 hr) to immediate surrounding conditions (e.g., Zwieniecki and Newton, 1999;
Rutherford et al., 2004; Hester et al., 2009). Some studies, however, do provide evidence of thermal
connections along river networks. The empirical evidence supporting thermal connections between
headwater streams and downstream waters includes studies that have gauged the spatial relationship of
water temperature over river networks and studies that have detected discontinuities in river
temperature associated with stream confluences. Geospatial analyses are used to assess the degree of
spatial dependence of a variable across a river network, and are particularly well suited for studying
connectivity within these systems. Studies of this type have shown that upstream water temperature is
significantly related to downstream water temperature, even over relatively long distances. For
example, water temperature data collected at 72 locations throughout a Catskill Mountain, NY
watershed were used to predict daily mean summer water temperatures spatially throughout
approximately 160 km of channel (Gardner and Sullivan, 2004). Results showed that water
temperatures at points along the river network separated by up to nearly 20 km were related. Johnson
et al. (2010) similarly used geostatistical analyses to determine the influence of headwater streams on
downstream physicochemistry, including water temperature. Water temperature within the eastern
Kentucky watershed was correlated across the river network over an average distance of approximately
5 km (Johnson etal., 2010).
Studies that have detected discontinuities in river temperature associated with stream confluences also
provide evidence of thermal connections along river networks. Ebersole etal. (2003) identified and
characterized cold patches along a river network in northeastern Oregon that largely had summer water
temperatures exceeding the tolerance limit of native salmonids. Floodplain springbrook streams were
among the cold patches identified and were determined to contribute the coldest water to the river
network (Ebersole etal., 2003). A subsequent study in northeastern Oregon determined that tributary
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confluences typically provided coldwater (>3 °C colder than mainstem temperatures) patches during the
summer (Ebersole et al., 2015). In addition, 39% of these tributary confluences were with streams that
contributed cold hyporheic water even when they lacked surface water—that is, they were ephemeral
and intermittent streams that were significantly connected to downstream waters even when the
streambed surfaces were dry (Ebersole et al., 2015). Unexpectedly, factors such as tributary size, flow
presence, and flowpath length were not important in predicting whether a tributary's confluence would
be a cold patch. Rather, the probability of a confluence's being a cold patch was largely explained by
amount of available water at the end of the snowmelt season (Ebersole etal., 2015).
Thermal infrared sensors are a recent remote-sensing tool that can provide snapshots of thermal
heterogeneity along river corridors (Torgersen et al., 2001; Torgersen et al., 2008; Cristea and Burges,
2009). Thermal maps and plots of longitudinal profiles overlaid by stream locations show that
confluences coincide with distinct peaks and troughs in river temperature (Figure 3-3). The effects of
streams were discernible when temperature differences of streams and the mainstem exceeded 1 °C and
streams had large symmetry ratios (Cristea and Burges, 2009). In most cases, the effect of the stream on
river-water temperature was minor relative to longitudinal changes over the course of the river
(Torgersen et al., 2001; Cristea and Burges, 2009). Despite having a relatively minor effect on
temperature over the length of entire rivers, however, streams provide persistent coldwater habitats
that are less susceptible to meteorological variation than other classes of thermal refuges and therefore
are particularly important for aquatic life (Section 3.5.2; Dugdale et al., 2013).
Although headwater stream temperatures are highly responsive to local conditions, they still can have a
cumulative effect on downstream waters. The fact that large-scale alteration of headwater streams has
been documented to affect downstream water temperature illustrates this point. For example,
reductions in baseflow (ground-water inputs) resulting from increased surface runoff from impervious
surfaces (Leopold, 1968) and reduced hyporheic exchange through the engineered piping, straightening,
and hardening of streambeds contribute to increased average and maximum summer water
temperatures and decreased average and minimum winter temperatures in downstream waters. The
combination of riparian vegetation removal, increased urban runoff, and storm sewer inputs results in
larger temperature swings associated with increased channel width-to-depth ratios and thus air-water
surface area available for radiant, evaporative, and convective fluxes (LeBlanc etal., 1997).
3.4 Chemical Connections
Chemical connections are linkages between headwater and other tributary streams and their
downstream waters based on the transport of chemical elements and compounds (e.g., nutrients,
dissolved and particulate organic matter, ions, and contaminants). Chemical connectivity between
streams and rivers involves the transformation, removal, and transport of these substances throughout
the river network; these processes, in turn, influence water quality, sediment deposition, nutrient
availability, and biotic functions in rivers.
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Figure 3-3. Airborne thermal infrared remotely sensed water temperature in the mainstem and at
tributary confluences of the North Fork John Day River, OR, on 4 August 1998. Line indicates
mainstem, black dots indicate tributary confluences, and dashed vertical lines indicate location of
tributary confluences along the mainstem. Reprinted with permission from Torgersen et al. (2008).
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Because water flow is the primary mechanism for downstream transport of chemical substances,
chemical connectivity is closely related to hydrologic connectivity (Sections 2.2 and 3.3.1). The
movement of water across and through landscapes and into river networks integrates potential solute
sources and sinks throughout the watershed. Thus, solute concentrations are an integration of upstream
mixing processes and transport processes in the stream channel. In simplest terms, streams generally
operate in two modes: a high-discharge throughput mode in which solutes and particles entering the
stream channel are quickly transported downstream, and a low-discharge processing mode whereby
solutes and particles are processed or stored near where they entered the river network (Meyer and
Likens, 1979).
Factors that affect hydrologic connectivity (including precipitation patterns and human alterations)
modify these upstream-downstream chemical linkages. For example, the spatial and temporal variability
of rainfall affects chemical connectivity between streams and rivers. Many headwater streams receive
pulsed inputs of water, sediment, organic matter, and other materials during rain events. Periodic flows
in ephemeral or intermittent streams can have a strong influence on biogeochemistry by connecting the
channel to other landscape elements (Valett et al., 2005), and this episodic connection can transmit
substantial amounts of material into downstream rivers (Nadeau and Rains, 2007).
The alternation of dry and flowing periods largely drives the temporal dynamics of chemical
connections between ephemeral and intermittent streams and downstream waters. The frequency,
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duration, magnitude, timing, and rate of change of flow further account for the variable connectivity
observed within and across river networks over space and time (Section 1.2.2). Materials accumulate on
and within dry streambeds where they are temporarily stored and can undergo transformations (Acuna
etal., 2005; Fritz etal., 2006a; Ademollo etal., 2011; Arce etal., 2014). Transmission losses, tributary
confluences, various channel forms, and retention structures also can contribute to the spatial
distribution of materials and processes in dry streambeds (Marcus, 1987; Graf etal., 1991; Reneau etal.,
2004; Taylor and Little, 2013). The onset of flows in ephemeral and intermittent stream channels,
particularly those following long dry periods and initiated by floods (i.e., first flushes), are important in
transporting and transforming large amounts of unique materials for long distances downstream, which
then can have significant effects (e.g., Obermann etal., 2009; Hladyz etal., 2011; David etal., 2012).
Human alteration of channel characteristics (e.g., channel shape and depth) and organic matter inputs
also affect the ability of streams to temporarily store and cycle materials before transport to
downstream waters.
Biogeochemical transformations control the mobility of different chemicals by altering chemical
properties, such as form (e.g., dissolved, colloidal, gravitoidal), bioavailability, and toxicity. Thus,
transformation is a key process influencing the downstream transport and attenuation of chemicals.
Physicochemical (e.g., pH, redox potential, chelator concentration, light, hydrologic residence time) and
biological (e.g., extracellular enzymes, physiology, lipid content) conditions control the location, rate,
and timing of chemical transformations in streams and downstream rivers. For example, the
introduction of stream restoration structures (e.g., small log dams) can affect the spatial distribution of
oxic and anoxic zones in streambeds and thus biogeochemical cycling and reaction rates for instream
biogeochemical processes throughout the river network (Lautz and Fanelli, 2008). These types of
human alterations, in turn, affect the form of chemical substances and the timing of their transport
downstream (Box 3-1). Data from the Baltimore Ecosystem Study Long-Term Ecological Research site
suggest that increased hydrologic connectivity from urban infrastructure (e.g., pipes, storm drains,
ditches) in headwaters increases the frequencies of occurrence and transport rates of nutrients, carbon,
and metals to downstream surface waters (Kaushal and Belt, 2012). Urbanization can cause complex
downstream responses, however, and sometimes creates longer travel times (i.e., reduced downstream
connections). For example, aging infrastructure can leak water and pollutants into ground water rather
than transporting these materials directly downstream.
3.4.1 Nutrients
Studies have documented nutrient-based chemical connections along river networks. Alexander et al.
(2007) investigated how stream size affected nitrogen transport in a northeastern U.S. river network.
First-order headwater streams contributed approximately 65% of the nitrogen mass in second-order
streams, and approximately 40% of that mass in fourth-order and higher order streams (Alexander et
al., 2007). Alexander et al. (2000) conducted a study of major regional watersheds of the Mississippi
River basin, which showed that instream nitrogen loss was inversely related to mean stream depth. This
finding most likely resulted from the reduced occurrence of denitrification and settling of particulate
nitrogen in deeper channels, due to reduced contact and exchange between stream water and benthic
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sediments (Alexander etal., 2000). Bb'hlke etal. (2009) used laboratory-, local-, and reach-scale studies
to describe the effect of seasonal and event-based variation of instream properties (e.g., stream depth,
flow rates, temperature) on denitrification rates in headwater streams, which can cause interannual
variations in rates of nitrate export to downstream waters. A dynamic transport model using a one-
dimensional version of the advection-dispersion equation was developed to estimate progressive
instream nitrate removal from first- to fourth-order streams (Alexander etal., 2009). Model simulations
indicated that denitrification rate constants in headwater tributaries varied strongly by season, based on
biogeochemical and hydrologic factors. This in turn had a cumulative effect on downstream nitrate
export (Alexander et al., 2009). These studies highlight how stream size affects nitrogen-based chemical
connections, with headwater streams within the network affecting downstream water quality.
Phosphorus-based chemical connections also have been documented. Doyle et al. (2003) modeled the
relative influence of hydrogeomorphic and uptake processes on longitudinal phosphorus retention
through a river network of first- through sixth-order streams. The model revealed greater variation in
uptake relative to hydrogeomorphic processes, and the authors concluded that uptake processes
influence downstream variation in phosphorus retention at the watershed scale more than
hydrogeomorphology.
Research on hydrologic control and seasonality of nutrient export from streams in the Mississippi River
basin similarly provides evidence of downstream connectivity (Section B.4.3.2.1). Export of dissolved
reactive phosphorus from second- and fourth-order streams in agricultural watersheds occurred mainly
during high-discharge conditions, with discharges equal to and greater than the 90th percentile
exporting 84% of the dissolved reactive phosphorus, primarily during January and June (Royer et al.,
2006). Similar patterns have been documented in total phosphorus concentrations of first- through
fourth-order streams from another Mississippi River basin (Bayless et al., 2003). In another study,
researchers modeled riverine dissolved reactive phosphorus yield of 73 watersheds within the
Mississippi River basin during the January to June period, as a function of nutrient sources and
precipitation (Jacobson et al., 2011). Riverine dissolved reactive phosphorus yield was positively related
to fertilizer phosphorus inputs, human sources of phosphorus (e.g., sewage effluent), and precipitation,
which generates surface runoff that moves fertilizer applied to the landscape into streams and rivers
that then transport it downstream (Jacobson et al., 2011). These studies demonstrate the connections
and processes by which nutrients exported from streams in the Mississippi River basin contribute to
anoxia in the Gulf of Mexico (Rabalais etal., 2002).
Other environmental and biological processes also can affect nutrient-based chemical connections. The
underlying geology of the Mokelumne River in California's central Sierra Nevada Mountains affected the
spatial and temporal variability in chemical connections. Holloway et al. (1998) examined water quality
in that watershed to identify primary sources of nitrate entering downstream reservoirs. They
conducted a paired watershed comparison with two ephemeral streams in nearby watersheds that were
underlain with different rock types (diorite vs. biotite schist) but had similar land-use, vegetation,
topography, and watershed area. Many samples from the diorite watershed had nitrate concentrations
below detection limits (<4 uM), with a median concentration of 3.3 uM; concentrations were not
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strongly associated with the start or end of the high precipitation period. In the biotite schist watershed,
maximum stream concentrations of nitrate (>300 uM) occurred at the start of the high precipitation
period, and concentrations decreased over time. A nearby perennial stream, also in a biotite schist
watershed, displayed this same temporal trend, with highest nitrate concentrations at the beginning of
the rainy season and decreasing concentrations during the spring. Holloway et al. (1998) concluded that
biotite schist streams contributed a disproportionately large amount of total nitrate to downstream
reservoirs, despite draining only a small area of the entire watershed.
In another study, nitrate concentrations were measured at 50 sites across the West Fork watershed of
the Gallatin River in southwestern Montana's northern Rocky Mountains under different hydrologic
conditions and across two seasons, growing and dormant (Gardner and McGlynn, 2009). Streams ranged
from first-order mountain streams to fourth-order streams near the West Fork-Gallatin River
confluence. In the dormant season, the distance over which nitrate concentrations were spatially
correlated ranged from 3.2 to 5.5 km. In the growing season, this range decreased to 1.9 to 2.7 km. This
seasonal difference likely resulted from greater biological uptake and use of nitrate during the growing
season, which then limited its downstream transport; during the dormant season, downstream
transport increased, resulting in greater spatial dependence in nitrate concentrations (Gardner and
McGlynn, 2009).
Seasonal variability in chemical connectivity also was observed in Arizona's San Pedro River. Differences
in dissolved organic nitrogen concentration were detected among three segments of the river during the
dry season, but stream water was well mixed, the system was hydrologically connected, and no
differences in dissolved organic nitrogen concentration were detected during the wet season (Brooks
and Lemon, 2007). These seasonal differences occur because nitrogen accumulates locally atvarying
levels during drier periods but is mixed and transported downstream during large, infrequent storm
events, making nitrogen concentrations more longitudinally uniform (Fisher et al., 2001).
Peterson et al. (2001) examined chemical connectivity by studying similar network components across
different types of river networks. After measuring nitrogen export from 12 headwater streams
distributed throughout the contiguous United States, Alaska, and Puerto Rico, they found that uptake
and transformation of inorganic nitrogen were most rapid in the smallest headwater streams (Peterson
etal., 2001). Given the prevalence of headwater streams on the landscape (Section 3.2) and their
hydrologic connectivity to other river network components (Sections 2.2 and 3.3.1), headwater stream
nitrogen processing can improve water quality in downstream waters. Many other studies also highlight
the importance of nitrogen processing in headwater streams (e.g., Hill et al., 1998; Hill and Lymburner,
1998; Triska etal., 2007). Mulholland et al. (2008) measured in situ rates of nitrate removal by
denitrification in 72 streams across different biomes and used those rates to model how headwater and
larger streams in a river network respond to simulated nitrate loading increases. At low loading rates,
the biotic removal of dissolved nitrogen from water is high and occurs primarily in headwater streams,
which reduces loading to larger streams and rivers downstream. At moderate loading rates, the ability
of headwater streams to remove nitrogen is reduced, but larger streams can remove the excess nitrogen.
At high loading rates, removal by headwater streams and larger streams in the river network is
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ineffective, resulting in high nitrogen export to rivers (Mulholland et al., 2008). Similar results were
obtained by Wollheim et al. (2008) in the Ipswich River, MA.
Helton etal. (2011) conducted simulation experiments that illustrated the effects of connectivity in the
Ispwich River (MA) and Flat Creek (WY) networks, via the use of river-network models of nitrate
dynamics. The nitrate models underpredicted nitrogen removal in many reaches, which was attributed
to connections between the river channels and neighboring wetlands that were not characterized by the
model and that functioned as nitrogen sinks (Section 4.3.3.2). By not representing the fine-scale
variability in nitrogen uptake in river-network models and assuming that nitrogen uptake decreases
with depth along a river network, simulations can potentially misrepresent the export of nitrogen from
headwater streams to downstream waters (Darracq and Destouni, 2005, 2007). The potential for this
misrepresentation, however, depends on the spatial scale of the study and the specific characteristics of
the river network.
The influences of headwater and other tributary streams on nutrient concentrations in larger
downstream waters, as detailed in the numerous examples above, reflect the combined processes of
nutrient cycling and downstream transport that occur throughout river networks, albeit most
intensively in headwater streams. The concept of nutrient spiraling provides an approach to quantifying
these cycling and transport processes and a relatively simple framework for understanding their
implications. As nutrients cycle through various forms or ecosystem compartments, being consumed
and regenerated for reuse, they complete a "cycle" only after having been displaced some distance
downstream, which stretches the cycle into a helix or "spiral" (Webster and Patten, 1979). The stretch of
the spiral, or the openness between its loops, is primarily determined by flow, whereas the diameter of
the loops is mainly determined by biological activity (Cummins et al., 2006). Nutrients such as dissolved
phosphorus and nitrogen, which enter the stream via ground-water or overland flow, are removed from
the water column by algae and microbial organisms. These nutrients are then consumed by organisms at
higher trophic levels, transported farther downstream as suspended particles, or returned to the
dissolved pool through cell death and lysis. Nutrients flowing through the food web also are regenerated
to the dissolved pool via excretion and microbial decomposition. Nutrients in the dissolved, particulate,
and living tissue phases of the cycling process are subject to downstream transport, such that each
phase transition moves some distance downstream. The average downstream distance associated with
one complete cycle—from a dissolved inorganic form in the water column, through microbial uptake,
subsequent transformations through the food web, and back to a dissolved available form—is termed the
"spiraling length."
Although measurement of total spiraling length requires detailed study of tracer dynamics through
multiple compartments of the stream ecosystem, Newbold et al. (1981; 1983a) have shown that it can be
approximated by "uptake length" or the distance traveled in the water column before algal and
microbial assimilation occurs. Uptake lengths for phosphorus and nitrogen can be estimated precisely
only from tracer additions of radioactive or stable isotopes, but they can be roughly estimated from
experimental additions that briefly raise the concentration of the natural form of the nutrient Ensign
and Doyle (2006) compiled results of 404 measurements of uptake length of phosphate, ammonium, and
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nitrate in streams and rivers ranging from first- to fifth-order. For a given stream order, they estimated
the number of cycles that each nutrient had undergone as the ratio of median uptake length to the
average length of stream for that stream order (from Leopold et al., 1964). They found that the three
nutrient forms cycle between roughly 8 (nitrate) and 40 (ammonium) times within the length of a first-
order stream, and between roughly 8 and 90 times within the respective lengths of first- to fourth-order
streams.
Withers and Jarvie (2008) also compared phosphorus uptake lengths among different streams. Shorter
uptake lengths are indicative of more rapid phosphorus cycling and greater efficiency of phosphorus
retention. The shortest uptake lengths (2-580 m) were in first-order streams that drained "pristine"
watersheds. Uptake lengths were longer (26-3,460 m) in second- to fourth-order streams that drained
agricultural watersheds, and longest (4,140-367,000 m) in fifth-order rivers that drained a mixture of
urban and agricultural land use (Withers and Jarvie, 2008).
These studies highlight the high nutrient-processing potential of headwater streams. This potential
results from their low water volume-to-bed sediment area ratio, which enhances conditions for key
nutrient uptake processes (e.g., adsorption, precipitation, assimilation) not only at the water-bed
interface but within the streambed sediments (Withers and Jarvie, 2008). Downstream ecosystems
depend on processes that occur in headwater streams. Given that roughly half the water reaching larger
tributaries and rivers originates from headwater streams (Section 3.3.1), the results of Ensign and Doyle
(2006) make clear that phosphorus and nitrogen arrive at downstream waters having already been
cycled many times in headwater and smaller tributaries. This cycling is, fundamentally, a complex of
ecosystem processes that intensively uses nutrients and then regenerates them for delivery to
downstream waters much in their original form. Because nutrients undergo transformations across
various forms (e.g., dissolved, particulate, inorganic, or in living organisms) while being transported
downstream (i.e., spiraling), explicitly identifying their exact origin in the network can be difficult.
Although headwater nutrient cycling, or spiraling, functions largely to deliver regenerated nutrients
downstream, headwater stream processes measurably alter the delivery of nutrients to downstream
waters in many ways. For example, if cycling has been seriously impaired such that nutrient
regeneration is inhibited or nutrients are generated in biologically unavailable or toxic forms, the
downstream effects could be large. Nutrients taken up as readily available inorganic forms can be
released back to the water column as organic forms (Mulholland et al., 1988) that are less available for
biotic uptake (Seitzinger etal., 2002). Similarly, nutrients incorporated into particles are not entirely
regenerated (Merriam et al., 2002; Hall et al., 2009), but rather accumulate and contribute to
longitudinally increasing particulate loads (Whiles and Dodds, 2002). The amount of phosphorus and
nitrogen delivered downstream by headwater streams cycles seasonally due to the accumulation of
nutrients in temporarily growing streambed biomass (Mulholland and Hill, 1997; Mulholland et al.,
2004). Such variations affect downstream productivity (Mulholland et al., 1995) and help explain the
seasonality in the spatial correlations of nutrient concentrations described above.
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Microbially mediated transformations affect the forms of nitrogen transported from headwater streams
to downstream waters, and these transformations can influence—and be influenced by—human
alterations of the landscape. Nitrification, or the transformation of ammonium to nitrate, occurs
naturally in undisturbed headwater streams (e.g., Bernhardt et al., 2002) but increases sharply in
response to ammonium inputs (e.g., Newbold et al., 1983b), thereby reducing potential ammonium
toxicity from pollutant inputs (Chapra, 1996). Denitrification, which removes nitrate from stream-water
through transformation to atmospheric nitrogen, is also widespread among headwater streams, as
demonstrated by stable isotope tracer additions to 72 streams in the conterminous United States and
Puerto Rico (Mulholland etal., 2008). Mulholland etal. (2008) estimated that headwater streams
(<100 L s"1, about third order or less) free from agricultural or urban impacts reduce downstream
delivery of nitrogen by 20-40%. Alexander etal. (2007) and Wollheim et al. (2008), using earlier and
less extensive measurements of denitrification rates, estimated nitrogen removal of 8 and 16% by
stream networks of first to third order and first to fifth order, respectively. In headwater agricultural
streams, denitrification in stream sediments might not be effective at removing nitrate from stream
water because of altered hydrology. In watersheds with tile drains and channelized headwaters, stream
nitrate concentration is positively correlated with stream discharge, suggesting that these altered
streams are in throughput mode, whereby nitrate inputs are rapidly transported downstream with little
retention or processing (Royer et al., 2004).
Small tributaries also affect the downstream delivery of nutrients through abiotic processes. Meyer and
Likens (1979) showed that phosphorus concentrations in a forested first-order New Hampshire stream
were reduced by sorption to stream sediments. A much stronger sorption of phosphorus by stream
sediments was observed by Simmons (2010) in first- to third-order West Virginia streams impacted by
acid mine drainage, where phosphorus sorbed to metal hydroxide precipitates introduced by mine
drainage. These examples further illustrate the potential for headwater streams to absorb nutrient
impacts to the benefit of downstream waters.
3.4.2 Dissolved and Particulate Organic Matter
Headwater streams supply downstream waters with dissolved and particulate organic carbon, which
support biological activity throughout the river network. Organic carbon enters headwater streams
from the surrounding landscape, including wetlands (Section 4.3.3.4 and 4.4.3.1), in the form of
terrestrial leaf litter and other seasonal inputs, dissolved organic carbon (DOC) in subsurface and
surface runoff, and fine particulate organic matter (including eroded soil) in surface runoff. Headwater
reaches also export organic carbon produced within the stream by photosynthesis, both as DOC (Kaplan
and Bott, 1982) and suspended particles (Marker and Gunn, 1977; Lamberti and Resh, 1987).
Agren etal. (2007) determined that headwater streams exported the largest amount of terrestrial DOC
on a per unit basis in the Krycklan watershed in Sweden. The amount of organic matter exported from
headwater streams to downstream waters varies with multiple factors, including surrounding land use.
For example, Schelker et al. (2014) developed a mixing-model approach and quantified that forest
harvesting at areal proportions of 11% and 23-25% of a northern Sweden watershed induced stepped
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increases in DOC delivery, due to disturbance of shallow forest soils and subsequent transport from
headwaters to downstream locations. Similarly, a 20% increase in downstream DOC concentrations was
predicted folio wing forest harvesting in the headwater areas of the H.J. Andrews Long Term Ecological
Research site, using the VELMA (Visualizing Ecosystems for Land Management Assessments) model
(Abdelnour etal., 2013). In southeast Arizona, Meixner etal. (2007) found that DOC consistently
doubled to tripled in the San Pedro River during storm events from a flush of terrestrial organic matter
and nutrients. This is comparable to the flush response observed by others (Fisher et al., 1982; Brooks et
al., 2007) during monsoon precipitation events in the southwestern United States. These examples
further demonstrate connectivity of headwater streams and their cumulative effects on downstream
water quality.
Fisher and Likens (1973) followed the fate of these inputs in a forested headwater stream in New
Hampshire. They concluded that 34% of inputs were mineralized through respiration by consumers and
microbes within the headwater stream, which represented the "ecosystem efficiency" of the reach. The
remaining 66% was exported downstream and constituted, as Fisher and Likens (1973) observed, "...
inputs to the next stream section where they are assimilated, or passed on (throughput) or both." Other
studies have reported similar amounts of export. Webster and Meyer (1997) compiled organic matter
budgets from 13 North American first- and second-order streams. The median ecosystem efficiency was
31%, implying a median export of 69% of organic matter inputs. A large body of literature has
demonstrated that headwater streams modify and export organic carbon that significantly affects
ecosystem processes throughout the river network.
Vannote et al. (1980) recognized that exported carbon was not simply the unutilized fraction but was
also greatly modified in character. A basic tenet of their River Continuum Concept is that longitudinal
variations in the structure of stream ecosystems reflect, in part, the cumulative effects of upstream
organic matter processing. Much or most of the organic carbon exported from headwater streams has
been altered either physically or chemically by ecosystem processes within the headwater reaches. Leaf
litter contributes an average of 50% of the organic matter inputs to forested headwater streams
(Benfield, 1997), but leaves and leaf fragments (>1 mm) accountfor only 2% or less of organic matter
exports (Naiman and Sedell, 1979; Wallace etal., 1982; Minshall etal., 1983). The conversion of whole
leaves to fine particles (<1 mm) involves physical abrasion, microbial decomposition, and invertebrate
feeding and egestion (Kaushik and Hynes, 1971; Cummins et al., 1973; Petersen and Cummins, 1974).
The rate of that conversion is affected by whether the leaves are in an aerobic environment, such as
riffles, or an anaerobic environment, such as depositional pools (Cummins et al., 1980). Feeding
activities of aquatic invertebrates called "shredders" break down leaves that have entered streams
(Cummins and Klug, 1979; Cummins etal., 1989). Invertebrate activity is particularly important, as
demonstrated by large reductions of fine particle export following experimental removal of
invertebrates from a headwater stream (Cuffney et al., 1990; Wallace et al., 1991). Strong invertebrate
influence on fine particle export also has been inferred from analysis of seasonal (Webster, 1983) and
daily (Richardson et al., 2009) variations.
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Downstream organisms consume organic carbon exported from headwater streams, supporting
metabolism throughout the river network. In part, this results from direct consumption of detrital
organic matter (Wallace et al., 1997; Hall et al., 2000), but much of the metabolic consumption of organic
matter in streams occurs via microbial decomposition (Fisher and Likens, 1973). The microbes
themselves are then consumed by other organisms (Hall and Meyer, 1998; Augspurger et al., 2008),
whose energy in turn supports the food web through what is known as the "microbial loop" (Meyer,
1994). In addition to transformations associated with microbial and invertebrate activity, organic
matter in streams can be transformed through other processes such as immersion (Corti et al., 2011)
and abrasion (Paul etal., 2006); photodegradation also can be important in ephemeral and intermittent
streams where leaves accumulate in dry channels exposed to sunlight (Dieter et al., 2011; Fellman et al.,
2013).
The organic carbon turnover length, derived from the nutrient spiraling concept (Section 3.4.1; Newbold
et al., 1982b), is a measure of the downstream fate of exported carbon. Carbon turnover length is the
ratio of the downstream flux of organic carbon to ecosystem respiration per length of stream. It
approximates the average distance that organic carbon would travel before being consumed and
mineralized by aquatic organisms. Carbon turnover length for first-order streams is on the order of 1 to
10 km (Newbold et al., 1982b; Minshall et al., 1983), suggesting that organic carbon exported from
headwater streams is likely to be used primarily in the somewhat larger streams to which they are
direct tributaries (i.e., second- or third-order streams). The carbon turnover length, however, actually
represents a weighted average of widely varying turnover lengths associated with the diverse array of
particulate and dissolved forms of organic carbon in stream and river ecosystems (Newbold, 1992).
Turnover lengths of specific organic carbon forms can be estimated if their rates of downstream
transport and mineralization (or assimilation) are known. For example, Webster et al. (1999) estimated
a turnover length of 108 m for whole leaves in a North Carolina second-order stream, but a much longer
turnover length of 40 km for fine (<1 mm) organic particles. Newbold et al. (2005) obtained similar
estimates of 38 and 59 km for the turnover lengths of two different size fractions of fine organic
particles in a second-order Idaho stream. Kaplan et al. (2008) concluded that DOC in a third-order
southeastern Pennsylvania stream consisted of a rapidly assimilated "labile" fraction with a turnover
length of 240 m, a more slowly assimilated "semilabile" fraction with a turnover length of 4,500 m, and a
"refractory" fraction with immeasurably slow assimilation, implying an indefinitely long turnover length
sufficient to carry the carbon to coastal waters.
Because turnover length increases with stream size, organic carbon that travels to a larger order stream
is likely to travel farther than its original turnover length predicts (Minshall et al., 1983; Webster and
Meyer, 1997). For example, the organic carbon turnover length of the Salmon River, ID increased from
3.7 km in a second-order headwater stream to 1,200 km in the eighth-order reach, about 600 km
downstream (Minshall et al., 1992). In a modeling study, Webster (2007) estimated that turnover length
increased from several hundred meters in the headwater streams to greater than 100 km in a large
downstream river. This progression of increasing turnover length from headwater streams to
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increasingly larger streams and rivers implies that organic carbon exported from headwaters supports
metabolism throughout the river network.
Although turnover length reflects the spatial scale over which upstream exports of organic carbon are
likely to support downstream metabolism, it does not provide direct evidence for or quantify the actual
use of organic carbon in downstream reaches. Studies of transport and mass balance throughout the
river network provide such evidence. Shih et al. (2010) applied the SPARROW (SPAtially Referenced
Regressions On Watershed attributes) model to organic carbon data from 1,125 monitoring sites
throughout the conterminous United States. They estimated that all river reaches (large and small)
delivered an annual average of 72 kg C ha-1 of incremental drainage area, whereas the river systems as a
whole exported 30 kg C ha-1. Thus, 58% of carbon inputs were respired within the river networks, while
the rest (42%) were transported downstream. Shih et al. (2010) did not specify the proportion of inputs
originating from headwater streams, but using their results (with certain assumptions), we can estimate
the amount of organic carbon in river networks that originates from headwater streams. We begin with
the proportion of carbon originating from allochthonous sources as 0.78 (Shih etal., 2010). If we assume
that the proportion of headwater streams in a drainage area is 0.50 (Section 3.2; Alexander etal., 2007;
Caruso and Haynes, 2011), headwater streams then provide 0.39 (= 0.78 x 0.50) of the total organic
carbon supply, with the input from the larger downstream network being 0.61 (i.e., 61%) of the carbon
supply. Using the ecosystem efficiency for headwater streams of 31% (Webster and Meyer, 1997), the
proportion of carbon originating from headwater streams that is delivered downstream is 0.39 x (1 -
0.31) = 0.27. The proportion of carbon exported from headwater streams (0.27), plus the proportion of
carbon input directly to the downstream network (0.61), equals the total carbon input to the
downstream network (0.88). Thus, 31% (= 0.27/0.88 x 100) of the total carbon supplied to downstream
reaches originates from headwater streams.
Most terrestrial organic matter that enters headwater streams is transported downstream (Gomi et al.,
2002; MacDonald and Coe, 2007), typically as fine particulate or dissolved organic matter (Bilby and
Likens, 1980; Naiman, 1982; Wallace et al., 1995; Kiffney et al., 2000). These headwater streams also can
export significant amounts of autochthonous organic matter via the downstream transport of benthic
algae (Swanson and Bachmann, 1976). Both allochthonous and autochthonous organic matter can be
transported significant distances downstream (Webster et al., 1999), especially during high flows
(Bormann and Likens, 1979; Naiman, 1982; Wallace etal., 1995). The importance of discharge in
determining organic matter transport dynamics highlights the interdependence of physical and
biological connections within the river network. For example, Wallace et al. (1995) examined coarse
particulate organic matter export in three headwater streams in North Carolina and found that 63-77%
of export over a 9-year period occurred during the 20 largest floods. This finding suggests that
headwater streams (including ephemeral and intermittent streams) can provide temporary storage for
organic matter (Gomi et al., 2002), which is then transported downstream during storms or snowmelt.
Exports also can vary seasonally, increasing in autumn and winter when deciduous trees drop their
leaves (Wipfli et al., 2007) and in the spring when flowers and catkins are shed.
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The amount of organic matter exported from headwater streams can be large, and often depends on
factors such as discharge, abiotic retention mechanisms within the channel (Bilby and Likens, 1980),
biological communities (Cuffney et al., 1990), and the quality and quantity of riparian vegetation in
headwater watersheds (Wipfli and Musslewhite, 2004). For example, Wipfli and Gregovich (2002) found
that organic matter export ranged from <1 to 286 g of detritus (dead organic matter) per stream per day
in 52 coastal headwater streams in Alaska. When debris dams were removed from a New Hampshire
headwater stream, export of fine particulate organic carbon increased by 632% (Bilby and Likens,
1980). The longitudinal discontinuities created by logjams and beaver dams slow the downstream
transport of organic matter, enabling instream organisms to process the carbon and slowly leak material
downstream (Wohl and Beckman, 2014). The strong links among organic matter storage, processing,
and downstream transport in ephemeral streams of the southwestern United States can be seen in the
distribution of organic matter of varying quality and mobility over periods with varying rainfall
intensities (Norton et al., 2007). Arroyos or ephemeral channels in northeastern New Mexico are
important in transporting and transforming organic matter that enhances the fertility of agricultural
areas along downstream alluvial fans. More frequent but low-intensity rainfall was important in driving
biochemical transformations that altered organic matter mobility and quality, which was subsequently
transported downstream by larger storms (Norton et al., 2007). Traditional farming practices in the
region relied on the temporary storage, transformation, and transport of organic matter from ephemeral
streams (Norton et al., 2007; Sandor et al., 2007).
Although organic matter clearly is exported from headwater streams, effects on downstream organisms,
and the distance over which these effects propagate are difficult to quantify (Wipfli et al., 2007). Many
downstream organisms rely on organic matter and its associated microbes for food, but demonstrating
where in the river network such material originates presents a challenge. Similarly, the conversion of
organic matter to other forms (e.g., invertebrate or fish biomass via consumption), each with its own
transport dynamics, makes tracking sources of downstream contributions difficult. Given the prevalence
of headwater streams in both the landscape and the river network (Leopold et al., 1964), and their
primacy in organic matter collection and processing, a logical conclusion is that headwater streams
exert a strong influence on downstream organic matter dynamics. Benstead and Leigh (2012) estimated
that headwater streams, including intermittent and ephemeral channels, result in a global carbon efflux
of 1.6 Pg C yr1, making the overall contributions of rivers and streams about equivalent to all inland
lakes and wetlands combined. In addition, headwater streams also serve as a source of colonists for
downstream habitats (Section 3.5). For example, headwater springs can provide algae a winter refuge
from freezing, then serve as a source of propagules for downstream reaches upon spring thaws (Huryn
etal.,2005).
3.4.3 Ions
Measurements of ions and conductivity from nested study designs provide additional evidence for
connectivity by various transport mechanisms. Rose (2007) collected data at 52 sampling stations in
Georgia's Chattahoochee River basin, which includes the heavily urbanized region of Atlanta, over a
2-year period. The study sought to characterize baseflow hydrochemistry across a rural-to-urban land-
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use gradient. A plot of the major ion (sodium, bicarbonate alkalinity, chloride, and sulfate)
concentrations versus downstream river distance showed distinct peaks relative to baseflow
measurements, with elevated concentrations persisting downstream.
In a study of mined and unmined streams in the Buckhorn Creek basin in Kentucky, water
measurements taken at several locations within the same tributary had similar conductivity values
(Johnson et al., 2010). As expected, confluences disrupted this spatial similarity along the river network.
Conductivity values along the mainstem decreased at confluences with unmined streams and increased
at confluences with mined streams, demonstrating that headwater streams were transporting ions
downstream and affecting downstream conductivity. This spatial pattern in conductivity was consistent
between spring and summer surveys of the river network.
In a study in Sweden, measurements of pH from the outlets of seven watersheds were statistically
related to headwater pH measurements in those watersheds (Temnerud et al., 2010). As pH at outlets
increased under low-flow conditions, so did median pH of the headwater streams. This study illustrates
the connectivity between the headwater components of the river network and the outlets of the
watersheds and the cumulative effects of headwater streams to downstream waters.
3.4.4 Contaminants and Pathogens
The movement of contaminants—that is, substances that adversely affect organisms when present at
sufficient concentrations—and waterborne pathogens provides another line of evidence for chemical
connectivity between tributaries and the river network. Existing information typically has been derived
from either empirical experiments that release tracer substances into streams to monitor movement
along a longitudinal gradient or the use of modeled projections and characterization of contaminants.
Studies also have examined trace metal data collected at multiple sites throughout a specific watershed,
relative to a point source or a complex mixture of point-source inflows (e.g., active mining areas,
wastewater treatment plant discharges). These studies provide a way to understand sediment transport
in streams and rivers and to determine how metals are spatially and temporally dispersed in the
watershed (Rowan etal., 1995).
The degree of surface-water and ground-water mixing or exchange in the hyporheic zone influences the
transport and uptake of trace metals. In a 7 km perennial stream segment contaminated by copper
mining in Arizona, 20% of the dissolved manganese load was removed by microbial activity that was
likely stimulated by the physicochemical conditions and increased residence time (compared with
surface channel residence time) associated with hyporheic exchange (Harvey and Fuller, 1998). That
oxidation of manganese enhanced the uptake of other trace metals and thereby decreased cobalt, nickel,
and zinc loads 12-68% over the 7 km reach (Fuller and Harvey, 2000). Modeling the contributions of
hyporheic exchange on contaminant dynamics over entire river networks requires further research.
Another example of chemical connections along the river network is how inputs of water associated
with natural gas (coalbed methane) extraction and hardrock mining can influence trace element and
dissolved solute concentrations in perennial rivers. Patz etal. (2006) examined trace elements and other
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water quality parameters in ephemeral streams resulting from coalbed methane extraction activities
that are connected to the perennial Powder River, WY. Iron, manganese, arsenic, fluoride, dissolved
oxygen, pH, and turbidity differed across sample locations, demonstrating connectivity between
wellhead discharge and ephemeral streams. The contribution of ephemeral streams was detected in the
Powder River, where pH was consistently elevated downstream of the confluence with a high-pH stream
(Patzetal., 2006).
In a broader study, Wang et al. (2007) used retrospective USGS data (1946-2002) to investigate spatial
patterns in major cation and anion concentrations related to coalbed methane development in the
Powder River basin (33,785 km2) in Wyoming and Montana. The study indicated that coalbed methane
development could have detrimental effects on the Powder River, especially in terms of sodium
adsorption ratio (sodicity). Although the authors indicated connectivity and adverse effects in stream
quality with increased sodium and stream sodicity, data also revealed inconsistent patterns associated
with complex spatial variability within the drainage basin due to the geographic distribution of the
coalbed methane wells.
The spatial extent of metal transport has been demonstrated in the upper Arkansas River of Colorado,
where the headwaters have been affected by past mining activities (Kimball et al., 1995). Bed sediments
sampled from the headwaters to approximately 250 km downstream showed an inverse relationship
between sediment cadmium, lead, and zinc concentrations and downstream distance. That same spatial
distribution pattern in bed sediment metal concentrations was observed from headwater streams to the
downstream Clark Fork River in Montana, which has been impacted by mining and smelting activities in
its headwaters (Axtmann and Luoma, 1991). Based on regression models, bed sediment metal
concentrations from river sites were inversely related to downstream distance, and predictions from
those models indicated that sediments with metals originating from headwater mining and smelting
areas were reaching Lake Pend Oreille, more than 550 km downstream. Hornberger et al. (2009) used a
19-year data set from the Clark Fork River, with sites from the headwater streams to 190 km
downstream, and found that bed sediment copper concentrations at downstream sites were positively
correlated with concentrations at upstream sites.
Lewis and Burraychak (1979) examined the downstream transport of heavy metals from ephemeral and
intermittent streams to a downstream perennial stream, due to the impacts of active and abandoned
copper mines. Water chemistry in Pinto Creek was monitored biweekly for 2 years at four stations, one
above and three below a point discharge associated with the Pinto Valley Mine in east-central Arizona
(Lewis and Burraychak, 1979). Surveys offish, aquatic macroinvertebrates, and vegetation were
conducted during the same period at 13 sampling stations along the total stream length. Contaminants
from the Pinto Valley Mine entered Pinto Creek via accidental discharge of waste from tailings ponds
(Lewis, 1977). Monitoring revealed that mine wastes comprised up to 90% of total flow in Pinto Creek,
and that most chemical parameters increased in concentration below the discharge point, then
decreased progressively downstream (Lewis and Burraychak, 1979). Increases in sulfate, conductivity,
and total hardness between above-mine and below-mine locations were most apparent, although
increases in heavy metals and suspended solids were considered most detrimental to organisms.
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Suspended solids settled in and buried intermittent channels, which contained up to 50 cm of mine-
waste sediment; these sediments were present all the way to the stream terminus. Increased heavy
metal concentrations in the food web and sediments also were detected below the discharge point
(Lewis and Burraychak, 1979).
Lampkin and Sommerfeld (1986) similarly showed that intermittent streams can contribute highly
mineralized, acidic waters to a downstream perennial reach, in a study that characterized acid mine
drainage impacts on water and sediment chemistry (particularly major cations, silica, sulfate, selected
heavy metals, and acidity) in Lynx Creek, a small intermittent stream in east-central Arizona. Six
stations, two above and four below an abandoned copper mine, were monitored (water and sediment
samples) monthly for 1 year. Specific conductance, pH, and dissolved ion concentrations varied with
proximity to the mining complex. Concentrations of most constituents were higher near the mine and
progressively decreased downstream toward the terminus of Lynx Creek, due to precipitation and
dilution by headwater streams. All heavy metal and sulfate concentrations were higher at the immediate
discharge location versus the above-mine stations; sulfate concentrations downstream of mine-drainage
inputs also significantly differed from the rest of the creek. Sediments throughout the creek were high in
metals, suggesting downstream transport of contaminated sediments. Acid-mine drainage from the
mine had a major but mostly localized impact on Lynx Creek.
As discussed in previous sections, headwater streams are connected to downstream waters through the
transport of chemicals but also through transformation processes. Boreal river networks, in which
headwater streams are sources of DOC and pH increases downstream, provide these transformations.
Iron exported from the acidic headwater tributaries is bound to DOC (mobile form). As pH increases,
iron-rich ground water enters the channel, and iron transforms to iron (oxy) hydroxides that aggregate
and precipitate out of solution (Neubauer etal., 2013). These iron (oxy) hydroxides can function as
carriers of toxic metals and metalloids (e.g., arsenic), thereby removing them from solution and
temporarily storing them in and along the river network (Neubauer et al., 2013).
Several studies also have projected the cumulative effect of headwater systems on downstream mercury
concentrations and loads in response to land use, climate, and atmospheric deposition. The Water
Quality Analysis Simulation Program and the Bioaccumulation and Aquatic System Simulator models
were used to predict changes in water, sediment, and fish-tissue mercury concentrations across water
bodies with varying upstream headwater drainage areas (Knightes et al., 2009). Simulations predicted
that watersheds with high headwater drainage densities would exhibit longer lag times for mercury
delivery downstream compared to those with low headwater drainage densities. This work suggests
that headwater streams can serve a mercury storage function, and that temporally varying connectivity
contributes to the transport of mercury from headwater streams to downstream waters.
The cumulative effects of land-cover change on total and methylmercury fluxes from a North Carolina
headwater watershed to the Cape Fear River were simulated using the Grid Based Mercury Model
(Golden and Knightes, 2011). The simulations estimated a 95% increase in total mercury fluxes from the
landscape to downstream waters in response to new suburbanization and a 7% decrease in total and
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methylmercury export in response to reforestation. Predicted changes in total mercury fluxes from the
landscape to the downstream assessment point resulted primarily from changes in landscape land
cover, rather than changes in connections within the river network.
The effects of climate change on total mercury export from headwater tributaries draining a Coastal
Plain watershed (79 km2) in South Carolina were simulated using multiple watershed models (Golden et
al., 2014). Results indicated increased total mercury export under the high-precipitation scenario and
decreased total mercury export under the low-precipitation scenario, showing that precipitation, and
thus hydrologic connections, drive mercury transport from headwater streams to downstream waters.
Contaminants are commonly transported from tributaries to downstream rivers bound to sediments.
Using isotopic fingerprinting, Gehrke et al. (2011) identified different tributaries as contributing to
downstream mercury contamination of surface sediments in San Francisco Bay. Historic gold mining in
the tributary watersheds of the San Joaquin and Sacramento Rivers contributed to contaminated
mercury sediments in the northern part of San Francisco Bay, whereas wastes from mercury mine
operations were delivered to the southern part of the bay via the Guadalupe River (Gehrke et al., 2011).
Studies of radionuclide (e.g., plutonium, thorium, uranium) distribution, transport, and storage provide
convincing evidence for long-distance chemical connections in river networks. Although the natural
occurrence of radionuclides is extremely rare, their production, use, and release for military and energy
applications have been monitored for more than 50 years. Like metals, radionuclides adsorb readily to
fine sediment; thus, the fate and transport of radionuclides in sediment generally mirrors that of fine
sediment. From 1942 to 1952, plutonium dissolved in acid was discharged untreated into several
intermittent headwater streams that flow into the Rio Grande at the Los Alamos National Laboratory,
NM (Graf, 1994; Reneau et al., 2004). These intermittent headwaters drain into Los Alamos Canyon
(152 km2 drainage area), which joins the Rio Grande approximately 160 km upriver from Albuquerque.
Also during this time, nuclear weapons testing occurred west of the upper Rio Grande near Socorro, NM
(Trinity blast site) and in Nevada. The San Juan Mountains in the northwestern portion of the upper Rio
Grande basin (farther upstream from the site where Los Alamos Canyon enters the Rio Grande) is the
first mountain range greater than 300 m in elevation east of these test locations. The mountains
therefore have higher plutonium concentrations than the latitudinal and global averages because of
their geographic proximity to the test sites. The mountain areas are steep with thin soils, so plutonium
from testing fallout was readily transported to headwater streams in the upper Rio Grande basin via
erosion and subsequent overland movement. The distribution of plutonium within the Rio Grande
illustrates how headwater streams transport and store contaminated sediment that has entered the
basin through both fallout and direct discharge. Although Los Alamos Canyon represented only 0.4% of
the drainage area at its confluence with the Rio Grande, its mean annual bedload contribution of
plutonium was almost seven times that of the mainstem (Graf, 1994). Much of this contribution occurred
sporadically during intense storms that were out of phase with flooding on the upper Rio Grande. Total
estimated contributions of plutonium to the Rio Grande are approximately 90% from fallout to the
landscape and 10% from direct effluent at Los Alamos National Laboratory (Graf, 1994). Based on
plutonium budget calculations, only about 10% of the plutonium directly discharged into Los Alamos
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Canyon and less than 2% of the fallout over the upper Rio Grande basin have been exported to the Rio
Grande. Much of the plutonium is adsorbed to sediment and soil that has either not yet been transported
to the river network or is stored on floodplains or in tributary channels (Graf, 1994). Approximately
50% of the plutonium that entered the Rio Grande from 1948 to 1985 is stored in the river and its
floodplain; the remaining amount is stored in a downriver reservoir. Similar export of radionuclides
through a river network has been traced following the Fukushima Dai-ichi Nuclear Power Plant accident
in Japan (Chartin etal., 2013). The highest levels of radionuclide falloutwere in areas drained by
headwater tributaries. Isotopic analysis of sediment-bound radionuclides collected from throughout the
river network over time documented the downstream transfer of contaminated sediment during a
succession of summer typhoons and spring snowmelt (Chartin etal., 2013).
Waterborne pathogens (bacteria, viruses, protozoa) are another class of contaminants of concern
because of the associated risks to human health and well-being. The principal origins of waterborne
pathogens to downstream waters are as point and diffuse sources from livestock and municipal wastes
via tributaries (Ferguson et al., 2003). Rainfall events and waterborne disease outbreaks in the United
States are strongly correlated, pointing to hydrologic connectivity through tributaries and stormwater
drains as a key link in transporting pathogens downstream, where they can overwhelm treatment plants
and eventually contaminate drinking water sources (Curriero et al., 2001). Ephemeral and intermittent
tributaries also transport waterborne pathogens downstream from livestock and human waste (e.g.,
Parker etal., 2010; Wilkes etal., 2013). Moist sediments in and near ephemeral and intermittent
streams can act as temporary pathogen reservoirs (Chase et al., 2012). Survival of fecal indicator
bacteria in dry sediments of an intermittent stream was high and remained constant over 1 month, but
declined to unculturable levels after 51 days at 20 °C and 163 days at 5 °C (Chahinian et al., 2012). As for
contaminants, various physicochemical (e.g., discharge, nutrient concentrations, temperature, humic
acids) and biological (predation, competition) conditions in tributaries can mediate the transport or
inactivation of pathogens (Ferguson et al., 2003).
3.5 Biological Connections
Biological connections are linkages throughout the river network, from headwater streams (including
those with intermittent and ephemeral flow) to their downstream waters, that are mediated by living
organisms or their products (e.g., seeds, exudates, or excreta; Lamberti and Resh, 1987).
Because biological connections often result from passive transport of organisms or their products with
water flow, biological connectivity often depends on hydrologic connectivity (Section 3.3.1). Many living
organisms, however, also can actively move with or against water flow; others disperse actively or
passively over land by walking, flying, drifting, or "hitchhiking." All of these organism-mediated
connections form the basis of biological connectivity between headwater streams and downstream
waters.
Biological connections between upstream and downstream reaches can affect downstream waters via
multiple pathways or functions. For organisms capable of significant upstream movement, headwater
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streams, including ephemeral and intermittent streams, can increase both the amount and quality of
habitat available to those organisms. Many organisms require different habitats for different resources
(e.g., food, spawning habitat, overwintering habitat), and thus move throughout the river network—
both longitudinally and laterally—over their life cycles (Schlosser, 1991; Fausch et al., 2002). For
example, headwater streams can provide refuge habitat under adverse conditions, enabling organisms
to persist and recolonize downstream areas once adverse conditions have abated (Meyer and Wallace,
2001; Meyer etal., 2004; Huryn etal., 2005). Headwater streams also provide food resources to
downstream waters: as Progar and Moldenke (2002) state, "...headwater streams are the vertex for a
network of trophic arteries flowing from the forest upland to the ocean."
In this section, we consider longitudinal biological connections in terms of both the aquatic organisms—
specifically invertebrates and fishes—that move along river networks and their consequent effects on
downstream waters (see Section 3.4.2 for discussion of particulate organic matter dynamics and Section
3.4.4 for discussion of waterborne pathogens). We then discuss the importance of organism movement
throughout the river network for genetic connectivity in a separate section. We also recognize the many
important biological connections between river networks and terrestrial systems (Lamberti and Resh,
1987), but as discussed in Chapter 1, these connections are outside the scope of this document. Lateral
biological connections between the river network and riparian and floodplain habitats are considered in
Chapter 4.
3.5.1 Invertebrates
Headwater streams provide habitat for diverse and abundant stream invertebrates (Meyer et al., 2007)
and serve as collection areas for terrestrial and riparian invertebrates that fall into them (Edwards and
Huryn, 1995; Kawaguchi and Nakano, 2001). These aquatic and terrestrial invertebrates can be
transported downstream with water flow and ultimately serve as food resources for downstream
organisms. Many fish feed on drifting insects (Nakano and Murakami, 2001; Wipfli and Gregovich,
2002), and these organisms can also settle out of the water column and become part of the local benthic
invertebrate assemblage in downstream waters. Drift, however, has been shown to increase
invertebrate mortality significantly (Wilzbach and Cummins, 1989), suggesting that most drifting
organisms are exported downstream in the suspended detrital load (Section 3.4.2).
The downstream drift of stream invertebrates (Miiller, 1982; Brittain and Eikeland, 1988) and the
contribution of terrestrial and riparian invertebrates to overall drift (Edwards and Huryn, 1995;
Kawaguchi and Nakano, 2001; Eberle and Stanford, 2010) have been well documented. For example,
drift estimates in 52 small coastal streams in Alaska ranged from 5 to 6,000 individuals per stream per
day (Wipfli and Gregovich, 2002). This export of invertebrates can be especially high in intermittent and
ephemeral streams, as terrestrial invertebrates accumulate in these channels during dry periods and are
then transported downstream upon channel rewetting (Corti and Datry, 2012; Rosado et al., 2015). The
amount of invertebrate drift often is closely related to stream discharge (e.g., Harvey et al., 2006), as
well as diel invertebrate behavioral patterns that are independent of flow (Rader, 1997). To compensate
for loss of individuals to downstream drift, invertebrate populations in headwater streams are
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maintained and replenished through a combination of high productivity and upstream dispersal
(Hershey et al., 1993; Humphries and Ruxton, 2002). This dispersal creates downstream to upstream
biological connections along the river network; for organisms capable of directed movement over long
distances (e.g., winged adult forms of aquatic invertebrate larvae), these connections can occur over
significant network distances.
Given this evidence, that headwater streams are biologically connected to downstream waters via the
active and passive export of invertebrates is clear, and the cumulative export of invertebrates from
numerous headwater streams to downstream waters can be substantial. As with organic matter,
however, assessing the effect of headwater invertebrate production and export on downstream waters
is difficult, given that these resources enter downstream waters at multiple points and times throughout
the river network. Nevertheless, some studies have documented the importance of drifting invertebrates
for downstream organisms. Wipfli and Gregovich (2002) estimated that drifting insects and detritus (i.e.,
particulate organic matter; Section 3.4.2) from fishless headwater streams in Alaska supported between
100 and 2,000 young-of-year salmonids per km in a large, salmon-bearing stream. This estimate of
headwater importance in systems where juvenile salmonids move into headwater streams to feed and
grow is likely conservative (Section 3.5.2). Other studies have shown increased fish growth with
increased invertebrate drift (Wilzbach et al., 1986; Nielsen, 1992; Rosenfeld and Raeburn, 2009),
indicating that drift does provide a valuable food resource, especially when food is limiting (Boss and
Richardson, 2002).
Headwater streams also serve as habitat for invertebrates. Many invertebrate species are well adapted
to seasonal or episodic periods of drying (Feminella, 1996; Williams, 1996; Bogan and Lytle, 2007) or
freezing temperatures (Banks, 2007) and can be found throughout a range of stream sizes (e.g., Hall et
al., 2001b) and flow regimes (intermittent and perennial, e.g., Feminella, 1996). Intermittent streams
also can provide refuge from adverse biotic conditions. For example, Meyer et al. (2004) found that
native amphipods can persist in intermittent reaches but are replaced by nonnative amphipods in
perennial reaches. After disturbance, these upstream habitats can provide colonists to downstream
reaches. This phenomenon can be especially important in intermittent streams, where permanent
upstream pools can serve as refuges during drying. For example, Fritz and Dodds (2002, 2004)
examined invertebrate assemblages before and after drying in intermittent prairie streams and found
that initial recovery of invertebrate richness, richness of invertebrate drift, and richness of aerially
colonizing insects were negatively related to distance from upstream perennial water. Dry stream
channels also can facilitate dispersal of aquatic invertebrates by serving as dispersal corridors for
terrestrial adult forms (Bogan and Boersma, 2012; Steward etal., 2012).
Headwater stream invertebrates also provide critical functional roles in maintaining physical and
chemical connectivity to downstream waters (Covich et al., 1999). Invertebrates accelerate the
breakdown of coarse particulate organic matter (e.g., leaves) to more mobile fine and dissolved forms
(Section 3.4.2; Wallace and Webster, 1996); promote algal productivity and microbial activity (and
nutrient uptake) by biofilm grazing (Feminella and Hawkins, 1995); and temporarily store and transfer
sediments, nutrients, and contaminants through their trophic and physical activity (e.g., via
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bioconsolidation and bioturbation; Pringle et al., 1993; Walters et al., 2008; Statzner, 2012). The
contribution of invertebrates in controlling sediment mobilization can be substantial. For example,
Statzner (2012) estimated that the discharge necessary to move approximately 0.4 kg of sediments-1 in
the Colorado River would increase by an order of magnitude in response to bioconsolidation by net-
spinning caddisflies and would decrease by an order of magnitude in response to bioturbation by
crayfish.
Diverse and abundant invertebrate assemblages also inhabit the hyporheic zone of river networks
(Stanford and Ward, 1988; Boulton, 2000). Hyporheic assemblages are composed of invertebrate
species that inhabit shallow subsurface sediments within streambeds to various degrees. Some taxa
spend their entire lives in the hyporheic zone (Boulton, 2000). Other taxa spend only part of their life
cycles, typically their earliest larval stages or periods of disturbances, in the hyporheic zone, and others
spend their entire aquatic stages in the hyporheic zone then migrate out for their aerial adult stages
(Boulton, 2000). These hyporheic assemblages make similar contributions to physical and chemical
connectivity with downstream waters as benthic invertebrates do, while also enhancing hyporheic
exchange (Section 3.3.1) through movements and migration within the hyporheic zone (Boulton, 2000).
3.5.2 Fishes
Although some fish species maintain resident headwater populations, many species move into and out
of headwater streams at some point in their life cycles (Ebersole etal., 2006; Meyer etal., 2007). Some
fish species occur only in headwater streams, contributing to regional aquatic biodiversity (e.g., Paller,
1994). As with invertebrates, however, certain fish species can be found throughout a range of stream
sizes (Freeman et al., 2007) and flow durations (Schlosser, 1987; Labbe and Fausch, 2000), and the fish
species found in headwater streams often are a subset of species found in downstream habitats
(Horwitz, 1978). Use of headwater streams as habitat is especially evident for the many diadromous
species that migrate between headwater streams and marine environments during their life cycles (e.g.,
Pacific and Atlantic salmon, American eels, certain lamprey species), and the presence of these species
within river networks provides robust evidence of biological connections between headwater streams
and larger rivers.
Through their activities, migratory fish can be important in modifying habitat, and transforming and
transporting materials (e.g., Taylor etal., 2006; Hassan etal., 2008). Return migration of diadromous
fishes provides a feedback loop in which marine-derived nutrients are transported upstream to
headwater streams, for subsequent processing and export (Section 3.4.1). This example illustrates how
biological connections also can create chemical connections throughout the river network. Migratory
fish also can bioaccumulate and transport contaminants long distances between headwater streams and
downstream waters (e.g., Krummel etal., 2003; Morrissey et al., 2011). Fish also can act as transport
vectors of other organisms (e.g., seeds, pathogens, glochidia), moving other organisms against flow or
extending their dispersal distances (e.g., Chick et al., 2003; Senderovich et al., 2010; Schwalb et al.,
2013). Even nonmigratory taxa can travel substantial distances within river networks throughout their
life cycles (Gorman, 1986; Sheldon, 1988; Hitt and Angermeier, 2008). As a result, the distribution and
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movement of fish throughout river networks can be highly variable, both spatially and temporally
(Schlosser, 1991; Labbe and Fausch, 2000; Fausch et al., 2002).
The importance of connectivity in structuring fish assemblages provides further evidence of biological
connections along river networks. Fish assemblages among connected streams tend to be more similar,
in that assemblages in reaches located closer together tend to have more species in common than
assemblages in distantly separated reaches (Matthews and Robinson, 1998; Hitt et al., 2003; Grenouillet
et al., 2004). Measures of river network structure also can explain fish assemblage structure, with
studies showing that metrics such as link magnitude (the sum of all first-order streams draining into a
given stream segment) and confluence link (the number of confluences downstream of a given stream
segment) are significant predictors offish assemblages (e.g., Osborne and Wiley, 1992; Smith and Kraft,
2005).
The importance of biological connections along river networks is often highlighted by human alterations
that affect these connections. For example, fish assemblages within highly connected river networks
were more homogeneous, whereas fragmentation by road crossings resulted in greater dissimilarity of
fish assemblages between upstream and downstream habitats (Perkin and Gido, 2012). Many studies
have documented statistically significant associations between impoundment of prairie streams and loss
of native fishes (e.g., Winston et al., 1991; Luttrell et al., 1999; Schrank et al., 2001; Falke and Gido, 2006;
Matthews and Marsh-Matthews, 2007), and fragmentation of river networks has been consistently
related to local extinction of salmonid populations (Morita and Yamamoto, 2002; Letcher et al., 2007).
For certain taxa, headwater streams—including intermittent and ephemeral streams— provide critical
habitat for specific portions of their life cycles. Many fish, both salmonids and nonsalmonids, spawn in
headwater streams, including those with intermittent flow (Erman and Hawthorne, 1976; Schrank and
Rahel, 2004; Ebersole etal., 2006; Wigington etal., 2006; Colvin etal., 2009). Kanno etal. (2014) found
that many brook trout moved between mainstem and tributary habitats over their life cycles. Because
reproductive success varied across these habitats, this movement resulted in substantial gene
movement into tributary habitats (Section 3.5.3).
After spawning, fish using headwater streams return downstream for feeding and overwintering. For
example, Bonneville cutthroat trout moved from less than 1 km to more than 80 km downstream
postspawning, typically within 30 days (Schrank and Rahel, 2004). Many salmonids also grow in
headwater streams (Brown and Hartman, 1988; Curry et al., 1997; Bramblett et al., 2002). In some
cases, these headwater streams, including intermittent streams, can provide higher quality habitat for
juvenile fish, as evidenced by increased growth, size, and overwinter survival in these habitats (Ebersole
etal., 2006; Ebersole etal., 2009), perhaps due to warmer temperatures and higher prey and lower
predator densities (Limm and Marchetti, 2009).
In prairie streams (Section B.4), the importance of hydrologic connectivity for biological connectivity is
especially evident, as many fishes broadcast spawn, or release eggs into the water column, which then
develop as they are transported downstream (Cross and Moss, 1987; Fausch and Bestgen, 1997).
Platania and Altenbach (1998) estimated that unimpeded eggs could travel as far as 144 km before
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hatching, and another 216 km as developing protolarvae (i.e., the swim-up stage), illustrating that
downstream transport of these drifting organisms can be extensive. Adult fish then migrate upstream
prior to egg release (Fausch and Bestgen, 1997). Thus, these fishes require hydrologic connectivity to
maintain both upstream and downstream populations (Fausch and Bestgen, 1997).
When abiotic or biotic conditions farther downstream in the river network are adverse, upstream
reaches can provide refuge habitat for downstream fishes. Examples of adverse abiotic conditions
include temperature (Curry et al., 1997; Cairns et al., 2005) and flow (Pires et al., 1999; Wigington et al.,
2006) extremes, low dissolved oxygen concentrations (Bradford et al., 2001), and high sediment levels
(Scrivener et al., 1994). Examples of adverse biotic conditions include the presence of predators,
parasites, and competitors (Fraser et al., 1995; Cairns et al., 2005; Woodford and Mclntosh, 2010).
Because headwater streams often depend on ground-water inputs, temperatures in these ecosystems
tend to be warmer in winter (when ground water is warmer than ambient temperatures) and colder in
summer (when ground water is colder than ambient temperatures), relative to reaches farther
downstream (Section 3.3.4; Power etal., 1999). Thus, these headwater streams can provide organisms
with both warmwater and coldwater refuges at different times of the year (Curry et al., 1997; Baxter and
Hauer, 2000; Labbe and Fausch, 2000; Bradford etal., 2001), again highlighting the spatial and temporal
variability of these fish-based biological connections. In some cases, loss of coolwater refuges can
facilitate invasion by species more tolerant of warmwater conditions (Karr et al., 1985).
Headwater streams also can provide refuge from flow extremes. Fish can move into headwater streams,
including intermittent streams, to avoid high flows downstream (Wigington et al., 2006); fish also can
move downstream during peak flows (Sedell et al., 1990), highlighting the bidirectionality of biological
connections within these systems. Low flows can cause adverse conditions for organisms, as well, and
residual pools that are often fed by hyporheic flow can enable organisms to survive dry periods within
intermittent streams (Pires et al., 1999; May and Lee, 2004; Wigington et al., 2006).
Biotic conditions within the river network—that is, the taxa found in the system—also can create an
adverse environment, as the presence of invasive species or other predators and competitors can
negatively affect native taxa. In some cases, headwater streams can provide these taxa refuge from other
species and enable populations to persist. For example, Fraser et al. (1995) found that prey fish moved
downstream when piscivores (fish-eating fish) were excluded, but moved upstream into headwater
streams when they were present. The role of headwater streams as refuges from adverse biotic
conditions can be closely related to where along the connectivity-isolation continuum these habitats fall,
with isolation allowing for persistence of native populations (Letcher et al., 2007). Physical barriers
(which reduce connectivity and increase isolation) have been used to protect headwater streams from
invasion (Middleton and Liittschwager, 1994; Freeman et al., 2007); similarly, most genetically pure
cutthroat trout populations are confined to small, high-elevation streams that are naturally or
anthropogenicallyisolated (Cook etal., 2010).
When adverse conditions have abated and these organisms move back down the river network, they can
serve as colonists of downstream reaches (Meyer and Wallace, 2001). For example, Hanfling and
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Weetman (2006) examined the genetic structure of river sculpin and found that upstream populations
were emigration biased (i.e., predominant movements were out of these reaches), whereas downstream
populations were immigration biased (i.e., predominant movements were into these reaches).
3.5.3 Genes
Genetic connectivity results from biotic dispersal and subsequent reproduction and gene flow, or the
transfer of genetic material within and among spatially subdivided populations. Populations connected
by gene flow have a larger breeding population size, making them less prone to inbreeding and more
likely to retain genetic diversity or variation—a basic requirement for adaptation to environmental
change (Lande and Shannon, 1996). Genetic connectivity exists at multiple spatial and temporal scales. It
can extend beyond a single river watershed (Hughes etal., 2009; Anderson etal., 2010), and in
diapausing organisms, can provide a direct link between distant generations (dispersal through time;
Bohonak and Jenkins, 2003).
Although physical barriers can protect headwater habitats and populations by isolating them from
colonization by and hybridization with invasive species (Section 2.3.2.1), isolation also can have serious
adverse effects on native species via reductions in genetic connectivity. For example, Hanfling and
Weetman (2006) found that artificial weirs intensified natural patterns of limited headwater
immigration, such that headwater (above-barrier) sculpin populations diverged genetically from
downstream (below-barrier) populations and lost significant amounts of genetic diversity. This pattern
of strong genetic divergence accompanied by loss of headwater genetic diversity above natural and
artificial barriers has been documented in multiple fish species and regions (Yamamoto et al., 2004;
Wofford et al., 2005; Deiner et al., 2007; Guy et al., 2008; Gomez-Uchida et al., 2009; Whiteley et al.,
2010). Loss of headwater-river genetic connectivity might be exerting selection pressure against
migrant forms in fish with life cycles requiring movement along the entire river corridor (Morita and
Yamamoto, 2002). Ultimately, tradeoffs exist between the risks associated with headwater-river genetic
connectivity (e.g., hybridization with nonnative species and hatchery fish) and those associated with
genetic isolation (e.g., reduced reproductive fitness, increased risk of local extinction, deterioration of
overall genetic variation, and selection against migratory traits; Fausch et al., 2009).
In general, genetic connectivity decreases with increasing spatial distance (Wright, 1943). Genetic
connectivity in river networks is also strongly influenced by the hierarchical structure of a river network
(Section 2.4.2), the direction of dispersal (upstream, downstream, or both), dispersal modes and
pathways used (e.g., swimming, flying), and species' life histories (Hudy et al., 2010).
Computer simulation approaches examine the spatial and temporal processes of genetic connectivity for
realistic behaviors and life histories of species inhabiting complex, dynamic landscapes and riverscapes
(Epperson et al., 2010). For example, Morrissey and de Kerckhove (2009) demonstrated that
downstream-biased dispersal in dendritic river networks (which by definition have more tributaries
than mainstems) can promote higher levels of genetic diversity than other geographical habitat
structures. Under these conditions, low-dispersing headwater stream populations can act as reservoirs
of unique genetic alleles (units of genetic variation) that occasionally flow into and mix with highly
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dispersing downstream populations. Although the number of headwater streams (i.e., potentially unique
genetic reservoirs) is important in maintaining genetic diversity, networks with more complex
hierarchical structures (Figure 3-4) are more efficient at maintaining genetic diversity than networks in
which all tributaries flow directly into the mainstem (Morrissey and de Kerckhove, 2009). In another
simulation, Chaput-Bardy et al. (2009) demonstrated that out-of-network gene flow (e.g., terrestrial
dispersal by insects or amphibians) or very high levels of within-network gene flow (e.g., fish that move
and reproduce throughout the network) can counteract the effects of network structure; thus, individual
species behavior can profoundly affect observed genetic patterns.
Figure 3-4. (A) A dendritic network with multilevel hierarchical structure, and (B) a uninodal
network with all headwater streams feeding directly into a river mainstem. Source: Reprinted from
The maintenance of genetic variation due to asymmetric gene flow in dendritic metapopulations,
(2009) by Morrissey and de Kerckhove with permission of The Univ of Chicago Press.
Most empirical evidence for the role of headwater streams in maintaining genetic connectivity and
diversity comes from studies of economically important fish species, but correlations of river network
structure or landscape alteration with genetic patterns have been reported for other species. Consistent
with the model of Morrissey and de Kerckhove (2009), Per and Hroudova (2008) found higher genetic
diversity in downstream populations of yellow pond-lily (Nuphar luted), which disperses over long
distances via water-mediated dispersal of detached rhizomes. Frequent dispersal and high gene flow
among headwater and downstream populations of the giant Idaho salamander (Dicamptodon aterrimus;
Mullen et al., 2010) are expected to contribute to genetic diversity of upstream and downstream
populations.
Headwater populations contribute to the maintenance of genetic diversity even in animals capable of
overland dispersal. In a field study of the common stream mayfly Ephemerella invaria, which emerges
into streamside forests to mate and disperse, Alexander et al. (2011) found that regional genetic
diversity was strongly correlated with tree cover in first-order (headwater) stream watersheds.
Observed loss of genetic diversity in this species could be related to degradation of stream habitats,
degradation of out-of-network dispersal path ways, or both (Chaput-Bardy et al., 2009; Grant etal., 2010;
Alexander etal., 2011).
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In summary, genetic connectivity in river systems reflects the breeding potential of a metapopulation.
The maintenance of genetic diversity is directly related to genetic connectivity, and thus is critical to a
species' regional persistence. Genetic connectivity is influenced by the landscape, riverscape, and
biology of the organisms involved; spatially subdivided stream and river populations can maintain
genetic diversity, provided they remain connected by at least low levels of gene flow (Waples, 2010).
3.6 Streams: Synthesis and Implications
Despite widespread human alterations, rivers are not simple conduits draining watersheds. A river,
including the water and material it carries and the organisms living in it, represents the cumulative
longitudinal, lateral, and vertical connections of its network of channels integrated over time (Section
1.2.3). Although we recognize that streams also exchange water and other materials with nearby
terrestrial and deep ground-water systems via lateral and vertical connections, this chapter focused on
longitudinal surface-water connections between streams and rivers, as well as shallow subsurface-water
interactions integral to surface-water connections and downstream water condition.
A substantial body of evidence unequivocally demonstrates connectivity between streams and
downstream rivers via both structural and functional connectivity (as defined in Wainwright et al.,
2011). Streams are structurally connected to rivers through the network of continuous channels (beds
and banks) that make these systems physically contiguous, and the very existence of a continuous bed
and bank structure provides strong geomorphologic evidence for connectivity (Section 2.2.1). A stream
must be linked to a larger, downstream water body by a channel for the two to have a surface-water
(hydrologic) connection. Although some streams lack a channel connection to larger water bodies (i.e.,
small endorheic basins), they are the exception. Streams that link larger water bodies through networks
of continuous bed and bank are the rule. The network structure reflects the aggregate and cumulative
nature of the connections between distant headwater streams and the downstream river.
Although not comprehensive or equally studied among all stream types, the existing science indicates
that connectivity with downstream waters varies among streams and over time. This variation in
connectivity to downstream waters can be described as a connectivity gradient, ranging from highly
connected to highly isolated (Section 1.2.2). A stream's position on the gradient is influenced not only by
distance to downstream waters but also by the frequency, magnitude, duration, timing, and rate of
change effluxes to downstream waters. Connectivity is dynamic: It changes with immediate, seasonal,
and interannual or interdecadal (e.g., climate oscillations) conditions that affect the availability and
distribution of water, materials, and biota. Because connectivity is dynamic, a complete understanding
of a stream's connections and consequences to downstream waters should aggregate connections over
relatively long time scales (multiple years to decades; Section 1.2.3). Although distance between streams
and downstream waters vary, other factors such as intervening resistance, relative size or chemical load,
and species assemblage also influence the degree of connectivity with and level of consequence on
downstream waters. Despite being distant from downstream waters, headwater streams make up the
majority of stream channels in most river networks and cumulatively supply most of the water in rivers.
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Streams are functionally connected to rivers by the movement of water and other materials through this
network of channels (Table 3-1). The longitudinal, vertical, and lateral connections within river
networks are inextricably tied. Even losing-stream reaches that at times lack sufficient flow for
hydrologic connection can still influence downstream waters by functioning as sinks for water and
materials carried by water. The river network and its flow of materials represent the integration of its
streams' cumulative contributions to downstream waters. Existing evidence indicates that headwater
streams (including intermittent and ephemeral streams) transform, store, and export significant
amounts of material (e.g., water, organic matter, organisms) to downstream waters. The most
compelling evidence linking headwater streams to downstream habitats supports source, sink (or lag),
and transformation functions (Section 2.3.1; Table 2-1). For example, studies that involved sampling
throughout river networks have documented headwater streams as sources of water (via floods and
baseflow) to rivers (Section 3.3.1). Nitrogen and carbon transported from headwater streams
cumulatively contribute to nitrogen and carbon levels in downstream rivers, and headwater streams can
function as nitrogen and carbon sinks for river networks (Sections 3.4.1 and 3.4.2). Studies documenting
the fate and transport of contaminants through headwater streams to downstream waters also
represent clear lines of evidence for headwater streams as sources and sinks (Section 3.4.4). Many
organisms, such as anadromous salmon, have complex life cycles that involve migration through the
river network, from headwater streams to downstream rivers and oceans, over the course of their lives
(Section 3.5). In fact, the importance of headwater streams (including intermittent and ephemeral
streams) in the life cycles of many organisms capable of moving throughout river networks provides
strong evidence for connectivity among these systems.
Most of the evidence relevant to issues of connectivity between headwater streams and large rivers is
based on data collected either in the upper (i.e., from headwater streams to intermediate tributaries) or
lower (i.e., from large tributaries to mainstem rivers) portions of the river network. Although few
studies have explicitly examined the movement of materials along entire river networks, the exchange of
materials among closely located stream reaches—which numerous studies have documented, for a
variety of materials—can be extended over large spatial scales.
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Table 3-1. Examples of mechanisms by which streams are connected to and influence downstream
waters, by functional type. See relevant section and appendix numbers in parentheses for greater detail.
Note that the distinction between types of functions is not always clear. For example, denitrification can
be considered a sink or transformation function. Bold letters represent the primary type of connection (B
= biological; C = chemical; and P = physical).
Source Function
Streams supply water downstream through baseflow and floods that influence discharge and habitat (3.3.1,
B.4.2.5, B.4.3.1.1, B.5.3, B.5.4.2, B.5.5.1). P
Streams supply downstream waters with sediment (3.3.2, 3.4.4, B.4.3.1.3, B.5.3, B.5.4.2). P
Streams supply downstream waters with nutrients and other ions (3.4.1, 3.4.3, B. 4.3.2.1, B.5.4.2). C
Streams can transport to downstream waters contaminants and pathogens that adversely affect organisms
and human health (3.4.4, B.4.3.1.3). C
Streams supply dissolved and particulate organic matter that can fuel heterotrophy in downstream waters
and influence physicochemical conditions (3.3.3, 3.4.2, B.4.3.2.2, B.5.4.2). C
Organisms actively and passively move between streams and downstream waters, carrying with them
nutrients, contaminants, pathogens, and other organisms (3.5, B.4.2.4, B.4.3.3). B
Organisms can enhance the supply of materials to downstream waters (3.5.1, 3.5.2). B
Sink Function
Streams can divert surf ace flow from downstream waters via infiltration into underlying alluvium and
evapotranspiration to the atmosphere (3.3.1, B.5.3, B.5.4.2, B.5.5.1). P
Streams can divert nitrate from downstream waters via denitrification (3.4.1, B.4.3.2.1). C
Streams can prevent sediment and associated contaminants from being transported to downstream waters
through deposition on floodplains (3.3.2, 3.4.4, B.5.3). C
Refuge Function
Streams can afford protection from temperature extremes, drying, predators, and competition with
nonnative species for organisms that inhabit downstream waters (3.5, B.4.3.3). B
Transformation Function
Streams can mediate the form and mobility of nutrients before they enter downstream waters via nutrient
spiraling (3.4.1, B.4.3.2.1). C
Streams can mediate the form and mobility of organic matter before they enter downstream waters via
carbon spiraling (3.4.2, B.4.3.2.2). C
Streams can mediate the form and mobility of contaminants before they enter downstream waters via
hyporheic exchange or exposure to other physicochemical gradients that lead to biogeochemical
transformations (3.4.4). C
Organisms can mediate the transformation of materials through their trophic and physical activities (3.4.1,
3.4.2,3.5.1,3.5.2, B.4.3.2.2). B
Lag Function
Streams can delay water from arriving at downstream waters through local and network structures, thus
reducingflood magnitudes, but increasing baseflows in downstream waters (3.3.1, 3.3.3, B.4.3.1.1, B.5.3,
B.5.4.2). P
Streams can delay sediment from arriving at downstream waters through local and network structures
(3.3.2, 3.3.3, 3.4.4, B.5.3). P
Streams can delay nutrients from arriving at downstream waters through local and network structures and
biological uptake (3.4.1, B.4.2.4, B.4.3.2.1). C
Streams can delay organic matter from arriving at downstream waters through local and network structures
and biological uptake (3.3.3, 3.4.2, B.4.3.2.2). C
Streams can delay contaminants from arriving at downstream waters through local and network structures
and exchanges that enhance mineralization and precipitation or adsorption to sediment, or both (3.4.4). C
Organisms can delay nutrients, organic matter, and contaminants from arriving at downstream waters
through consumption, assimilation, and bioconsolidation (3.4.1, 3.4.2, 3.5.1, 3.5.2, B.4.3.2.2). B
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CHAPTER 4. WETLANDS: PHYSICAL, CHEMICAL,
AND BIOLOGICAL CONNECTIONS TO RIVERS
4.1 Abstract
Wetlands are transitional ecosystems that occur between terrestrial and aquatic systems. They are
inundated or saturated by water at a frequency and duration sufficient to support hydrophytic
vegetation and development of hydric soils. The effects of wetlands on rivers and other downstream
waters depend on functions within the wetlands and connectivity between wetlands and downstream
waters. Riparian/floodplain wetlands can be hydrologically connected to streams and rivers through
unidirectional flows (i.e., from wetlands to rivers and streams, but not vice versa) of surface water and
ground water from upgradient areas (e.g., hillslopes and nearby uplands). In addition,
riparian/floodplain wetlands have bidirectional connections to streams and rivers (i.e., from wetlands to
streams and rivers and vice versa) through lateral movement of surface and ground water between the
channel and riparian/floodplain areas. Connections between riparian/floodplain wetlands and streams
or rivers occur over a gradient of connectivity, for example, they can be permanent, can occur frequently
(e.g., if the wetland is located within the mean high-water mark), or can occur infrequently (e.g., if the
wetland occurs near the edge of the floodplain; Sections 1.2.2 and 2.4.2). Even riparian/floodplain
wetlands that rarely flood can have important, long-lasting effects on streams and rivers.
Riparian/floodplain wetlands can reduce flood peaks by storing floodwaters, store large amounts of
sediment and nutrients from upland areas, influence stream geomorphology by providing woody debris
and sediment, and regulate stream temperature. Riparian/floodplain wetlands also are sources of food
for stream and river invertebrates and serve as rearing habitat for fish.
Wetlands in non-floodplain landscape settings lack bidirectional hydrologic connections with channels
(i.e., water flows from the wetland to the channel but not from the channel to the wetland). These
settings, however, have the potential for unidirectional hydrologic flows from wetlands to the river
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network through surface water or ground water. Non-floodplain wetlands can attenuate floods through
depressional storage and can recharge ground water and thereby contribute to baseflow. These
wetlands can affect nutrient delivery and improve water quality by functioning as sources (e.g., of
dissolved organic carbon) and as sinks for nutrients (e.g., nitrogen), metals, and pesticides. Non-
floodplain wetlands also can provide habitat or serve as sources of colonists for biological communities
in downstream waters, through movement of amphibians, reptiles, birds, and mammals. The extent to
which non-floodplain wetlands perform these functions depends on their hydrologic and biological
connectivity with downstream waters. Non-floodplain wetlands also occur on a hydrologic gradient,
from wetlands having permanent connections with perennial channels, to geographically isolated
wetlands having ground-water or occasional surface-water connections, to highly isolated wetlands
having minimal hydrologic connection to the river network (but which could include surface and
subsurface connections to other wetlands; Section 4.4.2). Non-floodplain wetlands that are connected to
the river network through a channel (i.e., wetlands that serve as stream origins) will have an effect on
downstream waters, regardless of whether the outflow is permanent, intermittent, or ephemeral. For
non-floodplain wetlands that do not connect to the river network through a stream channel (i.e.,
geographically isolated wetlands and wetlands that spill into losing streams that are completely
disconnected from the river network), the type and degree of connectivity with downstream waters will
vary with position in the watershed and over time.
This literature review is unable to provide evaluations of connectivity for specific groups or classes of
wetlands (e.g., prairie potholes or vernal pools). Evaluations of individual wetlands or groups of
wetlands, however, could be possible through case-by-case analysis. We can conclude the following:
1. A non-floodplain wetland having a surface-water outflow to a stream network (e.g., a wetland
that serves as a stream origin) is connected to the stream network and has an influence on
downstream waters.
2. Many non-floodplain wetlands interact with ground water, which can travel long distances and
affect downstream waters.
3. Even when wetlands lack a hydrologic connection to other water bodies, they can influence
downstream water through water and material storage and mitigation of peak flows (flood
reduction and flood attenuation). Sink functions of non-floodplain wetlands will have effects on
a downstream water when these wetlands are situated between the downstream water and
known point or nonpoint sources of pollution, thereby intersecting the flowpath between
pollutant source and downstream water. More generally, wetland sink functions are likely to be
greatest when the wetland is located downgradient from pollutant sources and upgradient from
a stream or river.
4. Within a watershed or wetland landscape setting, wetlands and open waters that are closer to
rivers and streams will have a higher probability of being connected than more distant areas,
assuming that conditions governing type and quantity of flows (e.g., slope, soil, and aquifer
permeability) are similar.
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5. Caution should be used in interpreting connectivity for wetlands that have been designated as
"geographically isolated."
4.2 Introduction
This chapter provides detailed information, based on a review of the pertinent peer-reviewed literature,
on how wetlands connect to and influence streams and rivers. In particular, we address two questions
(Section 1.1): (1) What are the connections to and effects of riparian/floodplain wetlands and open
waters (e.g., oxbow lakes) on downstream waters? (2) What are the connections to and effects of non-
floodplain wetlands and open waters on downstream waters?
In Chapter 1, we provided the scientific context for concepts and gradients of connectivity in hydrology
and ecology (Section 1.2). In Chapter 2, we provided definitions for wetlands, gave a rationale for
distinguishing between wetlands in riparian/floodplain and non-floodplain settings, and discussed
general hydrologic and biological mechanisms by which wetlands can connect to and affect streams and
rivers. Given that streams and rivers are the endpoints of interest, we limit our discussion of
riparian/floodplain wetlands to those occurring in riparian and floodplain settings. Below, we provide a
detailed review of the contributions of riparian/floodplain wetlands (Section 4.3) and non-floodplain
wetlands (Section 4.4) to rivers, followed by conclusions concerning these wetlands and their effects on
rivers (Section 4.5). Examples of some of the functions discussed in these two sections are found in
Table 4-1. In addition, four case studies on specific types of wetlands or lentic waters representing
different landscape settings and geographic regions are in Appendix B: Carolina and Delmarva bays
(Section B.I), oxbow lakes (Section B.2), prairie potholes (Section B.3), and vernal pools (Section B.6).
Much of the literature that we evaluate in this chapter does not specify the type or size of the stream or
river (or other water body) to which the wetland(s) are connected or which they influence. If available,
we note this information (e.g., whether riparian areas were located in floodplains or along portions of
river networks without floodplains), but often we can discuss only generic connections to streams,
rivers, or downstream waters. Given that rivers are connected to all upstream components of the river
network, including streams (Chapter 2), and the functional relationships between streams and rivers
(Chapter 3), however, we consider any evidence of connectivity with a stream (other than endorheic
streams; Sections 3.2 and B.5.5.1) to be evidence of connectivity with the river and other downstream
waters.
4.3 Riparian/Floodplain Wetlands
4.3.1 Introduction
This section focuses on the connections and influence of riparian/floodplain wetlands on downstream
waters. As previously defined in Section 2.2.1, riparian/floodplain wetlands are locations within riparian
areas and floodplains (Figures 1-1A, 2-2, and 2-3), respectively, that meet the Cowardin et al. (1979)
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Table 4-1. Examples of mechanisms by which riparian/floodplain wetlands and wetlands in non-
floodplain settings influence downstream waters, by functional type. See relevant section and appendix
numbers in parentheses for more detail. Note that the distinction between types of functions is not
always clear, for example, denitrification could be considered a sink or transformation function.
Source Function
Riparian/floodplain wetlands and non-floodplain wetlands connected to the stream network by channelized
flow—ranging from ephemeral to permanent—are sources of downstream water (4.3.2.1, 4.4.2.1, B.I.2.3,
B.2.3.1, B.3.3.1, B.6.3.1).
Wetlands that serve as origins for streams (e.g., seeps) can be sources of ground-water discharge,
contributing to stream baseflow (4.4.2.3).
Non-floodplain wetlands lacking a channel outlet can be sources of water via overland flow to the stream
network if wetland storage capacity is exceeded (4.4.2.1, B.3.3.1, B.6.3.1.1). They can also provide water
via subsurface drains ("tile drains") or surface ditches (4.4.2.1, B.I.3.1, B.3.3.1).
Riparian/floodplain wetlands and non-floodplain wetlands can be sources of nutrients and sediments to
downstream waters (4.3.2.2, 4.3.3, 4.4.3.1, B.l.3.2, B.3.3.2).
Riparian areas are a source of allochthonous inputs, the primary energy input into the food webs of small,
forested streams (4.3.3.4). They also are sources of woody debris that can affect stream morphology and
flow regime, and provide habitat for aquatic organisms (4.3.2.2).
Riparian areas and non-floodplain wetlands can be sources of dissolved organic matter that aquatic food
webs use, with additional potential effects on pH and mercury concentrations of downstream waters
(4.3.3.4, 4.3.3.6, 4.4.3.1).
Riparian/floodplain wetlands and non-floodplain wetlands can be sources of organisms, including plants,
invertebrates, amphibians, reptiles, and fish, to downstream waters transported via passive or active
dispersal (4.3.4, 4.4.4, B.2.3.3, B.3.3.3, B.6.3.2).
Riparian/floodplain wetlands can provide feeding habitat for riverine organisms, such as fish, during periods
of overbank flow (4.3.4.2, B.2.3.3).
Sink Function
Riparian/floodplain wetlands and non-floodplain wetlands can be sinks for water by intercepting overland or
subsurface flow, if available water storage capacity of the wetlands is not exceeded, which can reduce or
attenuate flow to downstream waters and flooding (4.3.2.1, 4.4.2.3, B.3.3.1).
Riparian areas and non-floodplain wetlands can be sinks for sediment and chemical contaminants, such as
pesticides, metals, mercury, and excess nutrients carried by overland or subsurface flow, potentially
reducing loading to downstream waters (4.3.2.2, 4.3.3, 4.4.3.2).
Riparian areas can be sinks for water, sediment, pesticides, and nutrients from overbank flow events,
reducing or attenuating downstream peak flows and materials entrained in the water column (4.3.2.1,
4.3.2.2, 4.3.3, B.2.3.2). They can also be sinks for seeds and plant fragments deposited via overbank flow
(4.3.4.1).
Riparian/floodplain wetlands and non-floodplain wetlands can be sinks for nitrogen by converting oxidized
forms of nitrogen to molecular nitrogen through denitrification, which is then lost to the atmosphere
(4.3.3.2, 4.4.3.2).
Refuge Function
Riparian/floodplain wetlands and non-floodplain wetlands can provide refuge for fish, aquatic insects, or
other lotic organisms from predators or other environmental stressors, facilitating individual or population
survival (4.3.4, 4.4.4).
Riparian/floodplain wetlands and non-floodplain wetlands can provide refuge during certain life stages for
lotic organisms. For example, they are breeding sites for frogs and other amphibians that reside in streams
as adults (4.4.4, B.I.3.3, B.6.3.2; Table 4-2); non-floodplain wetlands are additionally nesting and nursery
sites for American alligators that otherwise primarily reside in streams (4.4.4).
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Table 4-1. Examples of mechanisms by which riparian/floodplain wetlands and wetlands in non-
flood plain settings influence downstream waters, by functional type. See relevant section and
appendix numbers in parentheses for greater detail. Note that the distinction between types of
functions, is not always clear, for example, denitrification could be considered a sink or transformation
function (continued).
Transformation Function
Microbial communities in riparian/floodplain wetlands and non-floodplain wetlands can transform
elemental mercury to methylmercury before it enters a stream. Methylmercury is a particularly toxic and
mobile form that bioaccumulates in aquatic food webs (4.3.3.6, 4.4.3.1).
Riparian/floodplain wetlands and non-floodplain wetlands can transform nitrate to molecular nitrogen
through denitrification (4.3.3.2, 4.4.3.2).
Lag Function
Riparian/floodplain wetlands can temporarily store water following overbank flow, which then can move
back to the stream over time as baseflow (4.3.2.1).
Non-floodplain wetlands can contribute to ground-water recharge under low water table conditions, which
ultimately contributes to baseflow (4.4.2.3, B.3.3.1).
Non-floodplain wetlands can increase the time for stream discharge to rise and fall in response to a
precipitation event due to wetland storage capacity (4.4.2.3).
definition of having wetland hydrology, hydrophytic vegetation, or hydric soils. The terms "riparian
wetland" and "floodplain wetland" frequently describe the same geographic area. Because riparian areas
and floodplains also contain upland areas, some riparian/floodplain wetlands are geographically
isolated (i.e., completely surrounded by upland).
Although ample literature is available on riparian/floodplain wetlands—especially bottomland
hardwood and swamp wetlands—most papers on riparian areas and floodplains do not specify whether
the area is a wetland. This lack of specification occurs because riparian areas and floodplains also are
studied by stream ecologists and hydrologists who might not focus on whether their study site meets
the Cowardin et al. (1979) definition of a wetland. This situation creates a dilemma, because limiting our
literature review to papers that explicitly describe the area as a wetland would exclude a major portion
of this body of literature and greatly restrict our discussion of wetland science. Alternatively, if we
include papers that do not explicitly classify the area as a wetland, we could mistakenly incorporate
results that are relevant only to upland riparian areas. Our response to this dilemma was to survey the
floodplain and riparian literature broadly and include any results and conclusions that we judged
pertinent to riparian/floodplain wetlands. This judgment was based, in part, on: (1) the processes
described in the integrated systems perspective on interactions of watersheds, streams, wetlands and
downstream waters (Sections 2.2.2 and 2.2.3); (2) whether the information applies to all riparian areas,
regardless of whether they are wetlands or uplands (e.g., all riparian areas are subject to periodic
overbank flooding); and (3) an understanding of the specific processes. For example, riparian studies of
denitrification are likely to be either in a wetland or applicable to riparian/floodplain wetlands, because
the alternating oxidation/reduction conditions required for denitrification are present in wetlands.
Therefore, in our assessment of evidence regarding the connectivity and effects of riparian areas and
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floodplains, we have concluded that the processes and functions discussed occur in water bodies within
those areas.
As addressed in Chapter 2, much of the theory developed to explain how river systems function has
focused on linkages between system components (Vannote et al., 1980; Newbold et al., 1982a; Newbold
et al., 1982b; Junk et al., 1989; Ward, 1989; Power et al., 1995a; Power et al., 1995b; Huggenberger et al.,
1998; Ward, 1998; Fausch etal., 2002; Ward etal., 2002b; Wiens, 2002; Benda etal., 2004; Thorp etal.,
2006; Humphries et al., 2015). The integral connectivity between rivers and their floodplains and
riparian areas is a central tenet of stream hydrology and ecology, as is the substantial influence that this
bidirectional exchange has on the physical form, hydrology, chemistry, and biology of the river system
(Junk et al., 1989; Abbott et al., 2000; Tockner et al., 2000; Woessner, 2000; Amoros and Bornette, 2002;
Ward et al., 2002a; King et al., 2003; Naiman et al., 2005; Church, 2006; Kondolf et al., 2006; Poole et al.,
2006; Poole, 2010; Tockner et al., 2010; Vidon et al., 2010; Helton et al., 2011; Mclaughlin et al., 2011;
Humphries et al., 2015). For example, the flood pulse concept, which Junk et al. (1989) first articulated
and Tockner et al. (2000) extended, is a fundamental paradigm in river ecology, depicting the lateral
expansion and contraction of the river in its floodplain and the resulting exchange of matter and
organisms.
The influence of riparian/floodplain wetlands on downstream waters is especially notable because of
the potential magnitude and spatial extent of their interactions with rivers and their locations within
river networks. Although floodplains can form in modest size streams (Hughes and Lewin, 1982), they
typically form in the lower portion of river networks (Montgomery, 1999; Church, 2002, 2006), where
they can provide transient storage and subsequent release of river water and materials (Stanford and
Ward, 1993; Squillace, 1996; Mertes, 1997; Winter etal., 1998; Tockner etal., 2000; Fernald etal., 2001;
Amoros and Bornette, 2002; Malard etal., 2002; Claxton etal., 2003; Davis etal., 2011). Floodplain
patterns and river channel complexity are determined by sediment supply and character, river valley
slope, stream power, woody debris, and vegetation (Montgomery, 1999; Church, 2002; Coulthard, 2005;
Church, 2006; Osterkamp and Hupp, 2010; Sear etal., 2010; Collins etal., 2012). Circumstances
conducive to the formation of complex, rapidly changing channel forms (e.g., anastomosing, braided,
meandering) and the deposition of coarse sediment create conditions optimal for river-floodplain
interactions (Nanson and Croke, 1992; Mertes etal., 1995; Fernald etal., 2001; Fernald etal., 2006;
Poole et al., 2006; Whited et al., 2007).
Wetlands that occur in floodplains are referred to as riverine wetlands within the hydrogeomorphic
classification system (Smith et al., 1995). Although floodplain wetlands can occur as marshes (Villar et
al., 2001; Lee etal., 2005) or scrub-shrub wetlands (Chipps etal., 2006), these areas are known for
supporting forested wetlands. Mitsch and Gosselink (2007) classify floodplain forested wetlands as
freshwater swamps—for example, cypress-tupelo swamps (Taxodium distichum and Nyssa aquatica,
respectively) and white cedar swamps [Chamaecyparis thyoides)—if water is available throughout most
of the growing season, or as riparian ecosystems if the floodplain receives seasonal pulses of flooding.
Examples of the latter are bottomland hardwoods in the Southeast—for example, sycamore-sweetgum
(Platanus occidentalis and Liquidambar styraciflua, respectively) and cypress-tupelo forests—or
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cottonwood-willow (Populus spp. and Salix spp., respectively) and alder (Alnus spp.) riparian
communities in the Southwest (Mitsch and Gosselink, 2007).
This section provides further details on the connections between riparian/floodplain wetlands and
streams and rivers, and the resulting effects. Below, we examine the physical (Section 4.3.2), chemical
(Section 4.3.3), and biological (Section 4.3.4) effects of riparian/floodplain wetlands on rivers and other
downstream waters.
4.3.2 The Physical Influence of Riparian Areas on Streams
4.3.2.1 Hydrology
Riparian areas within and outside of floodplains are an important part of the overall riverine landscape
(Ward, 1998). Riparian areas are also connected to streams and rivers by a diverse set of hydrologic
inputs and outputs (Figure 2-6A; Junk etal., 1989; Winter and Rosenberry, 1998; Benke etal., 2000;
Tockner et al., 2000; Bunn et al., 2006). These inputs and outputs are described in Section 2.2 and have
been reviewed by various authors (National Research Council, 2002; Naiman etal., 2005; Vidon etal.,
2010).
Many studies document that riparian floodplains help attenuate flood pulses in streams and rivers by
capturing water from overbank flow and by storing excess water from streams (Mertes et al., 1995;
Poole et al., 2006; Rassam et al., 2006). Bullock and Acreman (2003) reviewed the wetland literature
and reported thatfloodplain wetlands reduced or delayed floods in 23 of 28 studies. Walton etal. (1996)
found that peak discharges between upstream and downstream water gages on the Cache River in
Arkansas were reduced 10-20%, primarily due to floodplain water storage. Gamble et al. (2007)
reported that 12 floodplain wetlands in Ohio stored an average of 3,654 m3 ha-1 of water. The authors
developed equations relating volume to area and depth for more than 650 regional wetlands and
reported that these systems could store approximately 1-2% of the daily flow of larger streams and
approximately 40% of the daily flow of small streams. As streamflow decreases after hydrologic events,
the water temporarily stored in riparian/floodplain areas can flow back into the channel, supporting
stream baseflow (Whiting and Pomeranets, 1997; Chen and Chen, 2003). Although not all
riparian/floodplain wetlands store the same amount of water, nearly all of them have the potential to
perform this function.
The potential for hydrologic connectivity between riparian/floodplain wetlands and rivers and streams
is high during periods of overbank flow and during periods of lower streamflow. Hyporheic exchange
occurs when water moves from river or stream channels into riparian or floodplain alluvial deposits and
back to the channels, and it occurs during flooded and non-flooded conditions (Sjodin et al., 2001;
Gooseff etal., 2008; Bencala, 2011) and on scales ranging from meters to kilometers (Stanford and
Ward, 1988; Bencala, 1993, 2005). Complex floodplains typically are environments with high levels of
hyporheic exchange (Woessner, 2000; Poole etal., 2006; Poole, 2010).
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Vegetation in riparian/floodplain wetlands can influence hyporheic and river water through
transpiration. Phreatophytes (plants that obtain their water from the saturated zone) can intercept
ground-water and overland flow before it enters a stream and decrease streamflow by directly taking up
stream water through their roots. For example, Meyboom (1964) studied two streams in the prairie
region of the United States to understand the effect of floodplain vegetation on streamflow fluctuations.
When the two streams decreased in flow, the floodplain vegetation accounted for 20% and 100% of this
reduction (Meyboom, 1964).
4.3.2.2 Geomorphology (Sediment-vegetation Interactions)
A bidirectional relationship exists between fluvial geomorphology and riparian and floodplain
vegetation (Corenblit et al., 2007). Distributions of vegetation communities often are shaped by river
flow dynamics and associated erosional and deposition processes, but the communities also exert
controls on geomorphic processes and riverine landforms.
Riparian/floodplain wetlands are key depositional environments for sediment that overland flow
carries from erosion of nearby uplands (Boto and Patrick, 1979; Whigham et al., 1988). Riparian areas
retain portions of this sediment before it enters the stream, especially if the overland flow enters the
riparian area as sheetflow runoff rather than as channelized flow, due to the greater volume of water
exposed to riparian-wetland soils and vegetation surfaces (Dabney et al., 1995; Meyer et al., 1995;
Naiman and Decamps, 1997; National Research Council, 2002; Naiman et al., 2005). Riparian open
waters (e.g., oxbow lakes; Section B.2) and wetlands are effective at retaining eroded clays, silts, and
sands that otherwise would enter stream channels (Cooper et al., 1987; Heimann and Roell, 2000).
Riparian areas were shown to remove 80-90% of sediments leaving agricultural fields in North Carolina
(Cooper et al., 1987; Daniels and Gilliam, 1996; Naiman and Decamps, 1997). Grassy riparian areas alone
can trap more than 50% of sediments from uplands when overland water flows are less than 5 cm deep
(Dillaha et al., 1989; Magette et al., 1989; Naiman and Decamps, 1997). Thus, riparian areas can buffer
stream channels against excessive sediment input.
Riparian areas and floodplains can be both sinks and sources for sediments in streams. When streams
flood their banks, increased surface contact and friction decrease the flow velocity. The slower moving
water has a diminished capacity for keeping material in the water column in suspension, which causes
the sediments to deposit (Church, 2002, 2006). Heavy particles such as sand are the first to be removed,
whereas finer, lighter particles such as clays and silts take longer to deposit. In southeastern Coastal
Plain systems, sediment deposition rates from the stream to the floodplain are high because of frequent
overbank flow and relatively high sediment loads of the rivers (Hupp, 2000).
Conversely, riparian areas and floodplains can also be a source of sediment to the stream, particularly
through streambank erosion. Although streambank erosion is a natural process, it can be accelerated
through vegetational changes because root tensile strength of riparian vegetation reinforces the soil
(Naiman and Decamps, 1997; Burt et al., 2002). Streambanks that are devoid of vegetation are often
highly susceptible to channel widening (Hupp et al., 1995; Naiman and Decamps, 1997). In a study of
748 bends in four southern British Columbia streams, for example, Beeson and Doyle (1995) reported
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that bank erosion was 30 times more prevalent on nonvegetated versus vegetated banks. In a
comparison of row-crop agricultural, grazing, and forested riparian areas in central Iowa, the forested
areas exhibited significantly reduced streambank erosion rates (Zaimes et al., 2004). Certain riparian
wetland vegetation types, such as black willow (Salix nigrd), maintain bank integrity and decrease
erosion so well that they are used in river restoration and bank stabilization projects (Pezeshki et al.,
2007).
Riparian vegetation also influences stream and river geomorphology through inputs of woody debris or
logs, which in turn shape stream channels (Brummer et al., 2006; Sear et al., 2010; Collins et al., 2012).
Woody debris can enter streams through tree mortality, bank undercutting, windthrow, wildfire, floods,
landslides, and debris flows (Gurnell et al., 2002; Reeves etal., 2003). Gurnell etal. (2002) reported that
the amount of woody debris deposited into streams can range from 12 to 40 t knr1 yr-1, depending on
the type of stream and nearby vegetation. As discussed in Section 3.3.3, woody debris can alter stream
channels, trap sediments, and form new aquatic habitat (Anderson and Sedell, 1979; Harmon et al.,
1986; Nakamura and Swanson, 1993; Abbe and Montgomery, 1996; Naiman and Decamps, 1997;
Gurnell etal., 2002).
4.3.2.3 Temperature and Sunlight
Riparian areas can modify stream temperatures and the amount of light available for photosynthesis in
stream and river environments through stream shading, particularly in forested settings (Barton et al.,
1985; Gregory et al., 1991; Blann et al., 2002). Dense, overhanging vegetation greatly reduces the
intensity of light, whereas open canopies allow light to penetrate (Gregory et al., 1991). This radiant
energy, or lack thereof, strongly influences stream temperature (Barton et al., 1985; Gregory et al., 1991;
Blann etal., 2002). The maximum temperature of a stream in Oregon, for example, was 7 °C higher in a
reach where the riparian vegetation was removed compared to its temperature when it was forested.
Fifteen years of regrowth in the harvested area was required for the stream temperature to return to
preharvest levels (Johnson and Jones, 2000).
By affecting stream temperatures, shading by riparian vegetation can alter fish growth, activity, and
mortality, while also influencing their prey species (Beschta et al., 1987). Higher temperatures, for
example, can lead to greater stream invertebrate biomass (Beschta et al., 1987). The net temperature
effect on fish growth, however, depends on the balance between food availability and higher metabolic
rates (Beschta et al., 1987). Riparian vegetation enhancement can be used by managers to promote fish
habitat for certain desired species. Blann etal. (2002) investigated the degree to which different types of
riparian vegetation could increase shade, reduce stream temperatures, and promote habitat for brook
trout (Salvelinus fontinalis) in Minnesota. The researchers concluded that both forested and herbaceous
riparian vegetation shaded the stream and buffered stream temperature, and could aid in creating
appropriate coldwater trout habitat (Blann et al., 2002).
Shading of the stream by riparian vegetation also directly influences the instream net primary
productivity of aquatic plants and other photosynthetic organisms, such as algae, by altering light
availability (Gregory et al., 1991). Net primary production is greatest in open reaches and is significantly
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less in reaches that are forested and shaded (Gregory et al., 1991). For example, Gregory et al. (1991)
reported that net primary productivity in open streams in Oregon averaged 210 mg carbon (C) nr2 d-1,
whereas forested reaches of streams with deciduous vegetation averaged 58 mg C nr2 d-1. Reduced net
primary production leads to lower densities of herbivores in streams (Hawkins and Sedell, 1981;
Gregory et al., 1991). Shading can limit stream productivity (Hill and Knight, 1988; Gregory et al., 1991),
but it can also be beneficial by reducing excessive algal production in nutrient-enriched waters. Algae
can lead to excessive biological oxygen demand and turbidity and can decrease water quality in
downstream systems (Volkmar and Dahlgren, 2006).
In addition to shading by riparian vegetation, riparian areas and floodplains can influence stream and
river water temperature through hyporheic exchange (Brosofske et al., 1997; Naiman and Decamps,
1997; Poole and Berman, 2001; Naiman et al., 2005). Hyporheic cooling of stream and river water
during warm summer periods has been observed in a wide range of settings, including large gravel bed
rivers in Oregon (Fernald etal., 2006; Burkholder etal., 2008; Seedangetal., 2008), an alpine stream in
the mountains of Colorado (Constantz, 1998), a boreal river in Sweden (Nyberg et al., 2008), and small
streams in Illinois (Peterson and Sickbert, 2006) and northern California (Loheide and Gorelick, 2006).
Important to note, however, is that hyporheic exchange can warm streams (Valett et al., 1990). Arscott
etal. (2001) found that hyporheic and other thermal regulating processes can lead to large thermal
heterogeneity of water bodies associated with complex floodplains. Hester and Gooseff (2010) argue
that, for streams impacted by human activities, restoration of hyporheic zones is essential for the
recovery of stream functions and ecosystem services.
4.3.3 The Chemical-nutrient Influence of Riparian Areas on Streams
Riparian areas in and outside of floodplains are instrumental in controlling the biogeochemistry of
riverine systems through (1) overbank flooding (flood pulse); (2) internal biogeochemical processes;
and (3) hyporheic exchange (Junk et al., 1989; Thurman et al., 1991; Heiler et al., 1995; Tockner et al.,
2000; Adair etal., 2004; Noe and Hupp, 2005; Valett etal., 2005; Noe and Hupp, 2007; Helton etal.,
2011; Powers etal., 2012; Bennett et al., 2015). All three mechanisms help shape nitrogen, carbon,
phosphorus, and pesticide cycling with the riverine environment.
Wetlands have been described as depositional areas in an eroding landscape (Brittain and Eikeland,
1988). Pollutants and materials relevant to discussions on water quality—such as nutrients, pesticides,
and metals—enter wetlands (e.g., Tiner, 2003c; Comer etal., 2005) through flowpaths that include dry
and wet (e.g., rain, snow) atmospheric deposition; point sources such as outfalls, pipes, and ditches; and
nonpoint sources, such as runoff from agricultural and urban fields and lawns, drift spray, and diffuse
near-surface water inputs (Nixon and Lee, 1986; Whigham and Jordan, 2003; Whitmire and Hamilton,
2008). For riparian/floodplain wetlands, transport from upstream reaches or through the hyporheic
zone (Figure 2-6) is another important source of these substances. Such materials can then be
sequestered via sorption (adsorption and absorption) or sedimentation processes, assimilated into the
flora and fauna, transformed into other compounds, or lost to the atmosphere through transformational
processes (Nixon and Lee, 1986; Johnston, 1991; Mitsch and Gosselink, 2007). These processes include
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conversion between particulate and dissolved forms of compounds via biologically mediated
degradation (e.g., Barlocher et al., 1978) and reduction-oxidation (redox) reactions (Nixon and Lee,
1986; Reddy and DeLaune, 2008). Redox reactions are essential to microbial respiration and are critical
to both defining wetland systems and understanding transformational processes that microbes mediate
(Boon, 2006; Reddy and DeLaune, 2008).
4.3.3.1 Hyporheic/Soil Processing of Nutrients
Riparian areas connect upland and aquatic environments through both surface and subsurface
hydrologic flowpaths (Figure 2-6; Naiman et al., 2005). Riparian areas act as buffers that are among the
most effective tools for mitigating nonpoint source pollution (Knight et al., 2010). These areas are
uniquely situated in watersheds to receive and process waters that pass through the root zone before
reaching streams (Gregory et al., 1991). These processes do not affect deep ground-water hydrologic
flowpaths (Figure 2-5) that enter a river or stream below the active riparian root zone. The focus of this
section, however, is on surface and shallow subsurface flows; we do not address deep ground-water
flowpaths here.
Riparian areas can significantly influence nutrients and other exports from watersheds (Gregory et al.,
1991) and can be considered areas of major nutrient transformation as subsurface waters move through
them (Dahm et al., 1998). Riparian areas remove nutrients such as nitrogen and phosphorus from water
as it flows from uplands to streams (Lowrance etal., 1997; Dosskey, 2001; Mayer etal., 2007). For
instance, Johnston (1993) reported that a floodplain wetland retained, 15.2,13.7, and 14.2% of the
solids, total nitrogen, and total phosphorus fluxes, respectively, from the watershed. The degree to
which a riparian area serves as either a source or a sink for nitrogen, phosphorus, organic matter,
pesticides, and mercury is controlled largely by the substance's concentration in riparian soils (Gregory
et al., 1991), soil redox conditions, and hydrology (Vidon et al., 2010). For example, riparian plant
communities can release seasonal pulses of dissolved leachates derived from stream litter (Fisher and
Likens, 1973). Riparian areas are therefore central to watershed water quality management (Burt, 1997;
Lowrance et al., 1997).
4.3.3.2 Nitrogen
Riparian areas can remove dissolved nitrogen (N) in subsurface flowpaths that would otherwise flow
into streams (Vidon et al., 2010). Removal occurs via plant uptake and microbial transformations (i.e.,
assimilative uptake, assimilatory nitrate reduction to ammonium, and dissimilatory nitrate reduction to
ammonium or nitrogen gases such as dinitrogen, nitric oxide, and nitrous oxide via denitrification). One
study demonstrated that intact riparian and hyporheic zones are critical in decreasing the amount of
dissolved inorganic nitrogen that moves from headwaters to larger, downstream waters (Triska et al.,
2007). Vidon et al. (2010) showed that riparian areas remove more than half the nitrogen from surface
and shallow subsurface water transporting ammonium and nitrate through the rhizosphere (Vidon et
al., 2010). Leaching from nitrogen-fixing plants (e.g., red alder, Alnus rubrd) in riparian systems,
however, also can be a major source of nitrogen to stream systems (Compton et al., 2003).
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Denitrification potential in surface and shallow subsurface flows is not homogeneous across the riparian
area, increasing markedly in the presence of organic carbon or anoxic conditions that create
denitrification "hot spots" (McClain et al., 2003; Orr et al., 2014). Therefore, for riparian areas to
appreciably increase nitrogen removal, flowpaths that convey nitrate-rich water into such
denitrification "hot spots" must be present (Vidon etal., 2010).
The highest denitrification potentials occur in floodplain systems where high organic matter levels,
denitrifying microbes, and saturated soil conditions are present (Vidon et al., 2010). Rates of
denitrification are greater in riparian soils nearer to streams (Gregory et al., 1991). Johnston (1993)
reported nitrate removal along a floodplain gradient of 6.6 g per 100-m distance from the stream. High
soil moisture and deposited organic matter enhance microbial activity, thereby tending to increase
denitrification (Reddy and DeLaune, 2008).
As subsurface flow passes through riparian areas, vegetative demand for dissolved nutrients also can
reduce nutrient loads (Vidon et al., 2010). More than three-quarters of the dissolved nitrate (NOs')
transported from agricultural fields to a Maryland river (Vidon et al., 2010) was removed by riparian
forests. Nitrogen was removed at a rate of 45 kg N ha-1 yr-1 as subsurface flow moved from agricultural
fields through riparian zones to nearby streams (Peterjohn and Correll, 1984). In the coastal plains of
Georgia, riparian forests retained more than 65% of the nitrogen and 30% of the phosphorus
contributed from nearby agriculture (Vidon et al., 2010). In southern Pennsylvania, a forested riparian
area had a subsurface nitrate budget with an average removal of 90 kg NOs-ha-1 yr-1, which was 26% of
the total nitrate input (Newbold et al., 2010).
4.3.3.3 Phosphorus
The movement and uptake of phosphorus in riparian areas are a function of phosphorus sources,
hydrology, and biogeochemistry (Vidon et al., 2010), with interactions between ground water and
surface waters driving the biogeochemical processes (Hoffmann et al., 2009). Phosphorus loss and
retention in riparian areas are related to the flowpath of the water through the riparian area to the
stream (e.g., overland flow of water from nearby agricultural fields, river-water inundation of floodplain
riparian areas). Flowpath dictates the confluence and interaction of phosphorus with minerals that drive
biogeochemical cycling of phosphorus in riparian areas (Hoffmann et al., 2009). The physical processes
of sedimentation and plant uptake are active in these flowpaths and can account for particulate
phosphorus retention rates as high as 128 kg P ha-1 yr-1 and 15 kg P ha-1 yr-1, respectively (Hoffmann et
al., 2009). Retention of dissolved phosphorus in riparian areas is more modest, with values less than 0.5
kg P ha-1 yr-1 often reported. Studies show, however, significantly higher numbers for the release of
dissolved phosphorus: up to 8 kg P ha-1 yr-1 (Hoffmann et al., 2009).
Although riparian soils generally serve as sources of phosphorus when soils are anoxic or when mineral
dissolution releases phosphorus (Baldwin and Mitchell, 2000; Chacon et al., 2008), riparian areas are
phosphorus sinks in oxic soils (Carlyle and Hill, 2001). Portions of riparian areas where agricultural
sediments are deposited are phosphorus sources to streams if the phosphorus is desorbed and leached
but can be sinks by adsorbing dissolved phosphorus if sediment phosphorus concentrations are low
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(Dillaha and Inamdar, 1997; Sharpley and Rekolainen, 1997). Riparian areas also serve as phosphorus
sinks when upland surface runoff travels through the riparian area or when fine-grained sediment
containing phosphorus is deposited overbank onto the riparian area (Dillaha and Inamdar, 1997). These
sediments, however, can become sources of phosphorus if they are later saturated with water and iron
and manganese are reductively dissolved during anoxic conditions, thus causing them to desorb
phosphorus (Reddy and DeLaune, 2008).
4.3.3.4 Carbon and Allochthonous Inputs
Both production and consumption of organic and inorganic carbon occur in riparian areas. In areas with
reducing conditions, microbes generally oxidize organic carbon and reduce available electron acceptors,
releasing carbon dioxide gas and making the soils more alkaline (Vidon et al., 2010). This process can
result in chemical gradients in which electron acceptor concentrations decrease and alkalinity increases
along subsurface flowpaths (Burns, 1996; Cirmo et al., 2000; Bailey Boomer and Bedford, 2008).
Riparian areas, especially those in low-lying flatlands, tend to have low subsurface flow velocities
resulting in anoxic conditions, shallow water tables, and slow organic matter decomposition, as is often
seen in riparian wetlands. This is why riparian areas are active areas for biogeochemical
transformations (Vidon etal., 2010).
Allochthonous inputs from riparian areas to streams are critical to aquatic food webs, particularly in
headwater catchments (reviewed in Tank et al., 2010). Allochthonous inputs are terrestrial organic
materials that enter the stream through vegetation litter (i.e., woody debris, leaves, and partially
decomposed plant parts), erosion, and hydrologic flows (Wetzel, 1992). In small forested watersheds,
overhanging trees provide organic matter inputs, while simultaneously reducing photosynthesis by
auto trophic organisms (Vannote et al., 1980). This dual effect makes allochthonous inputs the primary
source of energy flow into the food webs of these streams. For example, in a New Hampshire stream the
surrounding forest supplied more than 98% of the organic matter (Gregory et al., 1991). Organic matter
inputs are important because they affect food availability to aquatic organisms by releasing organic
carbon and nitrogen into streams (Wetzel and Manny, 1972; Mulholland and Hill, 1997). For example, in
a small headwater stream near Louisville, KY, macroinvertebrate communities, which are critical food
sources for fish (Wallace and Webster, 1996), relied almost exclusively on leaf inputs (Minshall, 1967).
Excluding litter from the riparian area changed the food web structure of a North Carolina stream
(Wallace et al., 1997) and decreased its dissolved organic carbon concentrations and loadings (Meyer et
al., 1998). In addition to the impacts of total inputs, the composition and timing of allochthonous inputs,
largely determined by riparian plant species composition, also can influence instream decomposition
and aquatic invertebrates (Cummins et al., 1989; Swan and Palmer, 2006).
Downstream, much less of the stream is directly influenced by streamside vegetation, due to larger
stream widths and consequently greater distances from the banks. This decreases the relative
importance of allochthonous inputs while concomitantly increasing the importance of instream
photosynthesis (Vannote et al., 1980). The macroinvertebrate community responds to this shift in input
types. For example, macroinvertebrate shredders that use large inputs, such as leaves, become less
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prevalent as streams increase in size. Besides changing longitudinally with stream size, riparian
allochthonous inputs also can vary seasonally, with a large pulse occurring in deciduous forests during
autumn leaf fall.
4.3.3.5 Pesticides
The roots in riparian areas can be important in removing pesticides from shallow subsurface flow,
because the labile organic matter and organic residues that accumulate near roots can increase
microbial biomass and activity (Vidon etal., 2010). Pesticides and their metabolites can be mineralized
and adsorbed where surface area contact is high and contact time with roots is sufficient (Krutz et al.,
2006). A study of the pesticides alachlor and atrazine in a riparian area notes the importance of plant
uptake in the fate of these pesticides, and suggests that vegetated buffer zones help protect water
supplies (Paterson and Schnoor, 1992). Studies examining specific pesticides—for example, isoproturon
(Benoit et al., 1999), metolachlor (Staddon et al., 2001), and atrazine (Mudd et al., 1995)-found that the
presence of vegetation, associated root zones, and accumulated organic matter increased the removal of
those pesticides (Vidon et al., 2010). Pesticide-degrading microbial populations increase after repeated
chemical applications (Gonod etal., 2006), suggesting that riparian areas can become better at
degrading pesticides that enter these zones (Vidon etal., 2010). In addition, microbial biomass has been
shown to be positively correlated with the loss of the herbicides 2,4-D (2,4-dichlorophenoxyacetic acid)
and dicamba, suggesting a relationship between the amount of microbial biomass in the soil and the
capacity of an ecosystem to degrade pesticides (Voos and Groffman, 1996).
4.3.3.6 Mercury
Mercury enters the global atmosphere primarily through waste incineration and coal combustion. It can
directly enter wetland systems or can be deposited on terrestrial areas and then transported into
riparian areas and wetlands via rainfall and runoff (St. Louis et al., 1994). Riparian soils and wetlands
are important both for mercury mobilization (Mierle and Ingram, 1991; Driscoll et al., 1995) and the
production of methylmercury, a particularly toxic and mobile form of the element. Mercury methylation
occurs in the presence of anoxic, saturated soils high in organic matter, mercury-methylating microbes,
and mercury from either atmospheric deposition or soils (St. Louis et al., 1996). The redox conditions
found in the presence of a fluctuating water table are thought to be a strong driver of mercury
methylation (Heyes etal., 2000; Branfireun and Roulet, 2002; Branfireun, 2004). Export of mercury and
methylmercury can expose organisms in downstream aquatic ecosystems to potential toxicity
(Thurman, 1985; Driscoll et al., 1995). Mercury bioaccumulates in fish, and consumption offish is the
main human pathway for exposure to mercury (Rypel et al., 2008).
The source-sink dynamics of riparian areas with respect to mercury are complex. Because soils
accumulate mercury, they buffer aquatic ecosystems against the full impact of this pollutant (Aastrup et
al., 1991). Because some of this mercury and methylmercury moves from soils to surface waters,
however, riparian areas might also be a source of the mercury that ends up in the aquatic food web.
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4.3.4 Biological Connections Between Riparian Areas and Streams
The dynamic nature of river systems is most apparent in riparian areas and floodplains, where a shifting
landscape mosaic supports diverse communities of aquatic, amphibious, and terrestrial plant and animal
species adapted to periodic or episodic inundation of riparian areas and floodplains (Power et al.,
1995a; Power et al., 1995b; Galat et al., 1998; Robinson et al., 2002; Toth and van der Valk, 2012;
Rooney et al., 2013; Granado and Henry, 2014). In unregulated rivers, floodplain inundation greatly
increases the area and diversity of aquatic habitats (Junketal., 1989; Tockner etal., 2000; Brooks and
Serfass, 2013). It also enables rapid cycling of nutrients imported from river channels (Section 4.3.3.2),
resulting in high primary productivity of plants and algae (Junk et al., 1989; Tockner et al., 1999). The
combination of diverse habitat types and abundant food resources makes floodplains important
foraging, hunting, and breeding sites for fish (Copp, 1989; Bestgen et al., 2000; Schramm and Eggleton,
2006; Sullivan and Watzin, 2009; Alford and Walker, 2013; Magana, 2013), aquatic life stages of
amphibians (Richardson etal., 2005), and aquatic invertebrates (Smocketal., 1992; Smock, 1994). Many
of these organisms have growth stages or reproductive cycles timed to coincide with seasonal
hydrologic connectivity between rivers and floodplains. Thus, lateral fluctuations in hydrologic
connectivity can increase overall levels of species productivity and biodiversity in river systems (Junk et
al., 1989) and can be integral to the viability of many riverine species (Bunn and Arthington, 2002).
Here, we review examples of adaptation to and exploitation of riparian habitats by aquatic species of
plants, fish, mammals, and invertebrates.
4.3.4.1 Vascular Plants and Phytoplankton
Channels and riparian/floodplain wetlands provide habitat for aquatic vegetation, emergent vegetation,
and phytoplankton. When seeds, plant fragments, or whole organisms move back and forth between
riparian/floodplain wetlands and the river network (via water, wind, or animal dispersal), these areas
become biologically connected. Species can disperse via overbank flow between channels and
riparian/floodplain wetlands (e.g., Schneider and Sharitz, 1988; Middleton, 2000; Nilsson et al., 2010).
Seeds from vegetation within the channel or that have been mobilized from upstream
riparian/floodplain wetlands can be deposited on bordering or downstream riparian areas and
floodplains (Nilsson et al., 2010), much like sediment and in many cases with sediment (Gurnell, 2007;
Gurnell et al., 2008). For example, in the southwestern United States, soil seed banks of wetland plants
can be established or replenished in floodplains when those areas are connected to a stream channel by
overbank flow (Boudell and Stromberg, 2008). In another example, 41% of plant species for which the
seeds were deposited on riparian areas during winter flood flow in two United Kingdom rivers were
wetland or aquatic plants (Gurnell et al., 2008). Overland flow or flooding also can dislodge viable plant
fragments in riparian/floodplain wetlands, which then are transported down the river network.
Fragments of seep monkeyflower [Mimulusguttatus) are easily dislodged by the relatively high flow
velocities along the riparian-channel interface, and fragments can survive and reestablish downstream
at rates exceeding 90% (Truscott et al., 2006).
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Floodplains can function as sinks for seeds and plant fragments. For example, in a forested floodplain
wetland in Illinois, many bald cypress (Taxodium distichuni) seeds dispersed by the river network were
deposited but did not germinate (Middleton, 2000). Alternatively, establishment and reproduction of
refuge floodplain populations can become important wetland seed sources for the river network,
especially if catastrophic flooding scours vegetation and seed banks that can exist on streambeds
(Gurnelletal., 2008).
Hydrologic connectivity between channels and riparian/floodplain wetlands can significantly enhance
riparian vegetation diversity (Jansson et al., 2005) and determine floodplain wetland community
structure (Boschilia et al., 2008). For nonnative species, however, connectivity can facilitate invasion,
resulting in changes in riparian vegetation community structure. In an intermittent stream in Illinois,
tubers of the nonnative Chinese yam (Dioscorea oppositifolid) were dispersed via stormflow and
overbank flow and became established along a narrow upstream riparian area and wider channel and
floodplain more than 1 km downstream; the presence of the nonnative plant significantly reduced native
plant cover (Thomas et al., 2006). Vegetation community composition, in turn, can affect the function of
riparian areas as nutrient sources or sinks to the river network (Sections 4.3.3.2 and 4.3.3.3). Invasion
by nonnative riparian plants also can result in altered stream invertebrate diversity, among other effects
(Lecerfetal., 2007).
Seeds of aquatic and riparian plants also can be actively dispersed by animals that consume them. For
example, seeds of the aquatic emergent bur-reed [Sparganium emersum) were ingested and viably
excreted by common carp [Cyprinus carpid) (Pollux et al., 2007), which elsewhere have been observed
using channel and floodplain wetland habitat (King et al., 2003). Riparian floodplain and wetland
vegetation can also disperse and exchange seeds via terrestrial animal vectors and the wind. Animals
that travel overland can also disperse ingested seeds or seeds adhering to fur, feathers, or limbs
between riparian/floodplain wetlands and the river network (see Sections 4.3.4.2, 4.4.4, and B.3.3.3 for
discussions of animal movement). Many macrophyte species have evolved for dispersal by wind,
including some of the most invasive in North America, cattail and reed canary grass (Barrat-Segretain,
1996; Soons, 2006 and references therein). Given the proximity of riparian/floodplain wetlands and the
river network itself, dispersal of pollen and seeds between these habitats could be quite frequent. For
example, seeds of some 20 species found in floodplain wetlands in bald cypress swamps in Illinois were
caught in aerial seed traps, and dispersal of three species averaged more than 100 seeds nr2 yr-1
(Middleton, 2000).
Phytoplankton also move via water between floodplain wetlands and the river network. A river with
overbank flow can homogenize the phytoplankton communities in floodplain wetlands separated by
more than 5 km (Angeler et al., 2010), and phytoplankton communities in river networks can be
bolstered by high-productivity conditions in temporarily connected floodplain wetlands. For example, a
portion of flow from California's Sacramento River is seasonally diverted from the main channel into the
Yolo Bypass, a nearby 240 km2 floodplain. From January to June 2003,14 and 31% of total diatom and
total green algae biomass, respectively, was produced in the floodplain (Lehman et al., 2008). This
considerable contribution of carbon to the aquatic food web, which ultimately supports downstream
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fisheries, resulted from the high net primary productivity of the floodplain. This observation is
particularly noteworthy because the median flow through the floodplain during the period of
measurement (23m s~l] was just 3% of the median flow through the main channel. Considered
collectively, these studies indicate riparian/floodplain wetlands can be both sources and sinks for
phytoplankton and water-, animal-, and wind-dispersed vascular plants with respect to the river
network.
4.3.4.2 Vertebrates
Animals, including many fish and mammals, move between riparian/floodplain wetlands and the river
network. The evidence is strong and abundant that fish can move between the main river channel and
riparian/floodplain wetlands when the channel and wetlands are hydrologically connected, even when,
in some cases, the connection is seasonal or temporary. Such wetlands provide refuge, feeding, and
rearing habitat for many fish species and augment recruitment to the river network (Boltz and Stauffer,
1989); examples include fish taxa in forested floodplain wetlands of the southeastern and southwestern
United States and salmonids of the northwestern United States such as coho salmon (Oncorhynchus
kisutch) and Chinook salmon [Oncorhynchus tshawytschd) (e.g., Wharton et al., 1982; Matheney and
Rabeni, 1995; Pease etal., 2006; Henning et al., 2007; Jeffres etal., 2008). In one section of the mainstem
Rio Grande in New Mexico, more than 90% of the larval and juvenile fish of six captured species were
from riparian areas with zero water velocity (backwaters, former side channels, and isolated pools;
Pease etal., 2006). Oxbow lakes are also important habitats for fish feeding and rearing. Based on a 5-
year study of fish in oxbow lakes, Shoup and Wahl (2009) concluded that the entire floodplain should be
considered a single functioning unit that supports the overall biological integrity of a river (Section B.2).
The use of riparian/floodplain wetlands by fish depends on many factors intrinsic to the particular river
system (e.g., periodicity and duration of floodplain inundation) and the characteristics of the resident or
migratory fish community (King et al., 2003).
Fish also move between lacustrine wetlands (wetlands associated with lakes) and large lakes when
hydrologic connections exist. Fish communities in the Great Lakes and their surrounding wetlands
become more homogeneous when surface connections between the wetlands and lake are present. Fish
use these wetlands for refuge from predators and as rearing habitat (Jude and Pappas, 1992). Miyazono
et al. (2010), studying floodplain lakes in the Yazoo River Basin, found that conditions that included
decreases in habitat connectivity, wetland buffers, and certain water quality parameters led to the
increased dominance of environmentally tolerant fish in those lakes. Fish assemblages in riparian
wetlands along the semiarid region of the Murray River, Australia showed a large decline in diversity
when those wetlands were disconnected from the river through hydrologic modifications. This trend
was reversed after a managed inundation treatment restored connections between the wetlands and the
river (Vilizzi et al., 2013). River-dwelling mammals also move between rivers and riparian/floodplain
wetlands, including river otters, which have been observed using wetlands extensively as latrines
(Newman and Griffin, 1994). In addition, both river otters and beavers have a strong preference for
riparian areas that are pond- and lake-dominated (Swimley et al., 1999). Thus, movement of animals,
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especially fish, connects riparian/floodplain wetlands to the river network and supplies streams and
rivers with a source of biological materials.
In addition to acting as sources, sinks, and refuges for individual species of organisms,
riparian/floodplain wetlands can improve the overall health of biological communities. For example, a
positive relationship between wetland cover and an index of biological integrity for fish communities in
rivers was observed in 23 sites in several small catchments of the River Raisin in Michigan (Roth et al.,
1996).
Besides providing a form of biological connectivity that can link riparian/floodplain wetlands and
downstream waters, vertebrates in riparian areas can affect stream characteristics and influence various
forms of connectivity. Perhaps the most familiar example of this is the beaver (Castor canadensis).
Although beaver damming would be expected to reduce hydrologic connectivity through impoundment,
their influence can be more complex. For example, Westbrook et al. (2006) found that beaver dams in
the Colorado River affected depth, extent, and duration of inundation resulting from a 10-year flood
event. In addition, beaver dams attenuated declines in water tables during drier summer periods in 25%
of their 58 ha study area. They concluded that the main hydrologic effects occurred downstream,
however, rather than near the dam (Westbrook et al., 2006). The hydraulic head generated by the dam
raised the water level above the banks, resulting in lateral and downstream spreading of flows during
high- and low-flow periods; these effects extended over hundreds of meters. For example, mottled soils
occurred throughout the study area, suggesting that the dams caused waterlogged soils for extended
periods. Increased overbank flooding increases hydrologic connectivity between riparian areas and
streams. In contrast, when no dams were present, flooding was limited to the area immediately near the
stream channel. Beaver dams also can affect stream biogeochemistry. For example, beaver dams modify
nutrient cycling and decomposition dynamics and can affect downstream transport of materials
(Naiman et al., 1988; Naiman et al., 1994). For example, beaver-dam wetlands can serve as a source of
methylmercury (Roy etal., 2009). Beaver dams also can affect fish species, such as coho salmon (Pollock
etal.,2004).
Vertebrates also can indirectly affect hydrologic connectivity through cascading effects on riparian plant
communities. Beschta and Ripple (2012) provide evidence from analyses at three western National
Parks for a trophic cascade model where large predators can affect the morphology of river channels
through intermediate effects on ungulate browsers and riparian plant community structure. For
example, extirpation of wolves (Canis lupus) at Yellowstone National Park by the mid-192 Os led to an
increase in elk [Cervus canadensis] numbers. This increase caused suppression and mortality of riparian
willow (Salix spp.) communities, ultimately resulting in changes to stream morphology such as bank
erosion, decreased sinuosity, increased active channel width, and increased amount of unvegetated
alluvium (Beschta and Ripple, 2012). Based on results from the three National Parks and other sites,
Beschta and Ripple (2012) concluded that the removal of apex predators due to extirpation increased
ungulate herbivory, which altered riparian plant communities, thereby increasing bank erosion that led
to either widening of the active channel or channel incision. These channel alterations, in turn, reduced
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the frequency of overbank flows, which decreases hydrologic connectivity between the riparian area
and downstream waters.
4.3.4.3 Invertebrates
Stream macroinvertebrates (e.g., insects, crayfish, mollusks) and microinvertebrates (e.g., cladocerans,
copepods, rotifers, gastropods) colonize nutrient-rich riparian areas and floodplains in large numbers
during seasonal or episodic immersion by rivers and streams (Junk et al., 1989; Ilg et al., 2008).
Macroinvertebrates and microinvertebrates (also called zooplankton) are the intermediate link between
primary producers (e.g., algae), detrital pools (e.g., leaf litter), and predators (e.g., fish, amphibians) in
river food webs (Malmqvist, 2002; Woodward and Hildrew, 2002; Stead et al., 2005; Woodford and
Mclntosh, 2010). The distribution of invertebrate populations in dynamic river systems is governed by
the location of resources required for different needs and life stages, and invertebrates actively
dispersing to find and exploit resources wherever they become available (Malmqvist, 2002). As with
vascular plants, hydrologic connectivity between channels and riparian/floodplain wetlands can
significantly influence macroinvertebrate community structure in riparian areas (Paillex et al., 2009;
Yetter, 2013). For example, the species diversity and abundance of macroinvertebrates in the wetlands
of a river delta have been found to be positively correlated with a gradient of connectivity (Dou et al.,
2015).
Invertebrates have evolved two basic strategies to exploit habitats near streams and rivers: (1) rapid
colonization of flooded areas and short life cycles that complete before floodplains dry again, or (2) use
of aquatic refuges or dormant life stages to persist in permanent waters, the hyporheic zone, or
floodplain soils between inundations (Tronstad etal., 2007). To evaluate the relative importance of each
strategy in the same river system, Jenkins and Boulton (2003) compared the abundance and species
composition of microinvertebrates emerging from floodplain sediments to those transported by
floodwater from instream habitats at reach and watershed scales. Initially, most colonizers of newly
flooded riparian habitats came from distant upstream reaches of the river network, washed
downstream by floodwaters. After a few days, however, species hatching from eggs diapausing in soils
greatly increased the diversity and size of the river/floodplain community. This study illustrates two
important points about biological connectivity of river/riparian habitats:
1. Stream invertebrate communities comprise species adapted to different stresses in their
environment (in this case, resilient species adapted to high flows and resistant species adapted
to desiccation).
2. Floods that periodically connect different parts of the river network generate potential for gene
flow across time and space by mixing individuals from different locations (e.g.,
upstream/downstream, channel/floodplain) and different years (e.g., eggs that might have
diapaused for tens or even hundreds of years).
The findings by Jenkins and Boulton (2003), that resting egg banks in riparian soils are important to the
persistence of aquatic species and the composition of river communities, were validated in a separate
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study by Frisch and Threlkeld (2005), who compared flood-pulse colonization in a field study with
laboratory hatching of copepod microcrustaceans from egg banks of inundated soils in Mississippi. The
laboratory samples showed that, in the absence of hydrologic connections, egg banks were sufficient for
persistence of copepod populations; the field samples showed that when hydrologic connections were
present, water dispersal and hatching from dormant stages were both important colonization pathways
for copepods. In a perched floodplain in Missouri, Fisher and Willis (2000) showed that flood-pulsed
movement of water and organisms between river channels and floodplains was bidirectional.
Adaptations by stream-dwelling invertebrates to variable moisture conditions, and rapid two-way
dispersal to exploit temporary or seasonal hydrologic connections, are strong evidence of long-term
biological connectivity between rivers and riparian areas.
Invertebrates that disperse by aerial means also take advantage of flooded riparian habitats. Tronstad et
al. (2007) investigated aerial colonization of floodplains by insects during multiple flood pulses having
different inundation periods in an unregulated river in Alabama's Coastal Plain. At least 41 genera in 21
families across 7 orders of flight-capable insects colonized floating trays placed in floodplain waters in
June, August, November, and April. Insect densities varied across the period and reached a maximum in
August of about 80,000 individuals nr2, most of which were seeking mates or oviposition sites rather
than foraging or hunting. High densities (21,291 individuals nr2) of passively dispersing (e.g., via wind
or animal vectors) microcrustaceans also were observed. Vanschoenwinkel etal. (2009) erected 9
windsocks (sampling devices for aerially dispersing organisms) near temporary rock pools for 1 month,
during which 850 viable dormant eggs, larvae, and adults from 17 invertebrate taxa were collected.
Results from these studies illustrate that aerial dispersal of multiple taxonomic orders and phyla is a
significant source of stream invertebrate colonists in newly inundated floodplain habitats.
4.4 Non-floodplain Wetlands
4.4.1 Introduction
This section focuses on the connections and influence of non-floodplain wetlands (defined in Section
2.2.1) on downstream waters. Brinson (1993), in his hydrogeomorphic classification system,
categorized wetlands according to four geomorphic settings. This system subsequently was expanded to
the following seven classes by Smith et al. (1995): riverine, depressional, slope, mineral soil flats,
organic soil flats, estuarine fringe, and lacustrine fringe. Non-floodplain wetlands consist of certain
depressional, slope, and flats wetlands (although some of these wetlands can occur in riparian and
floodplain wetland settings; Section 2.2.1). Depressional wetlands, as their name suggests, occur in
topographic depressions and might or might not have a surface water inlet or outlet. Common types of
depressional wetlands include kettles, potholes, vernal pools, playa lakes, and Carolina bays (Brinson,
1993). Slope wetlands (also known as seeps) are located in breaks of slopes and are sites of ground-
water discharge (Hall et al., 2001a; O'Driscoll and DeWalle, 2010). Slope wetlands include fens, which
typically are ground-water driven and have diffuse outputs (Brinson, 1993; Bedford and Godwin, 2003).
Mineral soil flats commonly occur on interfluves, relic lake bottoms, or large floodplain terraces.
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Precipitation dominates the water sources in mineral soil flats, with little ground-water input. Wet pine
flatwoods and large playas are examples of this wetland type. Non-floodplain wetlands also include
organic soil flats. These contain extensive peatlands, or peatbogs, where the accumulation of partially
decayed organic matter dominates (Mitsch and Gosselink, 2007). Precipitation also generally dominates
the water inputs to bogs, which can connect to downstream waters via a channel outlet or diffuse
overland flow (Brinson, 1993). Bogs are generally more acidic than fens (Bedford and Godwin, 2003).
Depressional, slope, or flats wetlands also can serve as stream origins (Figure 2-18A).
Below, we examine the physical (Section 4.4.2), water quality (Section 4.4.3), and biological (Section
4.4.4) effects of non-floodplain wetlands on rivers and other downstream waters. We then briefly
consider the issue of geographic isolation in non-floodplain wetlands (Section 4.4.5).
4.4.2 The Physical Influence of Non-floodplain Wetlands on Streams
Section 2.4.1 provided a general description of how non-floodplain wetlands can connect to downstream
waters via surface and ground-water flow (Figure 2-18). In this section, we provide further details on
these connections and discuss how such connections affect streamflow.
4.4.2.1 Surface-water Connections
Non-floodplain wetlands can be connected by perennial surface flows to river networks. For example,
seeps are likely to have perennial connections to streams that provide important sources of baseflow,
particularly during summer (Morley et al., 2011). In a study in Maine, seeps were found to provide 40-
80% of stream water during baseflow periods (Morley et al., 2011). In other cases, surface connections
between non-floodplain wetlands and streams can be intermittent or ephemeral. Rains et al. (2008) and
Rains et al. (2006) showed that California vernal pools, situated on both clay and hardpan soils,
connected with streams through channels containing transient water flow (Section B.6). The series of
vernal pools on the clay soils were filled with water for 200 days of the year, and water spilled from
these wetlands through swales and channels for 60% of those days (Rains et al., 2008). McDonough et al.
(2015) found that forested Delmarvabays had seasonally intermittent surface water connections to
streams; these connections occurred during periods of low evapotranspiration and high water tables,
that is, from mid-fall to late-spring. In contrast, surface-water connectivity of restored and prior
converted (wetlands converted to agriculture before 1985) bays was ephemeral, that is, it occurred in
response to rainfall. The cumulative duration of connections to perennial streams was greater and had
fewer transitions between connected and disconnected states for forested bays than for restored and
prior converted bays (McDonough et al., 2015). Drainage of wetlands via ditching also can produce
surface water outflows from depressional wetlands directly to streams (Section 2.4.4); ditches, however,
also can introduce nutrients and ions into downstream waters (Brunet and Westbrook, 2012).
Even non-floodplain wetlands that are considered to be geographically isolated (i.e., completely
surrounded by uplands), can have surface-water outflows that connect them to other water bodies
(Figure 2-18B). Tiner (2003b) identifies vernal pools as 1 of 10 types of geographically isolated
wetlands. Yet, as just discussed, the studies by Rains et al. (2008) and Rains et al. (2006) indicate that
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vernal pools can be connected to stream networks by channels. As another example, a recent study of
depressional wetlands in the Texas Gulf Coast area showed that, although classified as geographically
isolated, these wetlands are actually connected to nearby waterways via intermittent streams (Wilcox et
al., 2011). During a study period of almost 4 years, nearly 20% of the precipitation that fell on a wetland
complex flowed as surface runoff through the stream to a nearby water body, the Armand Bayou
(Wilcox et al., 2011). Non-floodplain wetlands also can have temporary hydrologic connections to each
other. Such connections can occur through the expansion and contraction of surface water that occur
between wet and dry periods (e.g., Figure 2 in Niemuth etal., 2010) and through fill and spill of surface
waters. One consequence of fill-and-spill behavior is that the contributing area of such a wetland is
dynamic and has a nonlinear relationship to potential storage area (Shaw et al., 2012; Shaw et al., 2013).
In the intermontane West, evidence suggests that depressional wetlands can connect to one another via
temporary overland or shallow ground-water flows (Cook and Hauer, 2007). In the prairie pothole
region, temporary overland connectivity between potholes has been observed in wet years. In 1996,
during heavy spring rains, an estimated 28% of the wetlands in the study area had surface-water
connections to at least one other wetland (Leibowitz and Vining, 2003). Le and Kumar (2014) analyzed
topographic depressions in five study areas across the United States and found that hydrologic
connectivity—as determined by nearest neighbor distances—followed a universal power law
distribution. One implication of this distribution is that, although most depressions are connected over
short distances, a few are connected by long distances, which could cause rapid increases in hydrologic
connectivity as the system wets up (Le and Kumar, 2014). However, the distribution can be altered
through wetland drainage (Van Meter and Basu, In press). Although some of these studies focused on
wetland-to-wetland connections, the findings illustrate (1) the potential for geographically isolated
wetlands to exhibit temporary surface water connections with other water bodies, and (2) that
interacting wetland complexes might best be understood as a functional unit (Section 4.4.5).
4.4.2.2 Ground-water Connections
In addition to surface-water connections, ground-water flow can connect non-floodplain wetlands with
other water bodies, potentially over great distances (Figures 2-5 and 2-18C). Many studies have shown
that non-floodplain wetlands can connect to ground water, either receiving ground-water discharge
(flow of ground water to the wetland), contributing to ground-water recharge (flow of water from the
wetland to the ground water), or both (e.g., Lide et al., 1995; Devito et al., 1996; Matheney and Gerla,
1996; Rosenberry and Winter, 1997; Pyzoha etal., 2008). For example, a 1989 study of four North
Dakota prairie pothole wetlands by Arndtand Richardson (1989) clearly demonstrated ground-water
connections as one wetland recharged ground water, one was a flow-through wetland, and one was a
discharge system. Hunt et al. (2006) found that benthic invertebrate communities were correlated with
amounts of ground-water discharge to stream-wetland complexes in northern Wisconsin. Using stable
hydrogen and oxygen isotopes in water, Matheney and Gerla (1996) concluded that, although most of
the water in a depressional prairie wetland came from precipitation, ground-water connections
accounted for the high salinity of the wetland soil. The high salinity is indicative of net ground-water
discharge to the wetland (Brinson, 1993). Min et al. (2010) reported that 38% of rainfall that entered
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four historically geographically isolated wetlands in Florida was recharged to ground water. A literature
survey by Bullock and Acreman (2003) found 69 studies making reference to ground-water recharge
from wetlands; of these, 32 studies observed ground-water recharge from a wetland, whereas 18 studies
did not
Ground-water flow-through wetlands are sites of both ground-water discharge and recharge, in essence
a surface expression of the ground-water system (Richardson et al., 1992; Kehew et al., 1998; Ferone
and Devito, 2004). In these wetlands, ground-water discharge generally flows into the wetland on one
side or area, and flows back into the ground water on the other side or area of the wetland. This dynamic
has been shown in many locations, including prairie potholes (Richardson et al., 1992), wetlands in
glacially formed landscapes in southwest Michigan (Kehew et al., 1998), Alaskan ponds (Rains, 2011),
Florida cypress dome systems (Sun et al., 1995), and small Wisconsin lakes (Born et al., 1979). The lakes
and wetlands of the Nebraska Sand Hills are also predominantly flow-through and an expression of a
large regional ground-water system (Winter, 1999). The flow-through wetland influences the chemistry
of the transiting, shallow ground water. Kehew et al. (1998) found a wetland of this type diluted
nitrogen concentrations in the ground water of an agricultural watershed.
Whether a wetland recharges ground water, is a site of ground-water discharge, or both, is determined
by topography, geology, soil features, and seasonal position of the water table relative to the wetland.
Shedlock et al. (1993), for example, concluded that ground water discharged into a bog along Lake
Michigan through a breach in the sediments underlying the wetland. In dry periods when water tables
are low, water tends to move from wetlands into the ground water, while in wetter periods with higher
water tables, water can flow in the opposite direction from shallow ground water into the wetlands
(Phillips and Shedlock, 1993; Pyzoha etal., 2008; Mclaughlin etal., 2014). Lide etal. (1995) observed
both ground-water flow into and from a Carolina bay wetland, with discharge to the wetland when the
water table was high and recharge to the ground water when the water table was low. Sun et al. (1995)
observed similar phenomena in a Florida cypress dome. This exchange and temporary storage of water
represents a lag function that can make wetlands particularly important for ground-water recharge
during dry periods. Rosenberry and Winter (1997) indicated that ground-water discharge to a wetland
often alternates with flow from the wetland to ground water, and the direction of flow is controlled by
the balance of recent precipitation with current evapotranspiration demands.
The magnitude and transit time of ground-water flow from a wetland to other surface waters depends
on the intervening distance and the properties of the rock or unconsolidated sediments between the
water bodies (i.e., the hydraulic conductivity of the material). In some carbonate or volcanic rocks, for
example, ground water can flow relatively freely through large openings; while in unconsolidated
material—such as gravel, sand, silt, or clay—the spaces between particles determine the time required
for water to flow a given distance (Winter et al., 2003). In porous material, such as gravel, water can
travel a distance of a kilometer in a few days; in fine-textured materials, such as silt or clay, hundreds to
thousands of years might be required for a single parcel of water to travel the same distance (Winter
and LaBaugh, 2003).
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In agricultural regions, the transit time of subsurface flows can be decreased substantially by artificial
subsurface drainage pipes, known as tile drains (Section 2.4.4; Schiller et al., 2012). Wetlands in these
areas are sometimes fitted with inlets that connect directly to tile drains, quickly moving temporarily
ponded water through the subsurface and to outlets that discharge directly to ditches or streams
(Tomer etal., 2010).
In summary, non-floodplain wetlands can have a range of hydrologic connectivity with other waters
(Figure 2-18). Non-floodplain wetlands can be connected by permanent, intermittent, or ephemeral
surface flows through swales or channels, or be connected to other water bodies via shallow or deep
ground-water flows. Conversely, a wetland can be isolated hydrologically if it lacks surface water and
ground-water connections entirely and evapotranspiration is the dominant form of water loss. A
wetland also can be hydrologically isolated from streams and rivers if it recharges a ground-water
aquifer that does not feed surface waters. Wetlands that lack surface connectivity in a particular season
or year can be connected, nevertheless, in wetter seasons or years. A wetland that serves as the origin of
a stream will have a permanent or temporary surface water connection with a stream network through
a stream channel, unless the wetland feeds an endorheic stream (Sections 3.2 and B.5.5.1).
4.4.2.3 Effects of Non-floodplain Wetlands on Streamflow
Non-floodplain wetlands can affect streamflow by altering baseflow or stormflow (Section 2.2.2; Figure
2-8) through several mechanisms, including surface storage and ground-water recharge. Depressional
wetlands effectively store water because the aboveground portion of the wetland contains a largely
empty volume for water storage, in contrast to belowground water storage where only part of the
volume is available for water storage, for example, due to soil particles (i.e., the specific yield; Johnson,
1967; McLaughlin etal., 2014). Large-scale studies have shown that wetlands, by storing water, reduce
peak streamflows, and thus, downstream flooding. Hubbard and Linder (1986), for example, calculated
the water retention capacity of more than 200 closed depressional prairie potholes in northeastern
South Dakota. They observed that a large amount of snowmelt and precipitation could be cumulatively
held by many small wetlands, reducing the potential for flooding at downstream locations. Similarly, a
USGS study in the prairie pothole region found that wetlands—including both depressional and
nondepressional types—stored about 11-20% of the precipitation thatfell in a given watershed, and
that storage could be increased by wetland restoration (Gleason et al., 2007). Vining (2002) concluded
that wetland storage in the Starkweather Coulee subbasin of North Dakota likely resulted in decreased
streamflow. Rovansek et al. (1996) found snowmelt to be the most important source of water for
wetlands and ponds in the Alaskan Arctic Coastal Plain, and that these wetlands and ponds functioned as
surface storage, thereby removing water from the snowmelt floods. However, Ford and Bedford (1987)
note that in permafrost-dominated areas of Alaska, wetland soils tend to be frozen during snowmelt
events, resulting in a significant proportion of these floodwaters running directly to streams, thus
rendering these wetlands unimportant in streamflow regulation. Likewise, Rouletand Woo (1986)
found that wetlands in the Continuous Permafrost Region of Canada tended to be unimportant for either
long-term water storage or streamflow regulation.
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Regression equations developed to predict peak flows during flooding events generally use lake and
wetland storage areas as variables. Using this approach for Wisconsin watersheds, Novitzki (1979)
estimated that peak flood flows were only 20% as large in watersheds with 40% lake and wetland area
relative to watersheds without lakes or wetlands. Johnston et al. (1990) found that small losses of
wetlands in watersheds with <10% wetlands could have major effects on flood flow in basins around
Minneapolis, Minnesota. Wang et al. (2010) modeled the influence of wetlands on hydrologic processes
in Manitoba and Minnesota and found that the loss of 10-20% of the wetlands in the study basins would
increase peak discharge by 40%. Similarly, Yang et al. (2010) calculated restoration of 600 ha of
wetlands in a 25,139 ha watershed would decrease peak stream discharge by 23%. Peak streamflows
were shown to be negatively correlated with lake and wetland storage in Minnesota (Jacques and
Lorenz, 1988), although a later study found peak flows to be correlated with lake storage only and not
wetland storage (Lorenz etal., 2010).
The ability of wetlands to reduce flooding via storage varies with topography, wetland type, antecedent
moisture conditions, and available water storage capacity. Using stable hydrogen and oxygen isotopes of
water, McEachern etal. (2006) found that snowmelt in boreal forests was discharged rapidly in a sloped
watershed. In contrast, in a lowland watershed, much of the snowmelt was stored by wetlands,
particularly by bogs with stream channel outlets. In northern Canada, stream runoff was positively
correlated with slope and the presence of channel fens, but negatively correlated with lowland
depressional bogs (Quinton et al., 2003). In a Light Detection and Ranging (LiDAR)-based assessment of
depressional wetlands in Florida, Lane and D'Amico (2010) found an average potential wetland water
storage capacity of 1,619 m3 ha"1, with values ranging from 1,283 m3 ha-1 for palustrine scrub-shrub
wetlands to 2,906 m3 ha-1 for palustrine aquatic-bed wetlands. A literature review found that four out of
four studies that examined surface water depressions having no direct connectivity to a river system
concluded that those wetlands reduced or delayed flooding (Bullock and Acreman, 2003). Findings were
more varied for slope wetlands with direct connectivity to a river: 26 of 62 studies found reduced
flooding, while 27 of the 62 studies concluded that those wetlands increased flooding.
In addition to wetland type, antecedent moisture conditions and available storage capacity also
influence wetland water retention. The wetlands noted above, that serve as stream origins, likely
increased flood peaks under saturated conditions, with low additional wetland water storage capacity
(due to spring rains or snowmelt, for example), and thus conveyed any additional precipitation rapidly
downstream (Bullock and Acreman, 2003). Similarly, Branfireun and Roulet (1998) concluded that prior
saturation of upland areas immediately surrounding a wetland produced increased stoonflows. This
might mean that wetlands have less attenuating effect on larger floods because floods commonly occur
during saturated conditions.
Besides affecting peak flows and downstream flooding, non-floodplain wetlands can alter baseflow or
stormflows during dry periods. Ground-water discharge wetlands that are connected to streams, such as
fens or seeps, are important sources of baseflow (Morley et al., 2011). Moreover, wetlands can be focal
points for ground-water recharge and thus might contribute to baseflow. Rains (2011), for example,
found that perched and flow-through ponds in southwestern Alaska were sites of net ground-water
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recharge. Given the high prevalence of ponds on the landscape (Rains, 2011), these wetland types
cumulatively could substantially affect stream baseflow via ground-water inputs.
Other wetlands, however, might actually reduce flows during dry periods. Bullock and Acreman (2003)
concluded that this was the case in two-thirds of the studies they surveyed. Antecedent moisture
conditions and available wetland storage could partially explain this finding, in combination with
relatively high evaporation rates from wetland-dominated landscapes (Bullock and Acreman, 2003).
One study cited in their review (Boelter and Verry, 1977) noted that two storms of nearly equal volume
and intensity produced different runoff responses from the same peatland. One storm occurring in the
spring at a time of already high water tables led to runoff. The other, in midsummer at a time of low
water tables, increased the water depth in the peatland but did not exceed the wetland's water storage
capacity, precluding runoff. This mechanism has been observed in simulations of prairie pothole
hydrology, in which wetlands reduced streamflow until storage capacity was exceeded (Haan and
Johnson, 1968). Thus, wetlands can function as a sink in dry periods if storage capacity is not exceeded
and evaporation rates surpass ground-water recharge. Where storage capacity is exceeded during storm
events in otherwise dry periods, watersheds containing extensive wetlands can require more time for
water discharge to rise and fall in response to storm events (Lindsay et al., 2004). This finding suggests
that watersheds with wetlands take longer to fill and exceed water-holding capacity than watersheds
without wetlands and so, in this case, they provide a lag function by releasing water downstream more
slowly.
Non-floodplain wetlands also can reduce the variability of baseflow through landscape hydrologic
capacitance (McLaughlin etal., 2014). McLaughlin etal. (2014) simulated the effects of geographically
isolated wetlands on the variation in baseflow and found that the magnitude of this effect increased with
total wetland area. Holding area constant and increasing the number of wetlands (while decreasing their
size) also increased this capacitance. The effect of these wetlands on baseflow was the result of
differences in specific yield (the change in output or input depth from evaporation or rain per change in
water level) between wetlands and uplands, which causes flow reversals between them (McLaughlin
and Cohen, 2013; McLaughlin et al., 2014). Specifically, water flows from upland areas to wetlands
(wetland discharge) during wet periods and from wetlands to uplands (wetland recharge) during dry
periods, thereby buffering water tables and baseflow.
4.4.3 Effects of Non-floodplain Wetlands on Water Quality
Non-floodplain wetlands can affect water quality of rivers and other aquatic systems through processes
that can be generalized as source and sink functions, often mediated by transformational processes (see
Section 4.3.3 for details on specific mechanisms). In some cases, non-floodplain wetlands directly modify
the water quality in downstream waters through their relative lack of surface water connections; this
modification is accomplished by removal, sequestration, or transformation of pollutants such as
nitrogen, phosphorus, and metals through processes described by Ewel and Odum (1984), Mitsch et al.
(1995), Reddy and DeLaune (2008), and Kadlec and Wallace (2009), among others. Although non-
floodplain wetlands can lack surface water connections to downstream waters, surface and near-surface
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hydrologic connections to downstream waters do occur in many non-floodplain systems (Section 4.4.2;
Figure 2-18; Sun etal., 1995; Whigham and Jordan, 2003; Wilcox etal., 2011), providing pathways for
materials transformed in non-floodplain wetlands (such as methylmercury or degraded organic matter)
to reach and affect other aquatic systems.
Below we show that non-floodplain wetlands are areas where extensive microbially mediated processes
occur that can affect downstream waters. In Section 4.4.3.1, we describe how non-floodplain wetlands
are sources for dissolved organic matter and entrained elements like carbon, nitrogen, and phosphorus,
which are important components of food webs in downstream waters. Dissolved organic matter is also
shown to be important in regulating whole-lake acidity and buffering capacity. Mercury is another
material affected by microbial processing in non-floodplain wetlands; mercury can be transported along
with dissolved organic matter to downstream waters, where it can become incorporated into the food
web with potentially deleterious effects. In Section 4.4.3.2, we discuss how non-floodplain wetlands
serve as sinks by sequestering or transforming materials, thereby affecting the chemical, physical, or
biological condition of downstream waters. Nitrogen, nitrate, ammonium, and phosphorus compounds
are shown to be removed or assimilated—often at high rates—in non-floodplain wetlands. Pesticides,
metals, and other potential pollutants also can be sequestered or assimilated in non-floodplain wetlands.
4.4.3.1 Non-floodplain Wetlands as Sources for Downstream Waters
Like all wetlands, non-floodplain wetlands contain diverse microbial populations that have adapted to
hydrologic, physical, and chemical extremes (Reddy and DeLaune, 2008). Microbial populations abound
in wetland systems; for example, Boon (1991) reported that Australian wetlands contained 100 times
more microbes in the water column than nearby rivers, with up to 157 x 109 cells Ir1. Functions that
occur in non-floodplain wetlands can affect streams, rivers, and lakes when compounds that are
transformed in wetland environments move to downstream waters through overland flow or shallow
ground water (Section 4.4.2; Winter et al., 2003). Two processes that occur in non-floodplain wetlands
(and in riparian/floodplain wetlands) are useful to illustrate the influence of non-floodplain wetlands on
downstream waters: the methylation and transport of the bioaccumulating pollutant mercury, and the
breakdown and transport of organic compounds to receiving waters.
Freshwater wetlands/peatlands are areas of active methylmercury (MeHg) production (Grigal, 2002).
Ullrich et al. (2001) noted that methylmercury production was linked to low pH, low salinity, and
presence of decomposable organic matter in reducing environments. Sulfate-reducing bacteria are
primarily responsible for biological mercury methylation and thrive in the reduced conditions at
wetland aerobic/anaerobic boundaries (Benoit et al., 1999); the addition of sulfate (e.g., through
atmospheric acid deposition) increases the formation of methylmercury in peatlands (Branfireun et al.,
1999). Once formed through microbial (or other) processes, mercury and methylmercury export is
controlled by the export of organic matter, such as dissolved organic compounds and humic and fulvic
acids (Linqvist et al., 1991; Mierle and Ingram, 1991; Driscoll etal., 1995). Methylmercury can be
translocated in watersheds having non-floodplain wetlands by entrainment with organic matter exports.
It also can move through near-surface and surface flows from non-floodplain peatlands to downstream
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waters. For example, Branfireun et al. (1996) reported 58% of MeHg-laden peat porewater leaving a
headwater catchment study area occurred during stormflow, 41% during baseflow, and 1% transported
via ground water. St. Louis et al. (1994) found that boreal forest catchments in Minnesota with non-
floodplain wetlands reduced total mercury concentrations, but had yields of methylmercury from
wetlands that were 26-79 times higher than upland areas. This yielded 1.84-5.55 mg MeHg ha-1 yr-1 to
streams in the Great Lakes basin, where mercury could be incorporated into lake-wide food webs.
Hurley et al. (1995) contrasted MeHg yields from different land use groups in Wisconsin and found that
wetland/forest sites were higher than agricultural/forested and agricultural-only sites. Similarly,
Porvari and Verta (2003) found that bioaccumulating methylmercury export from non-floodplain
peatlands to downstream waters ranged from 0.03 to 3.8 ng MeHg Ir1, and that catchments with greater
wetland abundances had greater methylmercury export.
Export of dissolved organic matter can have negative effects on downstream waters because
contaminants, such as methylmercury and other trace metals, can be adsorbed to it (Thurman, 1985;
Driscoll et al., 1995). Dissolved organic matter, however, is also an important source of energy for
downstream aquatic communities (Hobbie and Wetzel, 1992; Reddy and DeLaune, 2008). Wetlands are
the principal source of dissolved organic compounds to downstream waters in forested ecosystems
(Mulholland and Kuenzler, 1979; Urban et al., 1989; Eckhardt and Moore, 1990; Koprivnjak and Moore,
1992; Kortelainen, 1993; Clair etal., 1994; Hope etal., 1994; Dillon and Molot, 1997; Gergel etal., 1999).
Over prolonged periods, reductions in dissolved organic carbon (DOC) export (e.g., through wetland
conversion or degradation or alterations in hydrology) decrease the ability of downstream waters to
support primary productivity, due to reduced export of entrained carbon, nitrogen, sulfur, and
phosphorus (Hedin etal., 1995; Nuff and Asner, 2001). Changes in DOC export also affect the pH and
buffering capacity of downstream aquatic systems (Eshelman and Hemond, 1985) and their exposure to
damaging UV-B rays (Schindler and Curtis, 1997). Boreal forest basins composed of non-floodplain
wetlands in central Ontario were found to export between 11.4 and 31.5 kg C ha-1 yr-1 to downstream
waters (Creed etal., 2003). Furthermore, near-surface lateral transport of DOC explained 88% of the
variation in basin DOC export to lake systems where it directly affected pH and buffering capacity. Other
studies have similarly shown a relationship between the proportion of wetlands in a watershed and the
average annual concentration of DOC in the receiving streams of that area, and other areas of the boreal
forest/Precambrian Shield (Urban et al., 1989; Eckhardt and Moore, 1990; Koprivnjak and Moore, 1992;
Detenbeck et al., 1993; Clair et al., 1994; Hope et al., 1994; Dillon and Molot, 1997; Johnston et al., 2008).
The export of dissolved organic compounds from non-floodplain wetlands also can affect the acidity of
downstream waters. Gorham et al. (1986) addressed watershed factors associated with lake and forest
acidification in Nova Scotia, Canada. In addition to atmospheric deposition of acid precipitates, they
found that the ratio of non-floodplain muskeg peatlands to lakes was significantly correlated with lake
acidification, as muskeg wetland-dominated watersheds exported high-molecular-weight organic acids
via either overland or shallow ground-water flow. Further linking non-floodplain wetlands to lakes,
Gorham et al. (1986) reported that even small amounts of humic DOC can greatly affect lake water pH;
the pH of waters with a dissolved organic carbon value of 4.5 mg DOC L-1 (the log-normal mean) was
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100 times more acidic than waters with a dissolved organic carbon of <1 mg DOC Ir1 (the minimum
concentration).
4.4.3.2 Non-floodplain Wetlands as Sinks and Transformers for Downstream Waters
The wetland literature is replete with examples of wetlands improving water quality through
assimilation, transformation, or sequestration of nutrients and other pollutants (e.g., Ewel and Odum,
1984; Nixon and Lee, 1986; Johnston, 1991; Detenbeck et al., 1993; Mitsch and Gosselink, 2007; Reddy
and DeLaune, 2008; Kadlec and Wallace, 2009). These functions act on the large pool of pollutants that
are available through nonpoint sources. Non-floodplain wetland processes that affect pollutant
attenuation include denitrification, ammonia volatilization, and microbial and plant biomass
assimilation (Reddy and DeLaune, 2008). Other pollutants in wetland systems can be retained through
sedimentation, sorption and precipitation reactions, biological uptake, and long-term storage in plant
detritus (Reddy et al., 1999; Reddy and DeLaune, 2008).
Non-floodplain wetlands act as sinks and transformers for various pollutants. For example, high levels of
human sewage were applied to a forested non-floodplain wetland site for 4.5 years (Ewel and Odum,
1984 and chapters therein). More than 95% of the phosphorus (P), nitrate, ammonium, and total
nitrogen (N) were removed by the wetland during the study period (Dierberg and Brezonik, 1984), and
66-86% of the nitrate removed was attributed to the process of denitrification. In another example,
phosphorus retention in non-floodplain marshes of the lower Lake Okeechobee basin ranged from 0.3 to
8.0 mg soluble reactive P nr2 d-1 (Dunne etal., 2006). This retention represents a sizeable amount of
phosphorus removal, because only about 7% of the watershed comprised non-floodplain marsh.
Similarly, wetlands in the Lake Okeechobee, Florida basin were found to have greater storage of total
phosphorus than the uplands in which they were bedded, 236 kg ha-1 vs. 114 kg ha-1 (Cheesman et al.,
2010). These findings were echoed by Dunne etal. (2007), who reported that more phosphorus was
stored in wetland plant biomass and soil than in corresponding upland compartments, with wetland
surface soils (0-10 cm) representing the largest phosphorus reservoir (>87%) and soil organic matter
accounting for >69% of the soil total phosphorus variability. They further suggest that restoring 5-20%
of the geographical isolated wetland area in priority basins draining to Lake Okeechobee, Florida, could
increase phosphorus storage in geographical isolated wetlands by up to 13 kg P ha-1, mostly through
increased soil organic matter with its concomitant phosphorus in wetland soils (Dunne et al., 2007).
Marton et al. (2014) found that mean phosphorus sorption was approximately two to three times
greater in natural depressional wetlands than in restored wetlands and agricultural fields (297,114, and
86 mg P kg soil-1, respectively). Marton et al. (2014) also found that depressional wetlands sorbed twice
as much phosphorus as riparian systems. Craft and Casey (2000) reported similar accretion rates in
depression and floodplain wetlands of Georgia for sediment, organic carbon, and nitrogen, and
significantly highly floodplain storage of phosphorus. Cohen et al. (2007) found that riparian wetlands
had higher phosphorus-sorption capacities than non-riverine wetlands. Non-floodplain wetland flats
studied in Maryland and Delaware had microbially mediated denitrification enzyme activity (an
indicator of potential denitrification) rates of 0.06-0.76 mgN kg-1 d-1 (Jordan etal., 2007). Because flats
comprise greater than 70% of the wetland area in the basin, this value indicates a significant
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denitrification capacity. Marton et al. (2014) found that depressional wetlands denitrified at twice the
rate upland systems did, 12.3 ± 4.5 ng N g-1 hr1 versus 5.3 ± 1.7 ng N g-1 hr1. Craft and Chiang (2002)
determined that wetland soils stored a disproportionately large share of nitrogen, compared with
upland soils, in spite of uniform soil organic matter across the landscape. A non-floodplain bog in
Massachusetts was reported to sequester nearly 80% of the system's various nitrogen inputs, including
precipitation that had a range of 1.2-1.9 mg N Ir1 (Hemond, 1983). Prairie pothole wetlands in the
upper Midwest removed >80% of the nitrate load via denitrification (Moraghan, 1993). A large non-
floodplain prairie marsh removed 86% of nitrate, 78% of ammonium, and 20% of phosphate through
assimilation and sedimentation, sorption, and other mechanisms (Davis et al., 1981). Geographically
isolated, non-floodplain wetland systems in Michigan were found to remove nitrate-nitrogen (NOs-N)
and sulfate (S042-) at rates of 0.04-0.55 mgN03-N Ir1 ha-1 and 0.06-0.30 mgS042- L-1 ha-1. These rates
are significant, considering that nitrate-nitrogen pollution of ground water in Michigan was reported to
average 0.50 mgN03-N L-1 (Whitmire and Hamilton, 2008). Bhadha etal. (2011) found that infiltration
to the ground accounted for 14% of phosphorus loss from two historically isolated wetlands in a Florida
study area, suggesting that near-surface flow gradients are important to landscape-level phosphorus
dynamics. Together, these studies indicate that sink removal of nutrients by non-floodplain wetlands is
significant and geographically widespread.
Other pollutants and compounds can be mitigated by non-floodplain wetland sink and transformation
processes. For example, microbial methanogenesis completely removed the pesticide atrazine from a
mountainous bog in North Carolina (Kao et al., 2002). The environmental contaminants cobalt (Co) and
nickel (Ni) can be phytoremediated by wetland plants common in forested non-floodplain wetlands of
the Southeast; plant concentrations were found to range from 1 to 530 mg Co kg-1 and up to 250 mg
Ni kg-1 (Brooks et al., 1977). A bog in Massachusetts that Hemond (1980) extensively studied acted as a
sink and annually stored 54 mg magnesium nr2, 3 6 mg potassium nr2, and 46 mg lead nr2; the bog also
provided acid-rain buffering for downstream waters. Based on the literature, Boon (2006) concluded
that wetland microbial communities can mediate processes that degrade diesel fuel and other
hydrocarbons, pesticides, heavy metals and metalloids, and chlorinated solvents that can pollute ground
water.
4.4.4 Biological Connections Between Non-floodplain Wetlands and
Streams
Many of the same factors that affect movement of organisms between riparian/floodplain wetlands and
the river network (Section 4.3.4) govern movement of organisms between non-floodplain wetlands and
the river network. Non-floodplain wetlands, however, are generally farther from stream channels than
riparian/floodplain wetlands, which reduces hydrologic connectivity. The distance, number, and variety
of heterogeneous landscape patches (including barriers) over which organisms must disperse also can
be greater. Organisms have evolved numerous complex dispersal strategies to overcome non-floodplain
flows, reduced hydrologic connectivity, and increased geographic distance between habitats and
spatially subdivided populations. Passive transport (e.g., wind dispersal, "hitchhiking" on other animals)
and active movement (e.g., walking, crawling, flying) are common modes of dispersal that can establish
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connectivity in the absence of hydrologic flows. Such dispersal events are often sporadic and
asymmetric in non-floodplain wetland landscapes, making them more difficult to observe than surface
water flows. Their effects on community structure and diversity—including metapopulation effects of
wetland-to-wetland connectivity—have been well documented (e.g., Wellborn et al., 1996; Snodgrass et
al., 2000; Julian et al., 2013), especially for amphibians. Other effects, such as water quality and
population or species persistence, are not well understood. Below we review the various dispersal
mechanisms that operate in non-floodplain wetland landscapes.
Despite being sessile, plants have evolved many adaptations that facilitate dispersal. Considerable
attention has been given to waterborne dispersal of aquatic and emergent macrophytes (Nilsson et al.,
2010), which can play a role in non-floodplain wetlands that are periodically connected hydrologically
to river networks. In addition, significant numbers of such plants can be dispersed as seeds or pollen by
wind (Soons, 2006). Wind dispersal enables colonization of geographically isolated non-floodplain
wetlands such as prairie potholes (Galatowitsch and van der Valk, 1996). Given that geographically
isolated wetlands are surrounded by uplands, using wind as a vector carries the relatively high risk that
propagules of obligate wetland plants will land in unsuitable habitat. Plants have developed colonization
strategies to compensate for such risks. For example, Soons and Heil (2002) showed that producing
large numbers of seeds increased colonization success of short- and long-distance dispersing grassland
forbs; results from this and other studies are being applied to models of wetland dispersal and
colonization (e.g., Soons, 2006). Viable seeds or vegetative plant parts also can travel great distances
within the guts of or externally attached to migratory birds (Murkin and Caldwell, 2000; Amezaga et al.,
2002; Figuerola and Green, 2002), which move between non-floodplain wetlands and river networks,
depending on temporally dynamic habitat availability (Murkin and Caldwell, 2000; Haukos et al., 2006
and references therein).
Identifying specific source and recipient populations for any organism over these distances can be
challenging, but especially for plants having passively mobile life stages that cannot be precisely tracked.
Determining whether wetlands function as sources to or recipients of plant propagules from river
networks is especially difficult. Genetic similarity between populations can provide general evidence of
connectivity between non-floodplain wetlands and the river network. Sawgrass (Cladium jamaicense)
populations in Everglades wetlands showed low population genetic divergence at distances greater than
100 km; wind pollination and water dispersal of propagules through flooding likely keeps channel and
wetland populations genetically similar (Ivey and Richards, 2001). Another approach that can provide
evidence for dispersal is community-level surveying, which takes into account local determinants of
community composition and structure. Controlling for local conditions like rainfall and soil type, a study
in Connecticut (Capers et al., 2010) found that bodies of water—from small isolated wetlands to large
lakes—that were located closer together had more similar plant communities. This finding suggests
biological connectivity between proximal lakes and wetlands.
Recent evidence suggests that invertebrate hitchhiking on birds and mammals is more common than
previously thought (Figuerola and Green, 2002; Figuerola etal., 2005). Allen (2007) trapped
zooplankton dispersing from a pond in Illinois and found that animals wider than 3 cm were the primary
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vector of reproductive adult zooplankton forms. These results suggest that animals moving among
water bodies can be an important factor in structuring non-floodplain wetland invertebrate
metapopulations. Frisch et al. (2007) found that diapausing invertebrate eggs that dispersed by
hitchhiking on birds had higher incidences of hatchingin January (59.4%) than in November (11.5%).
These invertebrates included nematodes, zooplankton (i.e., rotifers, ostracods, copepods), and insects
(i.e., crane flies, nonbiting midges, hemipterans). This study indicates that winter migrations of aquatic
birds can be an important mechanism for spring colonization of habitats separated by hundreds or even
thousands of kilometers. Studies have thus shown that migratory birds can passively connect viable
plant matter, macroinvertebrates, and zooplankton from disparate habitats across the landscape, with
likely—although unresolved—impacts on food web dynamics (Polls etal., 1997).
The scientific literature has many examples of migratory birds—especially migratory waterfowl,
including cranes, geese, ducks, and shorebirds—actively moving between and using the different
available resources of estuarine, riverine, and riparian systems and non-floodplain wetlands. For
example, wood ducks (Aixsponsd) are found throughout freshwater deciduous forests of North America.
Preferred breeding sites include river floodplains, remote ponds, and woodland pools that receive
snowmelt and spring rain, the latter particularly indicative of non-floodplain wetland use (Haramis,
1990). Below we provide several examples of this type of biological connectivity that can connect non-
floodplain wetlands to each other and to other aquatic systems.
Approximately 80% of the entire North American population of redhead ducks (Aythya americand)
winters along coastal Texas and northern Mexico (Weller, 1964). Woodin (1994) identified more than
20,000 redheads using both estuarine systems and freshwater wetlands, reporting that the estuarine
systems were exclusively used for feeding, while freshwater coastal pond wetlands were used almost
exclusively for drinking water and courting (Mitchell et al., 1992). The coastal ponds redheads used
were seasonal basins, which frequently dried completely (Ballard etal., 2010). Ballard etal. (2010)
further noted that although the ponds were densely distributed in coastal Texas (up to 4.8 coastal basins
per km2), water availability varied year-to-year. As a result, during dry years redheads would use
available coastal ponds up to 8.1 km from the estuarine forging areas, while in wetter years closer ponds
would be used (likely to minimize energy expended through flying). Similarly, Adair et al. (1996)
reported that lesser scaup [Aythya affinis) and redheads avoided salt stress and metabolically expensive
osmoregulation through salt-gland excretory functions by feeding in estuaries; drinking, preening, and
resting in coastal basins; and then returning to estuaries. Grey teals (Anasgibberifronsgracilis) in
Australia that feed in saline areas similarly required freshwater to osmoregulate (Lavery, 1972). Mallard
ducks (Anas platyrhynchos) transiting Iowa during spring migration used seasonally flooded farmed
basins in agricultural fields (also known as sheetwater wetlands; LaGrange and Dinsmore, 1989) for
feeding and roosted in more permanent emergent wetlands at night. In the study, these shallow
sheetwater wetlands provided 19,530 mallard use-days during the daytime compared with 103 use-
days for the emergent wetlands.
Nebraska's Rainwater Basin historically had more than 11,000 playas, shallow wind-formed wetland
depressions, although human activities over the past 100 years have resulted in the loss of 90% of the
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number and approximately 88% of the area (Webb et al., 2010; Uden et al., 2014). Nevertheless, the
remaining basins are critical to dependent migratory waterfowl, with 7-10 million waterfowl using the
approximately 16,000 km2 area, including "virtually all of the 600,000 midcontinental greater white-
fronted geese (Anser albifrons), 500,000 Canada geese (Branta Canadensis), 50% of midcontinent
mallards (Anas platyrhynchos], and 30% of continental northern pintails (Anas acutd)" (Webb et al.,
2010, p. 109), 38 shorebird species, and the endangered whooping crane (Grus americand). In a 3-year
spring migration study of 36-40 playas, Webb et al. (2010) identified 72 migratory species and more
than 1.6 million birds actively using these playa basins. The abundance of all wetland bird taxa was
related to wetland area within 5-10 km of the study playas, although diving duck abundance (e.g.,
redhead, canvasback, lesser scaup) was specifically related to riparian area within 5 km, likely due to the
presence of open water within these systems (Webb et al., 2010; see their Table 1 for a complete list of
taxa found).
Many additional studies have identified Nebraska as an important staging and stopover area for
numerous species, perhaps due to its location on the Central Flyway. For example, almost the entire
population of midcontinent sandhill cranes (Grus canadensis] uses the Central Platte River Valley. Avian
researchers reported that cranes roost along both the current and former Platte River channel (Krapu et
al., 1984) and forage in grasslands on semipermanent (unconsolidated mud bottom) and temporary
palustrine wetlands (Folk and Tacha, 1990) and on frequently inundated soils—especially those within
4.8 km of roost sites (Anteau etal., 2011). Pearse etal. (2010) noted that after feeding in cornfields,
sandhill cranes roosted along the Central Platte River Valley in pastures with ponds. These pond
systems are likely either playas, as noted above, or palustrine wetlands often surrounded by croplands
(Austin and Richert, 2005). Austin and Richert (2005) further stated that the endangered whooping
crane was noted as roosting, feeding, and resting in both riverine and palustrine wetlands of the Great
Plains. Vrtiska and S.Sullivan (2009) found that lesser snow geese (Chen caerulescens] and Ross's geese
(Chen rossii), which numbered up to 7.3 million in 2001 during peak migration, used wetland habitats in
both the Rainwater Basin and Central Platte River Valley, depending on the availability of suitable (e.g.,
inundated) habitat.
Blanchong et al. (2006) found that this concentrated use of the Rainwater Basin by migratory lesser
snow geese resulted in greater contact between individuals, contributing to the spread ofPastruella
multocida, the bacterium that causes avian cholera. The loss of wetlands within the basin has resulted in
higher concentrations of migratory birds within the remaining wetlands, which has led to higher risks of
outbreaks of infectious diseases (Blanchong et al., 2006).
The U.S. Fish and Wildlife Service's Subcommittee on Rocky Mountain Greater Sandhill Cranes (SRMGSC,
2007) reviewed the literature on habitat use for the migratory population of Rocky Mountain sandhill
cranes. This population, one of five in North America, migrates from wintering areas in Arizona, New
Mexico, and central Mexico to breeding areas in Canada, Montana, Idaho, Wyoming, Utah, and Colorado.
SRMGSC (2007) reported that this population of sandhill cranes overwintered in multiple riverine,
riparian, and non-floodplain habitats, including playas in New Mexico and southeastern Arizona. Areas
used in the breeding range include non-floodplain wetlands, such as northern boreal forest bogs, and
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other habitat types (e.g., large marsh complexes, smaller, scattered marshes, intermittent streams,
beaver ponds, subirrigated wet meadows along riparian zones; SRMGSC, 2007).
Shorebirds also use multiple habitat types during their North American migration. Skagen and Knopf
(1993) concluded that dispersion and opportunism, rather than concentration and predictability,
characterize movements of shorebirds in the Great Plains. For example, Haig et al. (1998) noted that
large population declines of the endangered migratory piping plover (Charadrius melodus) along the
Missouri River were not actually declines, but a result of the birds moving to the Missouri Coteau (a 7.3
million ha region of the Upper Midwest and Canada replete with closed-basin prairie potholes; Phillips
et al., 2005), due to increased flooding along the Missouri. Farmer and Parent (1997) monitored pectoral
sandpipers [Calidris melanotus) migrating through non-floodplain sheetwater wetlands in Missouri and
small depressional wetlands of the Rainwater Basin in Nebraska and found that habitat connectivity
affected shorebird movements. Habitat patch density affected movements such that pectoral sandpipers
often perceived groups of wetlands as functionally connected and actively exploited the best feeding
habitat within that wetland complex. As the landscape became disconnected, however, the monitored
species altered their movement behavior, minimizing energy expenditure (Farmer and Parent, 1997).
Other taxa have been reported as linking downstream systems and non-floodplain wetlands. Fish tend
to disperse between non-floodplain wetlands and the river network during periodic surficial hydrologic
connections or when humans create surface-water connections via ditching (Snodgrass et al., 1996;
Langston and Kent, 1997; Zimmer et al., 2001; Baber et al., 2002; Hanson et al., 2005; Herwig et al.,
2010). Mammals that can disperse overland can also contribute to connectivity. Although muskrat
territories are usually restricted (Shanks and Arthur, 1952), dispersal between suitable river and non-
floodplain wetland habitat over longer distances that is seasonal, climate-induced, and density-
dependent has been observed (Serfass etal., 1999; Clark, 2000 and references therein). Spinola etal.
(2008) tracked translocated river otters (Lontra canadensis) in New York and found that, after release,
most otters inhabited a mosaic of isolated aquatic habitats distributed throughout the agriculture-
dominated landscape. As noted above for waterfowl, mammals (including muskrats) also can act as
transport vectors for hitchhiking organisms like algae (Roscher, 1967).
Numerous flight-capable insects, including mayflies, caddisflies, diving beetles, backswimmers, whirligig
beetles, water striders, water boatmen, scavenger beetles, crane flies, and nonbiting midges, use both
streams and non-floodplain wetlands (Williams, 1996). Aerial dispersal enables such insects to move
outside the stream network to seek suitable habitat for overwintering, refuge from adverse conditions,
hunting, foraging, or breeding (Williams, 1996; Bohonak and Jenkins, 2003).
Amphibians and reptiles also move between streams or rivers and non-floodplain wetlands to satisfy
part of their life-history requirements (Table 4-2). For example, Subalusky et al. (2009a) and Subalusky
et al. (2 009b) reported movement of adult female alligators (Alligator mississippiensis) from creeks to
shallow, seasonal limesink wetlands for nesting and use of the wetlands as nurseries for juveniles.
Subadults then shift to habitats within the river network by moving overland to the creek (Subalusky et
al., 2009a; Subalusky et al., 2009b). Lamoureuxand Madison (1999) used radio tracking to follow
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movements of green frogs (Rana clamitans) for 9 months in New York. Green frogs, which breed in
wetlands and then move into terrestrial habitats, are susceptible to freezing temperatures. In late
autumn, the frogs moved from upland habitats near breeding ponds to rapidly flowing streams and
seeps to overwinter. Boreal toads (Bufo boreas boreas} disperse long distances (>1 km) in streams
through home ranges (Adams etal., 2005). Knutson etal. (1999) found that the strongest land-use
predictor of anuran richness was urban land use. They speculated that, in addition to urban landscapes
being detrimental to anuran habitat quality, their tendency to fragment (i.e., disconnect) anuran habitats
is also a factor in the decline of these assemblages. In northwestern Ohio and southern Michigan
wetland complexes, the abundance of northern watersnakes [Nerodia sipedon sipedori) was positively
correlated with wetland size and wetland connectivity, defined by the authors as a wetland's distance to
other wetlands (Attum et al., 2007). The American toad [Anaxyrus [=Bufo] americanus) and eastern newt
[Notophthalmus viridescens) are widespread habitat generalists that move among streams and wetlands
to take advantage of both habitats, feed on aquatic invertebrate prey, and avoid predators (Table 4-2;
Babbitt et al., 2003; Green, 2005; Hunsinger and Lannoo, 2005; Petranka and Holbrook, 2006).
4.4.5 Geographic Isolation of Non-floodplain Wetlands
In defining non-floodplain wetlands (Section 2.2.1), we noted that this category could include wetlands
that are geographically isolated and those that are not. Further, we noted (Section 2.4.1) that certain
types of wetlands can be found with or without an outlet and can occur along a gradient of hydrologic
connectivity. This gradient can include non-floodplain wetlands that have permanent hydrologic
connections to the river network through perennial channels; wetlands that have losing streams that are
completely disconnected from the river network as output channels; geographically isolated wetlands
that have ground-water or occasional surface-water connections; and geographically isolated wetlands
that have minimal hydrologic connection to the river network (but which could include surface and
subsurface connections to other wetlands). The existence of this gradient (Section 1.2.2) can make
determining the degree to which particular non-floodplain wetlands are connected to or isolated from
downstream waters difficult.
A related issue is that spatial scale must be considered when determining geographic isolation. Tiner
(2003c) provided examples of how a wetland that was not isolated at a local scale could be
geographically isolated at a larger scale. Conversely, individual wetlands that are geographically isolated
could be connected to downstream waters when considered as a complex (a group of interacting
wetlands). This concept is demonstrated by Wilcox etal. (2011), who examined a depressional wetland
complex on the Texas Coastal Plain. Although the wetlands are hydrologically connected to each other
by shallow swales, they might be geographically isolated, because swales often are considered upland. In
fact, Tiner (2003c) classifies these Coastal Plain wetlands as geographically isolated. At the scale of the
wetland complex, however, the wetlands are connected to a nearby waterway via an intermittent
stream. During an almost 4-year study, nearly 20% of the precipitation that fell on the wetland complex
flowed as surface runoff through the channel to a nearby waterway, the Armand Bayou (Wilcox et al.,
2011). Although these wetlands might be geographically isolated at the local scale, the wetland
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^H Table 4-2. Partial list of amphibian and reptile species known to use both streams and non-floodplain wetlands or other lentic waters.
Stream and Wetland Connect
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1
f-
w
O
Common Name
Green frog
Leopa rd frog
Bullfrog
Columbia spotted frog
Southern leopard frog
Pacific chorus frog
American toad
Fowler's toad
Two-toed amphiuma
Greater siren
Eastern newt
Scientific Name
Rana clamitans
Rana pipiens
Rana catesbeiana
Rana luteventris
Rana sphenocephala
Pseudacris regilla
Anaxyrus [=Bufo]
americanus
Anaxyrus [=Bufo] fowleri
Amphiuma means
Siren lacertina
Notophthalmus viridescens
Habitat Use
Breeds in wetlands and pools; overwinters in streams (Lamoureux and Madison, 1999)
Breeds in wetlands and pools; overwinters in streams (Rorabaugh, 2005)
Uses seasonal pools as complementary nonbreeding habitat (Gahl et al., 2009)
Breeds in streams and wetlands; overwinters in streams (Pilliod etal., 2002)
Breeds in shallow pools and wetlands; adults inhabit many shallow freshwater habitats, including
temporary pools, cypress ponds, ponds, lakes, ditches, streams, river edges, floodplain pools, and
slightly brackish coastal wetlands (Butterfield, 2005)
Breeds in wetlands, ponds, temporary pools, streams, lakes, rivers, and other aquatic habitats
(Rorabaugh and Lannoo, 2005)
Breeds in lakes, ponds, streams, ephemeral wetlands, prairie potholes, ditches, and floodplain
pools (Green, 2005)
Breeds in ponds, temporary pools, streams, ditches, lake shores, and shallows of rivers (Green,
2005)
Adults inhabit a wide variety of aquatic environments, including ponds, lakes, ephemeral
wetlands, wet prairies, streams, and ditches (Gibbons and Semlitsch, 1991; Johnson and Owen,
2005)
Breeds in shallow pools and streams, adults live in lakes, streams, ponds, and wetlands (Gibbons
and Semlitsch, 1991; Hendricks, 2005)
Breeds in permanent and semipermanent pools, ponds, wetlands, and low-flow areas of streams;
adults live in pools, ponds, streams, and wetlands (Hunsinger and Lannoo, 2005; Timm et al.,
2007)
to
o
-------
£=?
sam and Wetland Connectivity:
eview and Synthesis
(continued).
Common Name
Yellow-bellied
watersnake
Copper-bellied
watersnake
Spotted turtle
Blanding's turtle
Painted turtle
Snapping turtle
American alligator
Scientific Name
Nerodia erythrogaster
flavigaster
Nerodia erythrogaster
neglecta
Clemmys guttata
Emydoidea blandingii
Chrysemys picta
Chelydra serpentina
Alligator mississippiensis
^^^^^^1
Habitat Use
Hunts in temporary pools and wetlands (Roe etal., 2004; Mitchell etal., 2007)
Hunts in temporary pools and wetlands (Roe etal., 2004; Mitchell etal., 2007)
Uses temporary wetlands for foraging, mating, basking, and aestivating (Joyal et al.,
Uses temporary wetlands for foraging, mating, basking, and aestivating (Joyal et al.,
2001)
2001)
Uses temporary wetlands for basking and foraging (Mitchell etal., 2007)
Uses temporary wetlands for basking and foraging (Mitchell etal., 2007)
Juveniles use seasonal wetlands as nurseries, subadults move back to river networks (Subalusky
etal., 2009a; Subalusky etal., 2009b)
to
o
-------
complex serves as the source of water for a headwater stream, and therefore, the complex is not
geographically isolated at a larger scale.
Besides the spatial scale of the wetland unit, assessments of non-floodplain wetland to stream
connectivity can be affected by the resolution and source of the spatial data that are used. For example,
higher connectivity was found in the Tuckahoe Creek watershed in Maryland, when wetland
connectivity was evaluated for streams determined from LiDAR compared to streams from both the
High Resolution National Hydrography Dataset (NHD) and NHD Plus (Lang et al., 2012). Yang and Chu
(2013) found that Digital Elevation Model (DEM) resolution also affected connectivity assessments, with
finer OEMs having a higher number of connected areas and less total connected area than coarser DEMs.
Given this discussion, caution should be used in interpreting connectivity for wetlands that have been
designated as "geographically isolated," because (1) the term can be broadly applied to a heterogeneous
group of wetlands that can include wetlands that are not actually geographically isolated, (2) wetlands
with permanent channels could be miscategorized as geographically isolated if the designation is based
on maps or imagery with inadequate spatial resolution (e.g., Lang et al., 2012), obscured views, etc., and
(3) wetland complexes could have connections to downstream waters through stream channels even if
individual wetlands within the complex are geographically isolated. The term "geographically isolated"
should be applied only to groups of wetlands if all those wetlands are, in fact, known to be
geographically isolated. Further, even geographically isolated wetlands can be connected to other
wetlands and downstream waters through ground-water connections, occasional spillage, or biological
connections. Thus, the term "geographically isolated" should not be used to infer lack of hydrologic,
chemical, or biological connectivity.
Finally, precisely this isolation is responsible for many of the functions that geographically isolated
wetlands provide to downstream waters. In particular, many of the sink and lag functions of these
wetlands result from their relative isolation from the river network. This relative isolation, combined
with the wetlands' storage capacity, enables them to store water and reduce peak streamflows and
downstream flooding (Novitzki, 1979; Hubbard and Linder, 1986; Vining, 2002; Bullock and Acreman,
2003; McEachern etal., 2006; Gleason etal., 2007). For example, depressional wetlands in Florida had
an average potential wetland water storage capacity of 1,619 m3 ha-1 (Lane and D'Amico, 2010). These
same sink and lag functions will also act on any materials associated with stored water, such as
sediments and pollutants. Increased isolation also can decrease the spread of pathogens (e.g., Hess,
1996) and invasive species (e.g., Bodamer and Bossenbroek, 2008) and increase the rate of local
adaptation (e.g., Fraser et al., 2011).
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4.5 Wetlands: Synthesis and Implications
4.5.1 Riparian/Floodplain Wetlands
Based on our review of the literature, riparian/floodplain wetlands are highly connected to streams and
rivers through surface water, shallow ground water, and biological connectivity. The effects of wetlands
on streams and rivers are a function of the magnitude of floodwaters, the geomorphic structure of the
floodplain, and the proximity of the channel. Although a gradient occurs in the frequency of connectivity
within the floodplain (Section 1.2.2), even riparian/floodplain wetlands that rarely flood can be
important because of long-lasting effects on streams and rivers. In fact, most of the major changes in
sediment load and river-channel structure—for example, movement of rivers through meander belts
and creation of oxbow lakes—that are critical to maintaining the health of the river result from large
floods that provide infrequent connections with more distant riparian/floodplain wetlands. Areas that
surface water infrequently floods also can be connected to the river more regularly through ground
water and the organisms. Key conclusions from our literature review on riparian/floodplain wetlands
are summarized in Table 4-3.
4.5.2 Non-floodplain Wetlands
Non-floodplain wetlands consist of depressional, slope, and flats wetlands that lack surface water inlets.
Non-floodplain wetlands can include regional wetland types such as prairie potholes, playa lakes, vernal
pools, and Carolina bays. Hydrologic flows through these wetlands are predominantly unidirectional, in
contrast to bidirectional flows that occur in riparian/floodplain wetlands.
The literature we examined on non-floodplain wetlands indicates that these systems have important
hydrologic, water-quality, and habitat functions that affect downstream waters and rivers provided a
connection exists between the wetland and downstream water (Table 4-4). The challenge is to identify
which non-floodplain wetlands have such a connection. Addressing this issue is difficult, because most
wetland studies do not investigate wetland effects on downstream waters or, if they do, they rarely
address connectivity explicitly.
Based on what is known about how water flows across the landscape (Chapter 2), hydrologists and
ecologists would generally agree that all non-floodplain wetlands are interconnected to some degree
and are connected with stream networks, which is why the water-cycle environment is referred to as
the hydrosphere. Hydrologists and ecologists also generally agree that some areas are more connected
or have a greater influence than others. The purpose of this review is to determine, based on the peer-
reviewed literature, the degree of connectivity and associated effects between different non-floodplain
wetlands and downstream waters.
Non-floodplain wetlands occur along the gradient discussed in Chapter 1, and can be described in terms
of the frequency, duration, magnitude, timing, and rate of change of water, material, and biotic fluxes to
downstream waters. With respect to hydrologic connectivity, this gradient includes wetlands that have
permanent hydrologic connections to the river network through perennial channels; wetlands that have
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Table 4-3. Key conclusions on the effects of riparian/floodplain wetlands on rivers.
Physical Connectivity and Function
Riparian areas are highly connected to streams, so much so that considering the riparian influence on
streams is essential to understanding their structure and function.
Riparian connectivity ranges from longitudinal flow and exchange in mountainous headwater streams to
increasing lateral flow and exchange in river valleys and coastal terrain.
Water storage by riparian areas, especially wetlands and lentic water bodies (such as oxbow lakes) that lack
surface channel connections to stream networks, attenuate downstream flood pulses.
Heterogeneous riparian areas that include wetlands and open waters remove large amounts of sediment
and nutrients from upland areas before they can enter the stream network.
Riparian areas influence stream geomorphology during periodic flooding by releasing stored sediments.
Forested riparian areas provide woody debris that helps shape stream morphology.
Riparian vegetation shades the stream and influences and regulates stream temperature and stream net
primary productivity.
Ground water that flows through riparian areas and into the stream helps moderate stream temperatures.
Chemical Connectivity and Function
Riparian areas, acting as buffers, are critical to protecting stream-water quality.
The structure of the riparian area (e.g., vegetation, wetlands, redox potential) influences its ability to
increase water quality before it reaches the stream.
The near-stream portion of a riparian area is often more important in protecting stream-water quality than is
the near-field (near uplands) portion.
Allochthonous inputs generally are most important to food webs in small headwater streams, especially in
forested areas. As rivers become larger, primary production becomes increasingly important.
Some of the best-documented functions of oxbow lakes are as sinks for nutrients from upland runoff that
might otherwise flow into rivers.
Biological Connectivity and Function
Many types of organisms move between riparian/floodplain wetlands and the river network; those
transported by water often move in response to flooding and those transported by other mechanisms (e.g.,
wind) move in response to seasonal cues or life-history stage requirements.
Riparian/floodplain wetlands and oxbow lakes can be sources or sinks of organisms; one of the most
important source functions is to provide rearing habitat for fish.
Riparian/floodplain wetlands provide food sources for stream and river invertebrates.
Many riparian/floodplain wetlands and open waters (e.g., oxbow lakes) are used by fish and other
organisms from the stream or river during flood ing.
output channels but are isolated from the river network; geographically isolated wetlands (i.e., wetlands
completely surrounded by uplands) that have local or regional ground-water or occasional surface-
water connections; and geographically isolated wetlands that have minimal hydrologic connection to the
river network (but which could include surface and subsurface connections to other wetlands).
Based on our literature review and basic hydrologic principles, we conclude thatnon-floodplain
wetlands that are connected to the river network through surface water will have an influence on
downstream waters, regardless of whether the outflow is permanent, intermittent, or ephemeral. Such
non-floodplain wetlands include wetlands that are the origins of streams or are connected downstream
to the river network through ditches. They also would include geographically isolated wetlands that are
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connected downstream to the river network through upland swales. Further, although the literature
review did not address other non-floodplain water bodies to the same extent as wetlands, our overall
conclusions also apply to these water bodies (e.g., ponds and lakes that lack surface water inlets)
because the same principles govern hydrologic connectivity between these water bodies and
downstream waters (Chapter 2).
Non-floodplain wetlands that do not connect to the river network through surface water include
wetlands that spill into losing streams that are completely disconnected from the river network; that is,
the wetland exports water through an output channel but the water is completely lost before it reaches
the river network due to evapotranspiration or loss to ground water. Also included are geographically
isolated wetlands that either do not spill, or spill into an upland swale that does not enter the river
network. Although such wetlands lack surface-water connections to streams and rivers, they can be
connected through local, intermediate, or regional ground-water flows or through biological movement.
Connectivity between these wetlands and downstream waters will vary within a watershed as a function
of local factors (e.g., position, topography, and soil characteristics; Sections 2.4.1 and 2.4.2), some of
which are identified and discussed in this section. Connectivity also will vary over time, as the river
network and water table expand and contract in response to local climate.
It is difficult to generalize about the specific downstream effects of non-floodplain wetlands that lack
surface water connections to downstream waters. In Chapter 2 we note that the influence of wetlands
and streams on downstream waters depends on two factors: (1) functions that affect material fluxes and
(2) connectivity (or isolation) that allows (or prevents) transport of materials between the systems
(Section 2.3). The literature we reviewed and summarized provides ample evidence that non-floodplain
wetlands provide hydrological, chemical, and biological functions that affect material fluxes. Thus, these
wetlands could affect downstream waters if they are connected to (or isolated from) the river network
in such a way that it allows (or prevents) transport of materials to downstream waters. However, the
more than 200 peer-reviewed references on non-floodplain wetlands we reviewed infrequently
evaluated connections between non-floodplain wetlands and river networks and rarely examined the
frequency, duration, magnitude, timing, and rate of change of these connections. Even if it is known from
an article that the study site is located near a downstream water, connectivity cannot be established
without specific information on frequency and magnitude of precipitation events, soil infiltration rate,
wetland storage capacity, hydraulic gradients, etc.—information that is only rarely available in
publications. Thus, the literature provided no evaluations of connectivity for specific groups or classes of
wetlands (e.g., prairie potholes or vernal pools). This lack of information applies to groups of these
wetlands within a particular watershed and to comparisons between different types of regional
wetlands. For example, our review did not reveal whether connectivity between vernal pools and
downstream waters is greater than connectivity between prairie potholes and downstream waters. We
emphasize that this does not mean these wetlands do or do not have connectivity with downstream
waters: It simply means the literature we reviewed does not enable us to distinguish connectivity of
these wetland types from each other. Literature that was not included in our review, such as reports
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Table 4-4. Key conclusions on the effects of non-floodplain wetlands on rivers.
Physical Connectivity and Function
The connections of non-floodplain wetlands with downstream waters exist along a spectrum from isolated
depressional wetlands, to those connected through ground water, to those connected via intermittent or
permanent surface flows.
The degree to which outputs (or connections) are dominated by surface water vs. ground water is controlled
in part by soil permeability: Permeable soils favor ground-water outputs, while impermeable soils result in
surface water outputs. Other factors, such as topographic setting, also can play a role.
Ground-water recharge is common in non-floodplain wetlands and can be a particularly important source of
water to aquifers during dry periods.
Ground-water networks extend from the local to the intermediate and regional scales, and provide a
mechanism by which non-floodplain wetlands can influence other water bodies over various periods.
Even when non-floodplain wetlands lack a connection to other water bodies, they can influence downstream
water through water storage and mitigation of peak flows (flood reduction and attenuation).
Chemical Connectivity and Function
Insofar as they often act as buffers between sources of pollution and riparian areas, non-floodplain
wetlands are a "first line of defense" in protecting streams from polluted waters.
Non-floodplain wetlands affect nutrient delivery and water quality.
Non-floodplain wetlands are a principal source for dissolved organic carbon (which supports primary
productivity) to some downstream waters; the area of a basin with non-floodplain wetlands is directly
correlated to the contribution of that basin to dissolved organic carbon in downstream waters.
Non-floodplain wetlands are sources of mercury: Microbial processes in non-floodplain wetlands methylate
mercury, which can be translocated through near-surface and surface flows to downstream waters where it
can bioaccumulate.
Non-floodplain wetlands are sinks for sediment, nutrients (including phosphorus, nitrate, and ammonium),
metals (e.g., nickel and cobalt), and pesticides (e.g., atrazine).
Non-floodplain wetlands can remove, retain, or transform many of the nutrient inputs to which they are
exposed.
Biological Connectivity and Function
Natural periodic and permanent human-engineered surface-water connections can connect biological
communities in non-floodplain wetlands and the river network; in addition, wind dispersal and overland
movement connect these types of water bodies with frequency decreasing as a function of distance,
landscape barriers, or both.
Migratory birds are vectors of plants and invertebrates between non-floodplain wetlands and the river
network, although their influence has not been quantified fully.
Non-floodplain wetlands promote biological interactions that can be critical to the life-history requirements
of some stream species.
Overland ("fill-and-spill") hydrologic connections can support biological connections. For example, stream
fish found in wetlands that periodically dry down indicate presence of surf ace flows sufficient for
colonization.
from local resource agencies, could allow the connectivity of these wetlands to be evaluated further, as
could analysis of existing or new data or field evaluation.
Further complicating our evaluation is that some of the effects that wetlands have on downstream
waters are due to their isolation, rather than their connectivity. Wetland functions that trap materials
and prevent their export to downstream waters (e.g., sediment and entrained pollutant removal, water
storage) result because of the wetland's ability to isolate material fluxes. As above, to establish that a
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wetland influences a downstream water through its isolation, it would have to be known that the
wetland intercepted materials that would otherwise reach the downstream water, and this information
is typically not provided in publications. The literature we reviewed does provide limited examples of
the direct effects of such isolation on downstream waters for some specific wetlands, but not for classes
of wetlands (e.g., vernal pools). However, the literature we reviewed allows us to conclude that sink
functions of non-floodplain wetlands, which result in part from their relative isolation, will have effects
on a downstream water when these wetlands are situated between the downstream water and known
point or nonpoint sources of pollution, and thus intersect the flowpath between pollutant source and
downstream water. For example, in cases where agricultural land use is a known contributor of
sediment to downstream waters, the presence of depressional wetlands along the flowpath between the
agricultural land and downstream water will result in reduced sediment loading to the downstream
water. These effects would also be realized from sink functions that do not result from the wetland's
isolation perse, but are emergent wetland properties (e.g., biogeochemical reactivity based on anoxic
conditions). Using the same example, if the agricultural land use is a known contributor of nitrogen to
downstream waters, depressional wetlands occurring along the flowpath will result in reduced nitrogen
loading to the downstream water. In such settings, wetland loss or increased connectivity (e.g., due to
ditching or tiling) is likely to reduce the effects of such functions on downstream waters (although
functions that depend on connectivity could be increased).
To provide more specific evaluations of the connectivity of non-floodplain wetlands to downstream
waters, studies are needed that: (1) further develop and validate methods for assessing wetland and
watershed connectivity; (2) apply such methods to different classes of non-floodplain wetlands,
especially those that lack channelized surface-water or regular shallow subsurface-water connections;
(3) evaluate the frequency, duration, magnitude, timing, and rate of change of water, material, and biotic
fluxes to downstream waters; and (4) consider aggregate functions and connectivity of wetland
complexes (groups of closely located and interacting wetlands). Such studies are needed throughout the
country to cover the breadth of wetlands in non-floodplain settings satisfactorily (e.g., across areas with
different climate, geology, and terrain).
Despite these limitations, we can make some conclusions:
1. A non-floodplain wetland having a surface-water outflow to a stream network (e.g., a wetland
that serves as a stream origin) is connected to the stream network and has an influence on
downstream waters.
2. Many non-floodplain wetlands interact with ground water, which can travel long distances and
affect downstream waters.
3. Even when wetlands lack a hydrologic connection to other water bodies, they can influence
downstream water through water and material storage and mitigation of peak flows (flood
reduction and flood attenuation). Sink functions of non-floodplain wetlands will have effects on
a downstream water when these wetlands are situated between the downstream water and
known point or nonpoint sources of pollution, thereby intersecting the flowpath between
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pollutant source and downstream water. More generally, wetland sink functions are likely to be
greatest when the wetland is located downgradient from pollutant sources and upgradient from
a stream or river.
4. Non-floodplain wetlands provide unique and important habitats for many species, both common
and rare. Some of these species require multiple types of waters to complete their full life cycles,
including downstream waters. Abundant or highly mobile species play important roles in
transferring energy and materials between non-floodplain wetlands and downstream waters.
5. Biological connections are likely to occur between most non-floodplain wetlands and
downstream waters through either direct or stepping stone movement of amphibians,
invertebrates, reptiles, mammals, and seeds of aquatic plants, including colonization by invasive
species. Many species in those groups that use both stream and wetland habitats are capable of
dispersal distances equal to or greater than distances between many wetlands and river
networks. Migratory birds can be an important vector of long-distance dispersal of plants and
invertebrates between non-floodplain wetlands and the river network, although their influence
has not been quantified. Whether those connections are of sufficient magnitude to impact
downstream waters will either require estimation of the magnitude of material fluxes or
evidence that these movements of organisms are required for the survival and persistence of
biota that contribute to the integrity of downstream waters.
6. Spatial proximity is one important determinant of the magnitude, frequency and duration of
connections between wetlands and streams that will ultimately influence the fluxes of water,
materials and biota between wetlands and downstream waters. However, proximity alone is
not sufficient to determine connectivity, due to local variation in factors such as slope and
permeability.
7. The cumulative influence of many individual wetlands within watersheds can strongly affect the
spatial scale, magnitude, frequency, and duration of hydrologic, biological and chemical fluxes or
transfers of water and materials to downstream waters. Because of their aggregated influence,
any evaluation of changes to individual wetlands should be considered in the context of past and
predicted changes (e.g., from climate change) to other wetlands within the same watershed
8. Caution should be used in interpreting connectivity for wetlands that have been designated as
"geographically isolated" because
a. the term can be applied broadly to a heterogeneous group of wetlands, which can include
wetlands that are not actually geographically isolated (e.g., some vernal pools are not
geographically isolated because they have output channels;
b. wetlands with permanent channels could be miscategorized as geographically isolated if the
designation is based on maps or imagery with inadequate spatial resolution, obscured
views, etc.; and
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c. wetland complexes could have connections to downstream waters through stream channels
even if individual wetlands within the complex are geographically isolated.
Thus, the term "geographically isolated" should be applied only to groups of wetlands if all those
wetlands are, in fact, known to be geographically isolated, something that we cannot determine
based on this literature review. As previously noted, additional information that was not
included in our literature review (e.g., reports from local resource agencies, analysis of existing
or new data, field evaluations) could allow some wetlands that are truly geographically isolated
to be distinguished from some of those that are not. Further, even geographically isolated
wetlands can be connected to other wetlands and downstream waters through ground-water
connections, occasional spillage, or biological connections. Thus, the term "geographically
isolated" should not be used to infer lack of hydrologic, chemical, or biological connectivity. Key
conclusions from our literature review on non-floodplain wetlands are summarized in
Table 4-4.
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CHAPTER
J: CONNECTIVITY CASE
STUDIES
5.1 Introduction
Chapters 3 and 4 of this report review evidence from the literature for the physical, chemical, and
biological connections of three broad categories of waters—streams, riparian/floodplain wetlands, and
non-floodplain wetlands—to, and their resulting effects on, downstream waters. In addition to the three
questions in Table 1-1, the EPA's Office of Water asked us to provide detailed information on six specific
water body types: Carolina and Delmarva bays, oxbow lakes, prairie potholes, prairie streams,
southwestern streams, and vernal pools (Appendix B).
In this chapter, we summarize the results of the six case studies, applying the concepts in Chapters 1 and
2 to the detailed evidence in Appendix B, for each habitat. The full body of evidence and supporting
citations, which we omitted here to improve readability, are provided in Appendix B. We summarize
evidence from the individual case studies in terms of (1) the descriptors of connectivity (i.e., the
frequency, duration, magnitude, timing, rate of change effluxes to and biological exchanges with
downstream waters; Section 1.2.2); (2) the consequences of different systems and degrees of
connectivity on downstream waters (Sections 1.2.3 and 2.3); (3) and the effects of typical human
alterations (Sections 1.2.4 and 2.4.4). We then use the information from these case studies and from
Chapters 3 and 4 to illustrate, hypothetically, where streams, riparian/floodplain wetlands, and non-
floodplain wetlands are positioned along a connectivity gradient, highlighting the primary lines of
evidence that support that positioning.
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5.2 Carolina and Delmarva Bays
Carolina bays are elliptical, ponded, depressional wetlands that occur along the Atlantic Coastal Plain
from northern Florida to New Jersey, although they are most abundant in North Carolina and South
Carolina (Section B.I). Carolina bays that are geographically specific to the Delmarva Peninsula are often
referred to as Delmarva bays. Carolina and Delmarva bays range in size from less than 1 ha to greater
than 3,600 ha and are densely concentrated in many areas. In the 1950s, roughly 500,000 bays existed,
although the number today is markedly less due to human modification of the landscape. Bays primarily
gain water from direct precipitation on their surfaces (with some water deriving from inlet channels,
surface runoff, shallow ground water, and natural springs) and lose water through evapotranspiration.
As a result, these relatively permanent bays experience fluctuating water levels. Their extensive
distribution and wet-dry cycles promote and support a diverse biota.
5.2.1 Connectivity and Consequences on Downstream Waters
Some Carolina and Delmarva bays connect to each other and some connect to downstream waters.
Delmarva bays inundate seasonally and connect hydrologically to other bays and to stream networks via
intermittent stream channels. Studies also document shallow ground-water connections, via both nearly
continuous shallow ground-water recharge and periodic shallow ground-water discharge.
When they occur, hydrologic connections are likely to result in effects on downstream waters. Carolina
and Delmarva bays can reduce the amount of nitrate transported between surface-water systems and
ground water via denitrification, which is promoted by the periodicity of wetting and drying that occurs
in bays, and dilution. Seasonal connections of Delmarva bays to stream networks export accumulated
organic matter from wetlands into tributaries of Chesapeake Bay. Hydrologic connections also export
methylmercury from these systems (see below).
Although the current published evidence for biological connections is limited and primarily indirect, the
potential for movement of organisms between bays and other water bodies is high. These bays provide
valuable habitat and food web support for numerous plant and animal species. Fish presence in bays
known to dry out periodically indirectly demonstrates that these bays must be connected to other
waters. Amphibians and reptiles use bays extensively for breeding and for rearing young. In bays that
lack fish, the absence of predators allows abundant amphibian populations to thrive, particularly those
with aquatic larval stages. These animals can then disperse many meters across the landscape and
colonize downstream waters. Bays also foster abundant aquatic insects, and their emergence can have
consequences for nearby waters. Many species documented in Carolina and Delmarva bays are known
to live in pond, wetland, and stream environments. As a result, species emerging from bays can become
important food sources for organisms in nearby streams after aerial or terrestrial dispersal. Cumulative
emergence from thousands of small bays across the landscape could create a significant food source for
downstream waters.
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5.2.2 Effects of Human Alteration
Human alteration of Carolina and Delmarva bays has affected their physical, chemical, and biological
connections to, and effects on, downstream waters. Agriculture, logging, and other human activities have
altered the vast majority of Carolina and Delmarva bays, affecting the frequency, duration, magnitude,
and timing of hydrologic connections between bays and other waters. Agricultural practices have greatly
reduced the number of bays over the past several decades. Channelization and ditching of bays for
agriculture is common. Draining bays for agricultural use disrupts or alters numerous wetland
functions: sediment and chemical storage and transformation, biological habitat and sources, and
organic matter export. Because the ditches commonly connect the surface water of bays that drain
agricultural fields to stream networks that drain into downstream water bodies, they serve as
conveyances for nutrients, sediment, and contaminants—thereby increasing physical, chemical, and
biological connections between bays and the downstream systems. The consequences of this increased
connectivity for downstream waters can be especially important in terms of nutrient and contaminant
transport. In addition to runoff from farmed fields, periodic drying and flooding of shallow Carolina and
Delmarva bays promote the bacteria-mediated methylation of mercury. Subsequent transport of
bioavailable methylmercury through ditches can pose a contamination risk to fish and piscivorous birds
inhabiting downstream water bodies.
5.3 Oxbow Lakes
Oxbow lakes are natural features of floodplains, originating from curves (meanders) in the river that
become cut off from the active river channel (Section B.2). They are located in flat, unconstrained
floodplains of river systems.
5.3.1 Connectivity and Consequences on Downstream Waters
The evidence for physical, chemical, and biological connectivity of oxbow lakes to downstream waters is
considerable. Because of their location within river floodplains, many oxbows are connected seasonally
or episodically to downstream waters during natural flood events via surface and shallow subsurface
flows. The frequency, duration, magnitude, and timing of these hydrologic connections depend on river
stage, lake geomorphology, and relative position along and distance from the river network. Despite this
spatial and temporal variability, oxbow lakes collectively are likely to influence downstream waters.
The frequency, magnitude, and duration of physical connection between oxbow lakes and the river
channel have important consequences on the river network. Physical surface connections facilitate
biological and chemical exchange between oxbow lakes and rivers. Oxbow lakes function as sinks,
because they intercept and store nutrients and other materials from upland runoff that otherwise would
flow directly into the river network. In these cases, the lack of a permanent connection between an
oxbow lake and a river helps to preserve the chemical integrity of the river network.
When oxbow lakes are connected, the biological material produced within them can subsidize riverine
food webs by passive or active transport from the lake to downstream waters. Oxbow lakes are
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important areas of biological productivity on floodplains. Periodic surface-water connections between
rivers and oxbow lakes facilitate the movement offish, allowing riverine fish to exploit these productive
floodplain water bodies before they move back to the river. In this way, connectivity between oxbow
lakes and rivers supports the biological integrity of the river network.
5.3.2 Effects of Human Alteration
Human alterations of the natural flow regime in rivers can influence connectivity between oxbow lakes
and the active river channel. In some cases, permanent channels are constructed between oxbows and
the river channel and connectivity is increased; in other cases, such as the creation of dikes or levees,
connectivity is reduced between oxbows and the altered area of the river network. Practices that alter
the natural flow regime of the river (e.g., dams) or inhibit periodic flooding of oxbow lakes (e.g., levees)
affect movement of water and sediment and the use of oxbow lakes by riverine fish. When cut off from
periodic inundation by the river channel, water in oxbow lakes can evaporate. Over time, these lakes can
dry up, be colonized by terrestrial vegetation, and eventually become dry land.
5.4 Prairie Potholes
The prairie pothole region, located in northern-central North America, is named for the abundant,
glacially formed wetlands that occur throughout the region, typically as depressions lacking natural
outlets (Section B.3). The prairie pothole region covers approximately 777,000 km2, a vast area that
varies in climate, terrain, geology, land use, and human alteration. These variations result in a gradient
of connectivity to and effects on downstream waters across the potholes themselves. For instance, the
three major physiographic areas within the prairie pothole region (Red River Valley, Drift Prairie, and
Missouri Coteau) vary in precipitation, distribution, and density of potholes and streams connecting
potholes to downstream waters. Potholes exhibit a wide range of hydrologic permanence, from holding
permanent standing water to wetting only in years with high precipitation. Differences in the frequency,
duration, and timing of pothole inundation across the region influence wetland function and the
diversity and structure of their biological communities.
5.4.1 Connectivity and Consequences on Downstream Waters
Individual prairie potholes span the continuum of isolation from and connection to the river network
and other water bodies. In addition to differences among individual potholes, interactions between
regional factors (e.g., precipitation) and local factors (e.g., landscape relief) can result in spatial patterns
of connectivity across the landscape (Sections 2.4.5 and B.3.2.1) that have consequences for the
downstream connectivity and effects of prairie potholes. Considered collectively, unaltered prairie
pothole systems have infrequent direct surface-water connections to downstream waters. Evidence of
the consequences of these connections on downstream waters is variable. Some studies document
measurable effects of water storage capacity of potholes on flood attenuation and maintenance of
stream baseflow, whereas other studies show no effect of pothole water storage on streamflows. These
differences in observed effects might be explained, in part, by the spatial variation observed within the
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prairie pothole region. Potholes can connect to downstream waters via ground-water flows when both
are within a continuous zone of a shallow local aquifer. In areas with restricted surface-water and
ground-water interactions, the magnitude of effects from such connections will be small.
The chemical connectivity of prairie potholes is largely mediated by their hydrologic connectivity. As
depressions on the landscape, potholes tend to accumulate nutrients, sediment, and pesticides that can
be chemically transformed and decrease potential effects on downstream waters (e.g., denitrification
frequently occurs in saturated pothole sediments). Although chemical sink (storage) functions and
periodic source functions of potholes have been documented in the literature, their overall influence on
lakes and river networks has been difficult to quantify. This difficulty exists in part because altered and
unaltered potholes co-occur in watersheds with different land use and management practices, and many
different parts of this complex landscape can affect the integrity of downstream waters. Thus, prairie
potholes can have substantial hydrologic and chemical consequences on downstream water levels and
flows, but this type of connectivity and its downstream effects are difficult to predict, demonstrate, and
quantify.
Although direct evidence is sparse, indirect evidence suggests that prairie potholes are highly
biologically connected. Prairie pothole systems have biological connections to downstream waters via
annual bird migrations—especially for migratory waterfowl such as cranes, geese, ducks, and
shorebirds, which actively move between and use multiple aquatic habitats, including prairie pothole
systems. For instance, the prairie pothole region has been identified as an area of global and regional
importance for migratory birds, and at least 15 duck species use prairie pothole wetlands. Mammals and
many species of amphibians also use potholes. Plants and invertebrates disperse to and from prairie
potholes via "hitchhiking" on waterfowl. That potholes lack an endemic aquatic and semiaquatic flora
and fauna indicates that communities in potholes are biologically well connected with other aquatic
ecosystems, but evidence for effects of biological connections on downstream waters is limited.
5.4.2 Effects of Human Alteration
Human alterations of the landscape affect the connectivity of prairie potholes. Land use in an upland
that drains to a wetland can alter the amount of runoff that wetland receives. Much of Upper Midwest
cropland is artificially drained to increase agricultural productivity. Filling potholes and lowering the
regional water table through agriculture tile drainage have increased the isolation of remaining potholes
by decreasing the density of depressions containing water. In some areas, extensive surface draining
and ditching has directly and dramatically increased connectivity between pothole basins and the river
network. Ditches create surface-water outlets from potholes, connecting potholes to streams and rivers;
drains and underground pipes fitted at the bottoms of potholes often discharge to open ditches or
streams. This increased hydrologic and chemical connectivity decreases water retention time, thereby
reducing storage and biogeochemical processing of nutrients, sediments, and pesticides. The cumulative
influence of human alterations on connectivity between potholes and downstream waters has not been
systematically studied or reported across the entire prairie pothole region.
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5.5 Prairie Streams
Prairie streams drain temperate grasslands in the Great Plains physiographic region of the central
United States and Canada (Section B.4). Eventually, these streams drain into the Mississippi River or
flow directly into the Gulf of Mexico or the Hudson Bay. Climate in the Great Plains region ranges from
semiarid to moist subhumid and intra- and interannual variation in precipitation and
evapotranspiration is high. This variation is reflected in the hydrology of prairie streams, which include
ephemeral, intermittent, and perennial streamflows. Row cropping and livestock agriculture are the
dominant land uses in the region, resulting in the withdrawal of water from stream channels and
regional aquifers and its storage in reservoirs to support agriculture.
5.5.1 Connectivity and Consequences on Downstream Waters
Prairie streams typically are connected to downstream waters. Like other types of streams, prairie
streams present strong fluvial geomorphic evidence for connectivity to downstream waters, in that they
have continuous channels (bed and banks) that make them physically contiguous with downstream
waters. Prairie river networks are dendritic and generally have a high drainage density, so they are
particularly efficient at transferring water and materials to downstream waters. Their pool-riffle
morphology, high sinuosity, and seasonal drying, however, also enhance material storage and
transformation. The timing of connections between prairie streams and downstream waters is seasonal
and therefore relatively predictable. For example, high-magnitude floods tend to occur in late fall into
later spring, although they also occur at other times during the year (Section B.4.2.1); this observation
indicates that the magnitude of connections to downstream also varies seasonally.
The frequent and predictable connections between prairie streams and downstream waters have
multiple physical, chemical, and biological consequences for downstream waters. Dissolved solids,
sediment, and nutrients are exported from the prairie river network to downstream waters. Ultimately,
the expansion of the hypoxic zone in the Gulf of Mexico is a downstream consequence of cumulative
nutrient loading to the Mississippi River network. Relative to small streams and large rivers draining the
moist eastern parts of the Mississippi River basin, small to midsized prairie streams deliver less than
25-50% of their nutrient load to the Gulf of Mexico. Nonetheless, given the large number and spatial
extent of headwater prairie streams connected to the Mississippi River, their cumulative effect likely
contributes substantially to downstream nutrient loading.
Organisms inhabiting prairie streams have adapted to their variable hydrologic regimes and harsh
physicochemical conditions via evolutionary strategies that include rapid growth, high dispersal ability,
resistant life stages, fractional reproduction, and life cycles timed to avoid predictably harsh periods.
Alterations in the frequency, duration, magnitude, and timing of flows—and thus hydrologic
connectivity—are associated with the extinction or extirpation of species in downstream systems.
Moreover, many fish species (e.g., Arkansas River shiner, speckled chub, flathead chub) in prairie river
networks require sufficient unfragmented (i.e., connected) channel length with adequate discharge to
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keep their nonadhesive, semibuoyant eggs in suspension for incubation and early development. When
these conditions are not met, the biological integrity of downstream waters is impaired.
5.5.2 Effects of Human Alteration
Human alteration of prairie river networks has affected the physical, chemical, and biological
connectivity to and their consequences for downstream waters. Impoundments and water removal,
through both surface flow diversions and pumping of ground-water aquifers, are common in this region.
These activities have reduced flood magnitude and variability, altered timing, and increased
predictability of flows to downstream waters. As a result, physical, chemical, and biological connections
to downstream waters have been altered. In addition to the altered land uses and application of
nutrients and pesticides for agriculture, human alteration of the river network itself, through
channelization, levee construction, desnagging, dredging, and ditching, has enhanced longitudinal
connectivity while reducing lateral and vertical connectivity with the floodplain and hyporheic zone,
respectively. Pumping from streams and ground water has caused historically perennial river segments
to regularly dry during summer months. Changes to the prairie's grazing (from bison to cattle) and
burning regimes increase nutrient and suspended sediment loading to downstream waters. Introduced
species have extirpated endemic species and altered food web structure and processes in prairie
streams, thereby affecting the biological integrity of downstream waters.
5.6 Southwestern Intermittent and Ephemeral Streams
Southwestern streams are predominantly ephemeral and intermittent (nonperennial) systems located
in the southwestern United States (Section B.5). Based on the National Hydrography Dataset, 94%, 89%,
88%, and 79% of the streams in Arizona, Nevada, New Mexico, and Utah, respectively, are nonperennial.
Mostof these streams connect to downstream waters, although 66% and 20% of the drainage basins in
Nevada and New Mexico, respectively, are closed and drain into playas (dry lakes). Southwestern
streams generally are steep and can be divided into two main types: (1) mountainous streams that drain
higher portions of basins and receive higher rates of precipitation, often as snow, compared to lower
elevations; and (2) streams located in valley or plateau regions that generally flow in response to high-
intensity thunderstorms. Headwater streams are common in both types of southwestern streams.
5.6.1 Connectivity and Consequences on Downstream Waters
Nonperennial southwestern streams, excluding those that drain into playas, are periodically connected
to downstream waters by low-duration, high-magnitude flows. In contrast to streams in humid regions
where discharge is typically supplemented by ground water as drainage area increases, many
southwestern streams lose streamflow to channel transmission losses as runoff travels downstream
(Figure B-10). Connection of runoff and associated materials in ephemeral and intermittent streams to
downstream waters is therefore a function of distance, the relative magnitude of the runoff event, and
transmission losses.
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Spatial and temporal variation in frequency, duration, and timing of southwestern stream runoff is
largely explained by elevation, climate, channel substrate, geology, and the presence of shallow ground
water. In nonconstraining substrate, southwestern rivers are dendritic and their watersheds tend to
have a high drainage density. When high flows are present, southwestern streams are efficient at
transferring water, sediment, and nutrients to downstream reaches. Due to the episodic nature of flow
in ephemeral and intermittent channels, sediment and organic matter can be deposited some distance
downstream, and then moved farther downstream by subsequent precipitation events. Over time,
sediment and organic matter continue to move downstream and affect downstream waters.
The southwestern streams case study (Section B.5) describes the substantial connection and important
consequences of runoff, nutrients, and particulate matter originating from ephemeral tributaries on the
integrity and sustainability of downstream perennial streams. Channel transmission losses can be an
important source of ground-water recharge that sustains downstream perennial stream and riparian
systems. For example, isotopic studies indicate that runoff from ephemeral tributaries like Walnut
Gulch, Arizona supplies roughly half the San Pedro River's baseflow through shallow alluvial aquifer
recharge.
5.6.2 Effects of Human Alteration
Human alterations to southwestern river networks affect the physical, chemical, and biological
connectivity to downstream waters. Impoundments trap water, sediment, and particulate nutrients and
result in downstream impacts on channel morphology and aquatic function. Diversion of water for
consumptive uses can decrease downstream baseflows but typically does not affect the magnitude of
peak flows. Excessive ground-water pumping can lower ground-water tables, thereby diminishing or
eliminating baseflows. Urbanization increases runoff volume and flow velocity, resulting in more erosive
energy that can cause bank erosion, streambed downcutting, and reduced infiltration to ground water.
5.7 Vernal Pools
Vernal pools are shallow, rain-fed, fishless pools situated on bedrock or low-permeability soils
(Section B.6). Vernal pools inundate seasonally and lack continuous surface-water connections to
downstream water bodies. Although they can occur in other parts of the United States, this case study
focuses on pools in the western states and the glaciated areas of northeastern states. Western vernal
pools typically occur in open grasslands; most northern vernal pools are detrital and are fully contained
within forest ecosystems. When inundation occurs, vernal pools can fill and overflow through swales or
intermittent streams, which connect them to downstream waters.
5.7.1 Connectivity and Consequences on Downstream Waters
Direct surface connection of vernal pools to downstream waters is infrequent. The duration and
magnitude of such connections are highly variable and depend on the climate, terrain, and geology of the
region and on the location of the vernal pool in the watershed. Vernal pools generally are clustered,
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forming wetland complexes. Pools located at the downgradient end of a complex can receive surface
water through stepping-stone spillage in addition to precipitation, and generally are inundated longer
than upper pools. Because they experience greater inundation and are likely to be located nearer to
streams, these downgradient pools are also more likely to be directly connected to streams. Temporary
storage of heavy rainfall and snowmelt in individually small vernal pool systems (pools plus soils) can
attenuate flooding, provide a reservoir for nearby vegetation during the spring growth period, and
increase nutrient availability.
The timing of seasonal inundation and lack of permanent surface connections make vernal pools
important biological refuges, which has consequences on the biological health of downstream waters.
Vernal pools are highly productive ecosystems that have evolved in a "balance between isolation and
connectedness" (Zedler, 2003; page 597). Because they are connected to other aquatic habitats through
dispersal, they provide rich reservoirs of genetic and species diversity. Food webs in vernal pools
include highly fecund amphibians and insects that convert detrital organic matter into biomass, which is
then exported to aquatic ecosystems in other parts of the watershed. Northern vernal pools can provide
alternative breeding habitat, refuge from predators or environmental stressors, hunting or foraging
habitat, or stepping-stone corridors for dispersal and migration.
5.7.2 Effects of Human Alteration
Vernal pools have been drained and converted to other land uses (e.g., agriculture, logging, urban
development). These activities have increased fragmentation of habitats for amphibians, plants, and
invertebrates, and had similar effects on the frequency, duration, magnitude, and timing of inundations,
surface-water outflows, and shallow subsurface-water connections to downstream waters as those
described in Section 5.2.1 (Carolina and Delmarva bays).
5.8 Synthesis
These case study summaries highlight the key connections between specific water body types and
downstream waters. The case study evidence provides further support that the structure and function of
downstream waters highly depend on constituent materials and organisms contributed by and
transported through water bodies located throughout the watershed. In addition, the studies support
that variation in the types and degrees of connectivity determines the range of downstream effects.
These case study summaries illustrate two key points. First, each type of water body addressed here
demonstrates variability in connectivity to and effects on downstream waters. Oxbow lakes, for
example, are more or less connected to the main river channel based largely on their relative position in
the landscape: Systems close to the river channel are highly connected and those farther away are
connected less often or the impact on the river takes longer to be realized. Evidence presented in the
prairie pothole case study also demonstrates variation in connectivity patterns across the region and
shows the consequences of this variability on downstream rivers and lakes. The prairie streams case
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study discusses functions and varying degrees of connectivity of streams and their cumulative effects on
downstream waters.
Second, the effects of human alteration on the connectivity to and effects on downstream waters depend
on the type of water body. Human alteration of different types of streams and wetlands can be complex,
either increasing or decreasing connectivity and subsequent effects on downstream waters. For
example, evidence shows that ditches in the prairie pothole region increase hydrologic connectivity, and
connectivity of oxbow lakes near active river channels can be reduced if that portion of the river is
leveed. Coupled human-natural systems are an area of active research and new information about the
effects of human activities on connectivity and water integrity is emerging in the peer-reviewed
literature.
Positioning the specific water body types in the case studies (Appendix B) along a gradient of
connectivity and effect proved to be premature for several reasons. First, the amount of documented
evidence (i.e., number of published studies) varied among the water body types. In some instances, a
large body of evidence exists and in others, only a few studies exist, limiting sound comparisons. Second,
variation in connectivity consistently was reported to be high within some water body types, creating
substantial overlap in ranges of connectivity among those water body types. In addition to a need for
more studies documenting connectivity in less studied regions, a more refined classification using the
descriptors of connectivity described in Chapter 1 (or others) and their controls (e.g., climate, geology,
and terrain) within wetland landscape settings are required.
Based on the evidence presented in Chapters 3 and 4, ordering the three broad categories of water
bodies considered in this report—streams, floodplain wetlands, and non-floodplain wetlands—along a
connectivity gradient (Figure 5-1) is possible. Of these three water body types, streams are, in general,
more connected to and have better-documented effects on downstream waters than either wetland
category. Floodplain wetlands, in turn, tend to be more connected to downstream waters, and have
better-documented downstream effects, than non-floodplain wetlands. This ordering must be
recognized as a broad generalization, and considerable overlap can occur among the types, given the
spatial and temporal variability in connectivity documented in these habitats (Figure 5-1). Nevertheless,
several key lines of evidence support this hypothesized ordering of water body types along the gradient.
1. Streams are connected to rivers by a continuous channel, which is a physical reflection of
surface connectivity. Formation of a channel indicates that connectivity, in terms of its combined
descriptors (frequency, duration, magnitude, timing) is sufficiently strong (or "effective") and
outweighs terrestrialization processes (e.g., revegetation, wind-mediated processes, soil
formation processes).
2. Within-channel flows are more efficient for moving water, sediment, pollutants, and other
materials than overland flow; for some aquatic organisms, channels are the only possible
transport routes. Channels are places where excess water and materials from the landscape are
concentrated as they are transmitted downstream. Recurrent flow of sufficient magnitude over a
given area of landscape selects routes with least resistance, which develop into branched
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channel networks with a repeating, cumulative pattern of smaller channels that join at
confluences to form larger channels.
3. The continuous channels connecting streams to rivers also represent areas of relatively high
shallow subsurface connectivity (shallow ground-water recharge and upwelling). Channels are
typically more permeable than surrounding soils, lack dense terrestrial vegetation (and thus
have lower uptake and evapotranspiration loss), and are topographic low points closer to
concentrated shallow ground water.
4. Floodplain wetlands and open waters are connected to rivers by historical and recurrent surface
connectivity. Riparian/floodplain wetlands are maintained by the recurrent inundation and
deposition of materials from streams and rivers during the peak and recession of flood flows.
5. Riparian/floodplain wetlands and open waters are close to river networks and thus more likely
to have strong connectivity with the downstream water than more distant wetlands, when all
other conditions are similar.
6. Non-floodplain wetlands are positioned outside the floodplain, and so are not subject to direct
flooding from the river or stream. Any hydrologic connections to the river system are therefore
unidirectional (from wetland to downstream water and not vice-versa). They are also likely to
be more distant from the network, increasing the flowpath lengths and travel time to the
network.
7. Because of their large numbers, headwater streams and associated wetlands cumulatively
represent a large portion of the landscape interface with a downstream water. These areas
provide functions that enhance both exchanges with and buffering of the downstream water,
making them critical to mediating the recognized relationship between the integrity of
downstream waters and the land use and stressor loadings from the surrounding landscape.
8. Connectivity to downstream waters is reflected in the distribution of aquatic organisms and
their dependence on particular aquatic habitats across different stages of their life cycles. For
example, the recurrent presence of completely aquatic organisms (i.e., organisms that lack
terrestrial life stages, overland dispersal, stages resistant to drying) in streams and wetlands
that periodically dry provides indirect evidence for surface-water connections. Because many
aquatic species can move and disperse overland, aquatic habitats can be highly connected
biologically in the absence of hydrologic connectivity.
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Figure 5-1. Relative positioning of streams, riparian and floodplain waters, and non-floodplain
waters along a gradient of connectivity. Ellipses are used to illustrate the degree of expected overlap
among water-body types based on the range of variation documented in the reviewed literature.
Non-floodplain
waters
Riparian & streams
Floodplain waters
Connection to downstream waters
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This chapter presents the five major conclusions of this report, with a summary of key findings from the
literature synthesized to develop these conclusions. It also discusses the relative abundance of literature
on topics reviewed in this report. Finally, it briefly discusses emerging research that can close some
current data gaps and help further clarify the role of connectivity in maintaining the integrity of
downstream waters.
Citations have been omitted from the text of the conclusions and key findings to improve readability;
please refer to individual chapters for supporting publications and additional information.
6.1 Major Conclusions and Key Findings
Based on our review and synthesis of the literature, we developed five major conclusions, which are
presented in this section with a summary of key findings for each conclusion.
6.1.1 Conclusion 1: Streams
The scientific literature unequivocally demonstrates that streams, individually or cumulatively, exert a
strong influence on the integrity of downstream waters. All tributary streams, including perennial,
intermittent, and ephemeral streams, are physically, chemically, and biologically connected to
downstream rivers via channels and associated alluvial deposits where water and other materials are
concentrated, mixed, transformed, and transported. Streams are the dominant source of water in most
rivers, and the majority of tributaries are perennial, intermittent, or ephemeral headwater streams.
Headwater streams also convey water into local storage compartments such as ponds, shallow aquifers,
or stream banks, and into regional and alluvial aquifers; these local storage compartments are important
sources of water for maintaining baseflow in rivers. In addition to water, streams transport sediment,
wood, organic matter, nutrients, chemical contaminants, and many of the organisms found in rivers. The
literature provides robust evidence that streams are biologically connected to downstream waters by
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the dispersal and migration of aquatic and semiaquatic organisms, including fish, amphibians, plants,
microorganisms, and invertebrates, that use both upstream and downstream habitats during one or
more stages of their life cycles, or provide food resources to downstream communities. In addition to
material transport and biological connectivity, ephemeral, intermittent, and perennial flows influence
fundamental biogeochemical processes by connecting channels and shallow ground water with other
landscape elements. Physical, chemical, and biological connections between streams and downstream
waters interact via integrative processes such as nutrient spiraling, in which stream communities
assimilate and chemically transform large quantities of nitrogen and other nutrients that otherwise
would be transported directly downstream, increasing nutrient loads and associated impairments due
to excess nutrients in downstream waters.
6.1.1.1 Conclusion 1, Key Findings
• Streams are hydrologically connected to downstream waters via channels that convey surface
and subsurface water either year-round (i.e., perennial flow), weekly to seasonally (i.e.,
intermittent flow), or only in direct response to precipitation (i.e., ephemeral flow). Streams are
the dominant source of water in most rivers, and the majority of tributaries are perennial,
intermittent, or ephemeral headwater streams. For example, headwater streams, which are the
smallest channels where streamflows begin, are the cumulative source of approximately 60% of
the total mean annual flow to all northeastern U.S. streams and rivers.
• In addition to downstream transport, headwaters convey water into local storage compartments
such as ponds, shallow aquifers, or stream banks, and into regional and alluvial aquifers. These
local storage compartments are important sources of water for maintaining baseflow in rivers.
Streamflow typically depends on the delayed (i.e., lagged) release of shallow ground water from
local storage, especially during dry periods and in areas with shallow ground-water tables and
pervious subsurfaces. For example, in the southwestern United States, short-term shallow
ground-water storage in alluvial floodplain aquifers, with gradual release into stream channels,
is a major source of annual flow in rivers.
• Infrequent, high-magnitude events are especially important for transmitting materials from
headwater streams in most river networks. For example, headwater streams, including
ephemeral and intermittent streams, shape river channels by accumulating and gradually or
episodically releasing stored materials such as sediment and large woody debris. These
materials help structure stream and river channels by slowing the flow of water through
channels and providing substrate and habitat for aquatic organisms.
• There is strong evidence that headwater streams function as nitrogen sources (via export) and
sinks (via uptake and transformation) for river networks. For example, one study estimated that
rapid nutrient cycling in small streams with no agricultural or urban impacts removed 20-40%
of the nitrogen that otherwise would be delivered to downstream waters. Nutrients are
necessary to support aquatic life, but excess nutrients lead to eutrophication and hypoxia, in
which over-enrichment causes dissolved oxygen concentrations to fall below the level necessary
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to sustain most aquatic animal life in the stream and streambed. Thus, the influence of streams
on nutrient loads can have significant repercussions for hypoxia in downstream waters.
• Headwaters provide habitat that is critical for completion of one or more life-cycle stages of
many aquatic and semiaquatic species capable of moving throughout river networks. Evidence
is strong that headwaters provide habitat for complex life-cycle completion; refuge from
predators, competitors, parasites, or adverse physical conditions in rivers (e.g., temperature or
flow extremes, low dissolved oxygen, high sediment); and reservoirs of genetic- and species-
level diversity. Use of headwater streams as habitat is especially critical for the many species
that migrate between small streams and marine environments during their life cycles (e.g.,
Pacific and Atlantic salmon, American eels, certain lamprey species). The presence of these
species within river networks provides robust evidence of biological connections between
headwaters and larger rivers; because these organisms also transport nutrients and other
materials as they migrate, their presence also provides evidence of biologically mediated
chemical connections. In prairie streams, many fishes swim upstream into tributaries to release
eggs, which develop as they are transported downstream.
• Human alterations affect the frequency, duration, magnitude, timing, and rate of change of
connections between headwater streams, including ephemeral and intermittent streams, and
downstream waters. Human activities and built structures (e.g., channelization, dams, ground-
water withdrawals) can either enhance or fragment longitudinal connections between
headwater streams and downstream waters, while also constraining lateral and vertical
exchanges and tightly controlling the temporal dimension of connectivity. In many cases,
research on human alterations has enhanced our understanding of the headwater stream-
downstream water connections and their consequences. Recognition of these connections and
effects has encouraged the development of more sustainable practices and infrastructure to
reestablish and manage connections, and ultimately to protect and restore the integrity of
downstream waters.
6.1.2 Conclusion 2: Riparian/Floodplain Wetlands and Open Waters
The literature clearly shows that wetlands and open waters in riparian areas and floodplains are
physically, chemically, and biologically integrated with rivers via functions that improve downstream
water quality, including the temporary storage and deposition of channel-forming sediment and woody
debris, temporary storage of local ground water that supports baseflow in rivers, and transformation
and transport of stored organic matter. Riparian/floodplain wetlands and open waters improve water
quality through the assimilation, transformation, or sequestration of pollutants, including excess
nutrients and chemical contaminants such as pesticides and metals, that can degrade downstream water
integrity. In addition to providing effective buffers to protect downstream waters from point source and
nonpoint source pollution, these systems form integral components of river food webs, providing
nursery habitat for breeding fish and amphibians, colonization opportunities for stream invertebrates,
and maturation habitat for stream insects. Lateral expansion and contraction of the river in its
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floodplain result in an exchange of organic matter and organisms, including fish populations that are
adapted to use floodplain habitats for feeding and spawning during high water, that are critical to river
ecosystem function. Riparian/floodplain wetlands and open waters also affect the integrity of
downstream waters by subsequently releasing (desynchronizing) floodwaters and retaining large
volumes of stormwater, sediment, and contaminants in runoff that could otherwise negatively affect the
condition or function of downstream waters.
6.1.2.1 Conclusion 2, Key Findings
• Riparian areas and floodplains connect upland and aquatic environments through both surface
and subsurface hydrologic flowpaths. These areas are therefore uniquely situated in watersheds
to receive and process waters that pass over densely vegetated areas and through subsurface
zones before the waters reach streams and rivers. When pollutants reach a riparian or
floodplain wetland, they can be sequestered in sediments, assimilated into wetland plants and
animals, transformed into less harmful or mobile forms or compounds, or lost to the
atmosphere. Wetland potential for biogeochemical transformations (e.g., denitrification) that
can improve downstream water quality is influenced by local factors, including anoxic
conditions and slow organic matter decomposition, shallow water tables, wetland plant
communities, permeable soils, and complex topography.
• Riparian/floodplain wetlands can reduce flood peaks by storing and desynchronizing
floodwaters. They can also maintain river baseflows by recharging alluvial aquifers. Many
studies have documented the ability of riparian/floodplain wetlands to reduce flood pulses by
storing excess water from streams and rivers. One review of wetland studies reported that
riparian wetlands reduced or delayed floods in 23 of 28 studies. For example, peak discharges
between upstream and downstream gaging stations on the Cache River in Arkansas were
reduced 10-20% primarily due to floodplain water storage.
• Riparian areas and floodplains store large amounts of sediment and organic matter from
upstream and from upland areas. For example, riparian areas have been shown to remove
80-90% of sediments leaving agricultural fields in North Carolina.
• Ecosystem function within a river system is driven in part by biological connectivity that links
diverse biological communities with the river system. Movements of organisms that connect
aquatic habitats and their populations, even across different watersheds, are important for the
survival of individuals, populations, and species, and for the functioning of the river ecosystem.
For example, lateral expansion and contraction of the river in its floodplain result in an exchange
of matter and organisms, including fish populations that are adapted to use floodplain habitats
for feeding and spawning during high water. Wetland and aquatic plants in floodplains can
become important seed sources for the river network, especially if catastrophic flooding scours
vegetation and seed banks in other parts of the channel. Many invertebrates exploit temporary
hydrologic connections between rivers and floodplain wetland habitats, moving into these
wetlands to feed, reproduce, or avoid harsh environmental conditions and then returning to the
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river network. Amphibians and aquatic reptiles commonly use both streams and
riparian/floodplain wetlands to hunt, forage, overwinter, rest, or hide from predators. Birds can
spatially integrate the watershed landscape through biological connectivity.
6.1.3 Conclusion 3: Non-floodplain Wetlands and Open Waters
Wetlands and open waters in non-floodplain landscape settings (hereafter called "non-floodplain
wetlands") provide numerous functions that benefit downstream water integrity. These functions
include storage of floodwater; recharge of ground water that sustains river baseflow; retention and
transformation of nutrients, metals, and pesticides; export of organisms or reproductive propagules to
downstream waters; and habitats needed for stream species. This diverse group of wetlands (e.g., many
prairie potholes, vernal pools, playa lakes) can be connected to downstream waters through surface-
water, shallow subsurface-water, and ground-water flows and through biological and chemical
connections.
In general, connectivity of non-floodplain wetlands occurs along a gradient (Conclusion 4), and can be
described in terms of the frequency, duration, magnitude, timing, and rate of change of water, material,
and biotic fluxes to downstream waters. These descriptors are influenced by climate, geology, and
terrain, which interact with factors such as the magnitudes of the various functions within wetlands
(e.g., amount of water storage or carbon export) and their proximity to downstream waters to
determine where wetlands occur along the connectivity gradient. At one end of this gradient, the
functions of non-floodplain wetlands clearly affect the condition of downstream waters if a visible (e.g.,
channelized) surface-water or a regular shallow subsurface-water connection to the river network is
present. For non-floodplain wetlands lacking a channelized surface or regular shallow subsurface
connection (i.e., those at intermediate points along the gradient of connectivity), generalizations about
their specific effects on downstream waters from the available literature are difficult because
information on both function and connectivity is needed. Although there is ample evidence that non-
floodplain wetlands provide hydrologic, chemical, and biological functions that affect material fluxes, to
date, few scientific studies explicitly addressing connections between non-floodplain wetlands and river
networks have been published in the peer-reviewed literature. Even fewer publications specifically
focus on the frequency, duration, magnitude, timing, or rate of change of these connections. In addition,
although areas that are closer to rivers and streams have a higher probability of being connected than
areas farther away when conditions governing the type and quantity of flows—including soil infiltration
rate, wetland storage capacity, hydraulic gradient, etc.—are similar, information to determine if this
similarity holds is generally not provided in the studies we reviewed. Thus, current science does not
support evaluations of the degree of connectivity for specific groups or classes of wetlands (e.g., prairie
potholes or vernal pools). Evaluations of individual wetlands or groups of wetlands, however, could be
possible through case-by-case analysis.
Some effects of non-floodplain wetlands on downstream waters are due to their isolation, rather than
their connectivity. Wetland sink functions that trap materials and prevent their export to downstream
waters (e.g., sediment and entrained pollutant removal, water storage) result because of the wetland's
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ability to isolate material fluxes. To establish that such functions influence downstream waters, we also
need to know that the wetland intercepts materials that otherwise would reach the downstream water.
The literature we reviewed does provide limited examples of direct effects of wetland isolation on
downstream waters, but not for classes of wetlands (e.g., vernal pools). Nevertheless, the literature we
reviewed enables us to conclude that sink functions of non-floodplain wetlands, which result in part
from their relative isolation, will affect a downstream water when these wetlands are situated between
the downstream water and known point or nonpoint sources of pollution, and thus intersect flowpaths
between the pollutant source and downstream waters.
6.1.3.1 Conclusion 3, Key Findings
• Water storage by wetlands well outside of riparian or floodplain areas can affect streamflow.
Hydrologic models of prairie potholes in the Starkweather Coulee subbasin (North Dakota) that
drains to Devils Lake indicate that increasing the volume of pothole storage across the subbasin
by approximately 60% caused simulated total annual streamflow to decrease 50% during a
series of dry years and 20% during wet years. Similar simulation studies of watersheds that feed
the Red River of the North in North Dakota and Minnesota demonstrated qualitatively
comparable results, suggesting that the ability of potholes to modulate streamflow could be
widespread across eastern portions of the prairie pothole region. This work also indicates that
reducing water storage capacity of wetlands by connecting formerly isolated potholes through
ditching or drainage to the Devils Lake and Red River basins could increase stormflow and
contribute to downstream flooding. In many agricultural areas already crisscrossed by extensive
drainage systems, total streamflow and baseflow are increased by directly connecting potholes
to stream networks. The impacts of changing streamflow are numerous, including altered flow
regime, stream geomorphology, habitat, and ecology. The presence or absence of an effect of
prairie pothole water storage on streamflow depends on many factors, including patterns of
precipitation, topography, and degree of human alteration. For example, in parts of the prairie
pothole region with low precipitation, low stream density, and little human alteration,
hydrologic connectivity between prairie potholes and streams or rivers is likely to be low.
• Non-floodplain wetlands act as sinks and transformers for various pollutants, especially
nutrients, which at excess levels can adversely impact human and ecosystem health and pose a
serious pollution problem in the United States. In one study, sewage waste waters were applied
to forested wetlands in Florida for 4.5 years; more than 95% of the phosphorus, nitrate,
ammonium, and total nitrogen were removed by the wetlands during the study period, and
66-86% of the nitrate removed was attributed to the process of denitrification. In another
study, sizeable phosphorus retention (0.3 to 8.0 mg soluble reactive P nr2 d-1) occurred in
marshes that comprised only 7% of the lower Lake Okeechobee basin area in Florida. A non-
floodplain bog in Massachusetts was reported to sequester nearly 80% of nitrogen inputs from
various sources, including atmospheric deposition, and prairie pothole wetlands in the upper
Midwest were found to remove >80% of the nitrate load via denitrification. A large prairie
marsh was found to remove 86% of nitrate, 78% of ammonium, and 20% of phosphate through
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assimilation and sedimentation, sorption, and other mechanisms. Together, these and other
studies indicate that onsite nutrient removal by non-floodplain wetlands is substantial and
geographically widespread. The effects of this removal on rivers are generally not reported in
the literature.
• Non-floodplain wetlands provide unique and important habitats for many species, both common
and rare. Some of these species require multiple types of waters to complete their full life cycles,
including downstream waters. Abundant or highly mobile species play important roles in
transferring energy and materials between non-floodplain wetlands and downstream waters.
• Biological connections are likely to occur between most non-floodplain wetlands and
downstream waters through either direct or stepping stone movement of amphibians,
invertebrates, reptiles, mammals, and seeds of aquatic plants, including colonization by invasive
species. Many species in those groups that use both stream and wetland habitats are capable of
dispersal distances equal to or greater than distances between many wetlands and river
networks. Migratory birds can be an important vector of long-distance dispersal of plants and
invertebrates between non-floodplain wetlands and the river network, although their influence
has not been quantified. Whether those connections are of sufficient magnitude to impact
downstream waters will either require estimation of the magnitude of material fluxes or
evidence that these movements of organisms are required for the survival and persistence of
biota that contribute to the integrity of downstream waters.
• Spatial proximity is one important determinant of the magnitude, frequency, and duration of
connections between wetlands and streams that will ultimately influence the fluxes of water,
materials, and biota between wetlands and downstream waters. However, proximity alone is
not sufficient to determine connectivity, due to local variation in factors such as slope and
permeability.
• The cumulative influence of many individual wetlands within watersheds can strongly affect the
spatial scale, magnitude, frequency, and duration of hydrologic, biological, and chemical fluxes
or transfers of water and materials to downstream waters. Because of their aggregated
influence, any evaluation of changes to individual wetlands should be considered in the context
of past and predicted changes (e.g., from climate change) to other wetlands within the same
watershed.
• Non-floodplain wetlands can be hydrologically connected directly to river networks through
natural or constructed channels, nonchannelized surface flows, or subsurface flows, the latter of
which can travel long distances to affect downstream waters. A wetland surrounded by uplands
is defined as "geographically isolated." Our review found that, in some cases, wetland types such
as vernal pools and coastal depressional wetlands are collectively—and incorrectly—referred to
as geographically isolated. Technically, the term "geographically isolated" should be applied only
to the particular wetlands within a type or class that are completely surrounded by uplands.
Furthermore, "geographic isolation" should not be confused with functional isolation, because
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geographically isolated wetlands can still have hydrologic, chemical, and biological connections
to downstream waters.
• Non-floodplain wetlands occur along a gradient of hydrologic connectivity-isolation with
respect to river networks, lakes, or marine/estuarine water bodies. This gradient includes, for
example, wetlands that serve as origins for stream channels that have permanent surface-water
connections to the river network; wetlands with outlets to stream channels that discharge to
deep ground-water aquifers; geographically isolated wetlands that have local ground-water or
occasional surface-water connections to downstream waters; and geographically isolated
wetlands that have minimal hydrologic connection to other water bodies (but which could
include surface and subsurface connections to other wetlands). This gradient can exist among
wetlands of the same type or in the same geographic region.
• Caution should be used in interpreting connectivity for wetlands that have been designated as
"geographically isolated" because (1) the term can be applied broadly to a heterogeneous group
of wetlands, which can include wetlands that are not actually geographically isolated; (2)
wetlands with permanent channels could be miscategorized as geographically isolated if the
designation is based on maps or imagery with inadequate spatial resolution, obscured views,
etc.; and (3) wetland complexes could have connections to downstream waters through stream
channels even if individual wetlands within the complex are geographically isolated. For
example, a recent study examined hydrologic connectivity in a complex of wetlands on the Texas
Coastal Plain. The wetlands in this complex have been considered to be a type of geographically
isolated wetland. Collectively, however, they are connected both geographically and
hydrologically to downstream waters in the area: During an almost 4-year study period, nearly
20% of the precipitation that fell on the wetland complex flowed out through an intermittent
stream into downstream waters. Thus, wetland complexes could have connections to
downstream waters through stream channels even when the individual wetland components
are geographically isolated.
6.1.4 Conclusion 4: Degrees and Determinants of Connectivity
Watersheds are integrated at multiple spatial and temporal scales by flows of surface water and ground
water, transport and transformation of physical and chemical materials, and movements of organisms.
Although all parts of a watershed are connected to some degree—by the hydrologic cycle or dispersal of
organisms, for example—the degree and downstream effects of those connections vary spatially and
temporally, and are determined by characteristics of the physical, chemical, and biological environments
and by human activities.
Stream and wetland connections have particularly important consequences for downstream water
integrity. Most of the materials—broadly defined as any physical, chemical, or biological entity—in
rivers, for example, originate from aquatic ecosystems located upstream or elsewhere in the watershed.
Longitudinal flows through ephemeral, intermittent, and perennial stream channels are much more
efficient for transport of water, materials, and organisms than diffuse overland flows, and areas that
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concentrate water provide mechanisms for the storage and transformation, as well as transport, of
materials.
Connectivity of streams and wetlands to downstream waters occurs along a continuum that can be
described in terms of the frequency, duration, magnitude, timing, and rate of change of water, material,
and biotic fluxes to downstream waters. These terms, which we refer to collectively as connectivity
descriptors, characterize the range over which streams and wetlands vary and shift along the
connectivity gradient in response to changes in natural and anthropogenic factors and, when considered
in a watershed context, can be used to predict probable effects of different degrees of connectivity over
time. The evidence unequivocally demonstrates that the stream channels and riparian/floodplain
wetlands or open waters that together form river networks are clearly connected to downstream waters
in ways that profoundly influence downstream water integrity. The connectivity and effects of non-
floodplain wetlands and open waters are more variable and thus more difficult to address solely from
evidence available in peer-reviewed studies.
Variations in the degree of connectivity influence the range of functions provided by streams and
wetlands, and are critical to the integrity and sustainability of downstream waters. Connections with
low values of one or more descriptors (e.g., low-frequency, low-duration streamflows caused by flash
floods) can have important downstream effects when considered in the context of other descriptors
(e.g., large magnitude of water transfer). At the other end of the frequency range, high-frequency, low-
magnitude vertical (surface-subsurface) and lateral flows contribute to aquatic biogeochemical
processes, including nutrient and contaminant transformation and organic matter accumulation. The
timing of an event can alter both connectivity and the magnitude of its downstream effect. For example,
when soils become saturated by previous rainfall events, even low or moderate rainfall can cause
streams or wetlands to overflow, transporting water and materials to downstream waters. Fish that use
nonperennial or perennial headwater stream habitats to spawn or rear young, and invertebrates that
move into seasonally inundated floodplain wetlands prior to emergence, have life cycles that are
synchronized with the timing of flows, temperature thresholds, and food resource availability in those
habitats.
6.1.4.1 Conclusion 4, Key Findings
• The surface-water and ground-water flowpaths (hereafter, hydrologic flowpaths), along which
water and materials are transported and transformed, determine variations in the degree of
physical and chemical connectivity. These flowpaths are controlled primarily by variations in
climate, geology, and terrain within and among watersheds and over time. Climate, geology, and
terrain are reflected locally in factors such as rainfall and snowfall intensity, soil infiltration
rates, and the direction of ground-water flows. These local factors interact with the landscape
positions of streams and wetlands relative to downstream waters, and with functions (such as
the removal or transformation of pollutants) performed by those streams and wetlands to
determine connectivity gradients.
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• Gradients of biological connectivity (i.e., the active or passive movements of organisms through
water or air and over land that connect populations) are determined primarily by species
assemblages, and by features of the landscape (e.g., climate, geology, terrain) that facilitate or
impede the movement of organisms. The temporal and spatial scales at which biological
pathways connect aquatic habitats depend on characteristics of both the landscape and species,
and overland transport or movement can occur across watershed boundaries. Dispersal is
essential for population persistence, maintenance of genetic diversity, and evolution of aquatic
species. Consequently, dispersal strategies reflect aquatic species' responses and adaptations to
biotic and abiotic environments, including spatial and temporal variation in resource availability
and quality. Species' traits and behaviors encompass species-environment relationships over
time, and provide an ecological and evolutionary context for evaluating biological connectivity
in a particular watershed or group of watersheds.
• Pathways for chemical transport and transformation largely follow hydrologic flowpaths, but
sometimes follow biological pathways (e.g., nutrient transport from wetlands to coastal waters
by migrating waterfowl, upstream transport of marine-derived nutrients by spawning of
anadromous fish, uptake and removal of nutrients by emerging stream insects).
• Human activities alter naturally occurring gradients of physical, chemical, and biological
connectivity by modifying the frequency, duration, magnitude, timing, and rate of change of
fluxes, exchanges, and transformations. For example, connectivity can be reduced by dams,
levees, culverts, water withdrawals, and habitat destruction, and can be increased by effluent
discharges, channelization, drainage ditches and tiles, and impervious surfaces.
6.1.5 Conclusion 5: Cumulative Effects
The incremental effects of individual streams and wetlands are cumulative across entire watersheds and
therefore must be evaluated in context with other streams and wetlands. Downstream waters are the
time-integrated result of all waters contributing to them. For example, the amount of water or biomass
contributed by a specific ephemeral stream in a given year might be small, but the aggregate
contribution of that stream over multiple years, or by all ephemeral streams draining that watershed in
a given year or over multiple years, can have substantial consequences on the integrity of the
downstream waters. Similarly, the downstream effect of a single event, such as pollutant discharge into
a single stream or wetland, might be negligible but the cumulative effect of multiple discharges could
degrade the integrity of downstream waters.
In addition, when considering the effect of an individual stream or wetland, all contributions and
functions of that stream or wetland should be evaluated cumulatively. For example, the same stream
transports water, removes excess nutrients, mitigates flooding, and provides refuge for fish when
conditions downstream are unfavorable; if any of these functions is ignored, the overall effect of that
stream would be underestimated.
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6.1.5.1 Conclusion 5, Key Findings
• Structurally and functionally, stream-channel networks and the watersheds they drain are
fundamentally cumulative in how they are formed and maintained. Excess water from
precipitation that is not evaporated, taken up by organisms, or stored in soils and geologic
layers moves downgradient by gravity as overland flow or through channels carrying sediment,
chemical constituents, and organisms. These channels concentrate surface-water flows and are
more efficient than overland (i.e., diffuse) flows in transporting water and materials, and are
reinforced over time by recurrent flows.
• Connectivity between streams and rivers provides opportunities for materials, including
nutrients and chemical contaminants, to be transformed chemically as they are transported
downstream. Although highly efficient at the transport of water and other physical materials,
streams are dynamic ecosystems with permeable beds and banks that interact with other
ecosystems above and below the surface. The exchange of materials between surface and
subsurface areas involves a series of complex physical, chemical, and biological alterations that
occur as materials move through different parts of the river system. The amount and quality of
such materials that eventually reach a river are determined by the aggregate effect of these
sequential alterations that begin at the source waters, which can be at some distance from the
river. The opportunity for transformation of material (e.g., biological uptake, assimilation, or
beneficial transformation) in intervening stream reaches increases with distance to the river.
Nutrient spiraling, the process by which nutrients entering headwater streams are transformed
by various aquatic organisms and chemical reactions as they are transported downstream, is
one example of an instream alteration that exhibits significant beneficial effects on downstream
waters. Nutrients (in their inorganic form) that enter a headwater stream (e.g., via overland
flow) are first removed from the water column by streambed algal and microbial populations.
Fish or insects feeding on algae and microbes take up some of those nutrients, which are
subsequently released back into the stream via excretion and decomposition (i.e., in their
organic form), and the cycle is repeated. In each phase of the cycling process—from dissolved
inorganic nutrients in the water column, through microbial uptake, subsequent transformations
through the food web, and back to dissolved nutrients in the water column—nutrients are
subject to downstream transport. Stream and wetland capacities for nutrient cycling have
important implications for the form and concentration of nutrients exported to downstream
waters.
• Cumulative effects across a watershed must be considered when quantifying the frequency,
duration, and magnitude of connectivity, to evaluate the downstream effects of streams and
wetlands. For example, although the probability of a large-magnitude transfer of organisms
from any given headwater stream in a given year might be low (i.e., a low-frequency connection
when each stream is considered individually), headwater streams are the most abundant type of
stream in most watersheds. Thus, the overall probability of a large-magnitude transfer of
organisms is higher when considered for all headwater streams in a watershed—that is, a high-
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frequency connection is present when headwaters are considered cumulatively at the
watershed scale, compared with probabilities of transport for streams individually. Similarly, a
single pollutant discharge might be negligible but the cumulative effect of multiple discharges
could degrade the integrity of downstream waters. Riparian open waters (e.g., oxbow lakes),
wetlands, and vegetated areas cumulatively can retain up to 90% of eroded clays, silts, and
sands that otherwise would enter stream channels. The larger amounts of snowmelt and
precipitation cumulatively held by many wetlands can reduce the potential for flooding at
downstream locations. For example, wetlands in the prairie pothole region cumulatively stored
about 11-20% of the precipitation in one watershed.
• The combination of diverse habitat types and abundant food resources cumulatively makes
floodplains important foraging, hunting, and breeding sites for fish, aquatic life stages of
amphibians, and aquatic invertebrates. The scale of these cumulative effects can be extensive;
for example, coastal ibises travel up to 40 km to obtain food from freshwater floodplain
wetlands for nesting chicks, which cannot tolerate salt levels in local food resources until they
fledge.
6.2 Strength of Evidence for Conclusions and Data Gaps in
the Available Literature
This report synthesizes a large body of scientific evidence to address the questions in Table 1-1 of this
report. The major conclusions (Section 6.1) reflect the strength of evidence currently available in the
peer-reviewed scientific literature for assessing the connectivity and downstream effects of water
bodies identified in Table 1-1.
The conclusions of this report were corroborated by two independent peer reviews by scientists
identified in the front matter of this report.
The term connectivity is defined in this report as the degree to which components of a watershed are
joined and interact by transport mechanisms that function across multiple spatial and temporal scales
(Sections 1.2.2 and 2.3.2.1). Our review found strong evidence supporting the central roles of the
physical, chemical, and biological connectivity of streams, wetlands, and open waters—encompassing
varying degrees of both connection and isolation—in maintaining the structure and function of
downstream waters, including rivers, lakes, estuaries, and oceans. Our review also found strong
evidence demonstrating the various mechanisms by which material and biological linkages from
streams, wetlands, and open waters affect downstream waters, classified here into five functional
categories (source, sink, refuge, lag, and transformation), modify the timing of transport and the
quantity and quality of resources available to downstream ecosystems and communities. Thus, the
currently available literature provided a large body of evidence for assessing the connections and
functions by which streams and wetlands produce the range of observed effects on the integrity of
downstream waters.
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The body of literature on functions provided by streams and riparian/floodplain wetlands was abundant
in all five categories (Table 6-1). The body of literature on functions of non-floodplain wetlands was
abundant in two categories (sink and transformation) and moderate in the other three categories
(source, refuge, and lag; Table 6-1). The evidence unequivocally demonstrates that the stream channels
and wetlands or open waters that together form river networks are clearly connected to downstream
waters in ways that profoundly influence downstream water integrity. The body of literature
documenting connectivity and downstream effects was most abundant for perennial and intermittent
streams, and for riparian/floodplain wetlands (Table 6-2). Although less abundant, the available
evidence for connectivity and downstream effects of ephemeral streams was strong and compelling,
particularly in context with the large body of evidence supporting the physical connectivity and
cumulative effects of channelized flows that form and maintain stream networks.
As stated in Conclusion 3 (Section 6.1.3), the connectivity and effects of wetlands and open waters that
are not structurally linked to other waters by stream channels and their lateral extensions into riparian
areas and floodplains are more difficult to address solely from evidence available in peer-reviewed
studies. One limitation was the relatively small number of published, peer-reviewed studies examining
the relationships of non-floodplain wetlands to downstream waters (Table 6-2). The literature on non-
floodplain wetlands that is available shows that these systems have important hydrologic, water-quality,
and habitat functions that can affect downstream waters where connections to them exist; the literature
also provides limited examples of direct effects of non-floodplain wetland isolation on downstream
water integrity. Currently available peer-reviewed literature, however, does not identify which types of
non-floodplain wetlands have or lack the types of connections needed to convey the effects on
downstream waters of functions, materials, or biota provided by those wetlands. These limitations of the
literature, considered in context with comments from the Science Advisory Board on an external review
draft of this report (U.S. EPA, 2014), are reflected in the lower strength of evidence expressed in the
conclusions (Section 6.1.3).
Additional information from other sources not included in this report (e.g., field assessments, analysis of
existing or new data, reports from local resource agencies) could be used in case-by-case analysis of
non-floodplain wetlands. Importantly, information from emerging research into the connectivity of non-
floodplain wetlands, including studies of the types identified in Section 4.5.2 of this report, could close
some of the current data gaps in the near future. Recent scientific advances in the fields of mapping (e.g.,
Heine etal., 2004; Tiner, 2011; Langetal., 2012), assessment (e.g., McGlynn and McDonnell, 2003;
Gergel, 2005; McGuire etal., 2005; Ver Hoef etal., 2006; Leibowitz etal., 2008; Moreno-Mateos etal.,
2008; Lane and D'Amico, 2010; Ver Hoef and Peterson, 2010; Shook and Pomeroy, 2011; Powers etal.,
2012; McDonough et al., 2015), modeling (e.g., Golden et al., 2013; Mclaughlin et al., 2014), and
landscape classification (e.g., Wigington et al., 2013) indicate that increasing availability of high-
resolution data sets, promising new technologies for watershed-scale analyses, and methods for
classifying landscape units by hydrologic behavior can facilitate and improve the accuracy of
connectivity assessments. Emerging research that expands our ability to detect and monitor ecologically
relevant connections at appropriate scales, metrics to accurately measure effects on downstream
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integrity, and management practices that apply what we already know about ecosystem function, will
contribute to our ability to identify waters of national importance and maintain the long-term
sustainability and resiliency of valued water resources.
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1-3
§ ffl
body of literature documented in the report, in the evidence for source, sink, refuge, lag, and transformation functions of streams and wetlands
and their associated effects on downstream waters. A small dot (•) indicates relatively lower confidence, a medium dot (•) indicates relatively
intermediate confidence, and a large dot (^)) indicates a relatively high level of confidence. The dot size does not necessarily correspond with
many references. The dot size also does not correspond with the level of confidence in particular conclusions.
a
Type of water body
Streams
Riparian/floodplain wetlands
Non-floodplain wetlands
Function
Source
O
o
o
Sink (Storage)
O
o
o
Refuge
O
o
o
Lag
O
o
o
Transformation
O
o
o
Uncertainty discussion
(Section)
3.6
4.5.1
4.5.2
to
o
-------
1-3
a
" 3
<2. O
(/) o
Table 6-2. Relative abundance of literature by review topic area. The table shows the relative size of the body of literature documented in the
report that addresses the physical, chemical, or biological connectivity to and effects on downstream waters. A small dot (•) indicates a relatively
smaller body of literature, a medium dot (•) indicates a relatively intermediate body of literature, and a large dot (^)) indicates a relatively
large body of literature. The dot size does not necessarily correspond with the number of associated citations in this report because some
level of confidence in particular conclusions.
Topic
Streams
Riparian/
Floodplain
Wetlands
Non-
floodplain
wetlands
Question
What are the physical, chemical,
and biological connections to and
effects of ephemeral, intermittent,
and perennial streams on
downstream waters?
What are the physical, chemical,
and biological connections to and
effects of riparian orfloodplain
wetlands and open waters (e.g.,
riverine wetlands, oxbow lakes) on
downstream waters?
What are the physical, chemical,
and biological connections to and
effects of wetlands and open
waters in non-floodplain settings
(e.g., most prairie potholes, vernal
pools) on downstream waters?
ephemeral
intermittent
perennial
Biological
Connection
O
o
o
o
o
Effect
o
O
o
o
o
Chemical
Connection
O
o
o
o
o
Effect
O
O
O
o
o
Physical
Connection
O
o
o
o
o
Effect
O
O
O
o
o
to
o
-------
Aastrup, M., J. Johnson, E. Bringmark, L. Bringmark, and A. Iverfeldt. 1991. Occurrence and transport of
mercury within a small catchment area. Water, Air, and Soil Pollution 56:155-167.
Abbe, T. E., and D. R. Montgomery. 1996. Large woody debris jams, channel hydraulics and habitat
formation in large rivers. Regulated Rivers: Research & Management 12:201-221.
Abbott, M. D., A. Lini, and P. R. Bierman. 2000.180, D, and 3H measurements constrain groundwater
recharge patterns in an upland fractured bedrock aquifer, Vermont, USA. Journal of Hydrology
228:101-112.
Abdelnour, A., R. B. McKane, M. Stieglitz, F. Pan, and Y. Cheng. 2013. Effects of harvest on carbon and
nitrogen dynamics in a Pacific Northwest forest catchment. Water Resources Research 49:1292-
1313.
Acreman, M., and J. Holden. 2013. How wetlands affect floods. Wetlands 33:773-786.
Acuna, V., I. Munoz, A. Giorgi, M. Omella, F. Sabater, and S. Sabater. 2005. Drought and postdrought
recovery cycles in an intermittent Mediterranean stream: Structural and functional aspects. Journal
of the North American Benthological Society 24:919-933.
Adair, E. C., D. Binkley, and D. C. Andersen. 2004. Patterns of nitrogen accumulation and cycling in
riparian floodplain ecosystems along the Green and Yampa rivers. Oecologia 139:108-116.
Adair, S. E., J. L. Moore, and W. H. Kiel. 1996. Wintering diving duck use of coastal ponds: An analysis of
alternative hypotheses. The Journal of Wildlife Management 60:83-93.
Adams, S. B., D. A. Schmetterling, and M. K. Young. 2005. Instream movements by boreal toads (Bufo
boreas boreas}. Herpetological Review 36:27-33.
Ademollo, N., S. Capri, L. Patrolecco, A. Puddu, S. Polesello, M. Rusconi, S. Valsecchi, and J. Froebrich.
2011. Fate and monitoring of hazardous substances in temporary rivers. Trends in Analytical
Chemistry 30:1222-1232.
Stream and Wetland Connectivity:
A Review and Synthesis
7-1
January 2015
-------
Agren, A., I. Buffam, M. Jansson, and H. Laudon. 2007. Importance of seasonality and small streams for
the landscape regulation of dissolved organic carbon export. Journal of Geophysical Research:
Biogeosciences 112:G03003.
Alexander, L. C., D. J. Hawthorne, M. A. Palmer, and W. 0. Lamp. 2011. Loss of genetic diversity in the
North American mayfly Ephemerella invaria associated with deforestation of headwater streams.
Freshwater Biology 56:1456-1467.
Alexander, R., J. Bb'hlke, E. Boyer, M. David, J. Harvey, P. Mulholland, S. Seitzinger, C. Tobias, C. Tonitto,
and W. Wollheim. 2009. Dynamic modeling of nitrogen losses in river networks unravels the
coupled effects of hydrological and biogeochemical processes. Biogeochemistry 93:91-116.
Alexander, R. B., E. W. Boyer, R. A. Smith, G. E. Schwarz, and R. B. Moore. 2007. The role of headwater
streams in downstream water quality. Journal of the American Water Resources Association 43:41-
59.
Alexander, R. B., R. A. Smith, and G. E. Schwarz. 2000. Effect of stream channel size on the delivery of
nitrogen to the Gulf of Mexico. Nature 403:758-761.
Alford, J. D., and M. R. Walker. 2013. Managing the flood pulse for optimal fisheries production in the
Atchafalaya River basin, Louisiana (USA). River Research and Applications 29:279-296.
Ali, G. A., and A. G. Roy. 2010. Shopping for hydrologically representative connectivity metrics in a humid
temperate forested catchment. Water Resources Research 46:W12544.
Allan, J. D. 1995. Stream ecology - Structure and function of running waters. Chapman & Hall, New York,
NY.
Allen, M. R. 2007. Measuring and modeling dispersal of adult zooplankton. Oecologia 153:135-143.
Amezaga, J. M., L. Santamaria, and A. J. Green. 2002. Biotic wetland connectivity—supporting a new
approach for wetland policy. Acta Oecologica 23:213-222.
Amoros, C., and G. Bornette. 2002. Connectivity and biocomplexity in waterbodies of riverine
floodplains. Freshwater Biology 47:761-776.
Anderson, C. D., B. K. Epperson, M. J. Fortin, R. Holderegger, P. M. A. James, M. S. Rosenberg, K. T.
Scribner, and S. Spear. 2010. Considering spatial and temporal scale in landscape-genetic studies of
gene flow. Molecular Ecology 19:3565-3575.
Anderson, N. H., and J. R. Sedell. 1979. Detritus processing by macroinvertebrates in stream ecosystems.
Annual Review of Entomology 24:351-377.
Angeler, D. G., M. Alvarez-Cobelas, C. Rojo, and S. Sanchez-Carrillo. 2010. Phytoplankton community
similarity in a semiarid floodplain under contrasting hydrological connectivity regimes. Ecological
Research 25:513-520.
Anteau, M. J., M. H. Sherfy, and A. A. Bishop. 2011. Location and agricultural practices influence spring
use of harvested cornfields by cranes and geese in Nebraska. The Journal of Wildlife Management
75:1004-1011.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis
-------
Arce, M. I., M. d. M. Sanchez-Montoya, M. R. Vidal-Abarca, M. L. Suarez, and R. Gomez. 2014. Implications
of flow intermittency on sediment nitrogen availability and processing rates in a Mediterranean
headwater stream. Aquatic Sciences 76:173-186.
Armsworth, P. R., and J. E. Roughgarden. 2005. The impact of directed versus random movement on
population dynamics and biodiversity patterns. The American Naturalist 165:449-465.
Arndt, J. L., and J. L. Richardson. 1989. Geochemistry of hydric soil salinity in a recharge-throughflow-
discharge prairie-pothole wetland system. Soil Science Society of America Journal 53:848-855.
Arrigoni, A. S., G. C. Poole, L. A. K. Mertes, S. J. O'Daniel, W. W. Woessner, and S. A. Thomas. 2008.
Buffered, lagged, or cooled? Disentangling hyporheic influences on temperature cycles in stream
channels. Water Resources Research 44:W09418.
Arscott, D. B., K. Tockner, and J. V. Ward. 2001. Thermal heterogeneity along a braided floodplain river
(Tagliamento River, northeastern Italy). Canadian Journal of Fisheries and Aquatic Sciences
58:2359-2373.
Ash worth, A. C. 1999. Climate change in North Dakota since the last glaciation—review of the
paleontological record. Proceedings of the North Dakota Academy of Science 53:171-176.
Atkinson, E. 1995. Methods for assessing sediment delivery in river systems. Hydrological Sciences
Journal 40:273-280.
Attum, 0., Y. M. Lee, J. H. Roe, and B. A. Kingsbury. 2007. Upland-wetland linkages: Relationship of
upland and wetland characteristics with watersnake abundance. Journal of Zoology 271:134-139.
Augspurger, C., G. Gleixner, C. Kramer, and K. Kusel. 2008. Tracking carbon flow in a 2-week-old and 6-
week-old stream biofilm food web. Limnology and Oceanography 53:642-650.
Austin, J., and A. Richert 2005. Patterns of habitat use by whooping cranes during migration: Summary
from 1977-1999 site evaluation data. Proceedings North American Crane Workshop 9:79-107.
Axtmann, E. V., and S. N. Luoma. 1991. Large-scale distribution of metal contamination in the fine-
grained sediments of the Clark Fork River, Montana, USA. Applied Geochemistry 6:75-88.
Babbitt, K. J., M. J. Baber, and T. L. Tarr. 2003. Patterns of larval amphibian distribution along a wetland
hydroperiod gradient. Canadian Journal of Zoology 81:1539-1552.
Babbitt, K. J., and G. W. Tanner. 2000. Use of temporary wetlands by anurans in a hydrologically modified
landscape. Wetlands 20:313-322.
Baber, M. J., D. L. Childers, K. J. Babbitt, and D. H. Anderson. 2002. Controls on fish distribution and
abundance in temporary wetlands. Canadian Journal of Fisheries and Aquatic Sciences 59:1441-
1450.
Bailey Boomer, K. M., and B. L. Bedford. 2008. Influence of nested groundwater systems on reduction-
oxidation and alkalinity gradients with implications for plant nutrient availability in four New York
fens. Journal of Hydrology 351:107-125.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis
-------
Baillie, M., J. F. Hogan, B. Ekwurzel, A. K. Wahi, and C. J. Eastoe. 2007. Quantifying water sources to a
semiarid riparian ecosystem, San Pedro River, Arizona. Journal of Geophysical Research:
Biogeosciences 112:G03S02.
Baldwin, D. S., and A. M. Mitchell. 2000. The effects of drying and re-flooding on the sediment and soil
nutrient dynamics of lowland river-floodplain systems: A synthesis. Regulated Rivers: Research &
Management 16:457-467.
Ballard, B., J. Dale James, R. Bingham, M. Petrie, and B. Wilson. 2010. Coastal pond use by redheads
wintering in the Laguna Madre, Texas. Wetlands 30:669-674.
Banks, E. W., C. T. Simmons, A. J. Love, and P. Shand. 2011. Assessing spatial and temporal connectivity
between surface water and groundwater in a regional catchment: Implications for regional scale
water quantity and quality. Journal of Hydrology 404:30-49.
Barling, R. D., I. D. Moore, and R. B. Grayson. 1994. A quasi-dynamic wetness index for characterizing the
spatial distribution of zones of surface saturation and soil water content. Water Resources Research
30:1029-1044.
Barlocher, F., R. J. Mackay, and G. B. Wiggins. 1978. Detritus processing in a temporary vernal pool in
southern Ontario. Archivfiir Hydrobiologie 81:269-295.
Baron, J. S., N. L. Poff, P. L. Angermeier, C. N. Dahm, P. H. Gleick, N. G. Hairston, R. B. Jackson, C. A.
Johnston, B. D. Richter, and A. D. Steinman. 2002. Meeting ecological and societal needs for
freshwater. Ecological Applications 12:1247-1260.
Barrat-Segretain, M. H. 1996. Strategies of reproduction, dispersion, and competition in river plants: A
review. Vegetatio 123:13-37.
Barton, D. R., W. D. Taylor, and R. M. Biette. 1985. Dimensions of riparian buffer strips required to
maintain trout habitat in southern Ontario streams. North American Fisheries Management 5:364-
378.
Battin, T. J., L. A. Kaplan, S. Findlay, C. S. Hopkinson, E. Marti, A. I. Packman, J. D. Newbold, and F. Sabater.
2009. Biophysical controls on organic carbon fluxes in fluvial networks. Nature Geoscience 1:95-
100.
Battin, T. J., and D. Sengschmitt. 1999. Linking sediment biofilms, hydrodynamics, and river bed
clogging: Evidence from a large river. Microbial Ecology 37:185-196.
Baxter, C. V., K. D. Fausch, M. Murakami, and P. L. Chapman. 2004. Fish invasion restructures stream and
forest food webs by interrupting reciprocal prey subsidies. Ecology 85:2656-2663.
Baxter, C. V., and F. R. Hauer. 2000. Geomorphology, hyporheic exchange, and selection of spawning
habitat by bull trout [Salvelinus confluentus). Canadian Journal of Fisheries and Aquatic Sciences
57:1470-1481.
Bayless, M. A., M. G. McManus, and J. F. Fairchild. 2003. Geomorphic, water quality and fish community
patterns associated with the distribution of Notropis topeka in a Central Missouri watershed.
American Midland Naturalist 150:58-72.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis
-------
Bedford, B. L, and K. S. Godwin. 2003. Fens of the United States: Distribution, characteristics, and
scientific connection versus legal isolation. Wetlands 23:608-629.
Beeson, C. E., and P. F. Doyle. 1995. Comparison of bank erosion at vegetated and non-vegetated channel
bends. Journal of the American Water Resources Association 31:983-990.
Bencala, K. E. 1993. A perspective on stream-catchment connections. Journal of the North American
Benthological Society 12:44-47.
Bencala, K. E. 2005. Hyporheic exchange flows. Pages 1733-1740 in Encyclopedia of hydrological
sciences. M. G. Anderson, editor. John Wiley and Son, Ltd., New York, NY.
Bencala, K. E. 2011. Stream-groundwater interactions. Pages 537-546 in Treatise on water science. P.
Wilderer, editor. Academic Press, Oxford, UK.
Bencala, K. E., M. N. Gooseff, and B. A. Kimball. 2011. Rethinking hyporheic flow and transient storage to
advance understanding of stream-catchment connections. Water Resources Research 47:WOOH03.
Benda, L. 2008. Confluence environments at the scale of river networks. Pages 271-300 in River
confluences, tributaries and the fluvial network. S. P. Rice, A. G. Roy, and B. L. Rhoads, editors. John
Wiley & Sons, Chichester, UK.
Benda, L., and T. Dunne. 1987. Sediment routing by debris flow. Pages 213-223 in Erosion and
sedimentation in the Pacific Rim. IAHS Publication 165. R. L. Beschta, T. Blinn, G. E. Grant, F. J.
Swanson, and G. G. Ice, editors. International Association for Hydrological Science, Wallingfor, UK.
Benda, L., M. A. Hassan, M. Church, and C. L. May. 2005. Geomorphology of steepland headwaters: The
transition from hillslopes to channels. Journal of the American Water Resources Association 41:835-
851.
Benda, L., N. L. Poff, D. Miller, T. Dunne, G. Reeves, G. Pess, and M. Pollock. 2004. The network dynamics
hypothesis: How channel networks structure riverine habitats. BioScience 54:413-427.
Benda, L. E., and T. W. Cundy. 1990. Predicting deposition of debris flows in mountain channels.
Canadian Geotechnical Journal 27:409-417.
Benda, L. E., D. J. Miller, T. Dunne, G. H. Reeves, and J. K. Agee. 1998. Dynamic landscape systems. Pages
261-288 in River ecology and management: Lessons from the Pacific coastal ecoregion. R. J. Naiman
and R. E. Bilby, editors. Springer-Verlag, New York, NY.
Benfield, E. F. 1997. Comparison of litterfall inputto streams. Journal of the North American
Benthological Society 16:104-108.
Benke, A. C., I. Chaubey, G. M. Ward, and E. L. Dunn. 2000. Flood pulse dynamics of an unregulated river
floodplain in the southeastern US coastal plain. Ecology 81:2730-2741.
Bennett, M. G., K. A. Fritz, A. Hayden-Lesmeister, J. P. Kozak, and A. Nickolotsky. 2015. An estimate of
basin-wide denitrification based on floodplain inundation in the Atchafalaya River basin, Louisiana.
River Research and Applications: doi: 10.1002/rra.2854.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis
-------
Benoit, P., E. Barriuso, P. Vidon, and B. Real. 1999. Isoproturon sorption and degradation in a soil from
grassed buffer strip. Journal of Environmental Quality 28:121-129.
Benstead, J. P., and D. S. Leigh. 2012. An expanded role for river networks. Nature Geoscience 5:678-679.
Bergsten, A., and A. Zetterberg. 2013. To model the landscape as a network: A practitioner's perspective.
Landscape and Urban Planning 119:35-43.
Bernhardt, E. S., R. 0. Hall, and G. E. Likens. 2002. Whole-system estimates of nitrification and nitrate
uptake in streams of the Hubbard Brook Experimental Forest. Ecosystems 5:419-430.
Beschta, R. L, R. E. Bilby, G. W. Brown, L. B. Holtby, and T. D. Hofstra. 1987. Stream temperature and
aquatic habitat: Fisheries and forestry interactions. Pages 191-232 in Streamside management:
Forestry and fishery interactions. E. 0. Salo and T. W. Cundy, editors. College of Forest Resources,
University of Washington, Seattle, WA, USA.
Beschta, R. L., and W. J. Ripple. 2012. The role of large predators in maintaining riparian plant
communities and river morphology. Geomorphology 157:88-98.
Best, J. L. 1988. Sediment transport and bed morphology at river channel confluences. Sedimentology
35:481-498.
Bestgen, A. C., I. Chaubey, G. M. Ward, and L. Dunn. 2000. Flood pulse dynamics of an unregulated river
floodplain in the southeastern U.S. Coastal Plain. Ecology 81:2730-2741.
Beven, K., and P. Germann. 1982. Macropores and water flow in soils. Water Resources Research
18:1311-1325.
Beven, K. J., and M. J. Kirkby. 1979. A physically based, variable contributing area model of basin
hydrology. Hydrological Sciences Bulletin 24:43-69.
Bhadha, J., J. Jawitz, and J.-H. Min. 2011. Phosphorus mass balance and internal load in an impacted
subtropical isolated wetland. Water, Air, & Soil Pollution 218:619-632.
Bigelow, P. E., L. E. Benda, D. J. Miller, and K. M. Burnett. 2007. On debris flows, river networks, and the
spatial structure of channel morphology. Forest Science 53:220-238.
Bilby, R. A., and G. E. Likens. 1980. Importance of organic debris dams in the structure and function of
stream ecosystems. Ecology 61:1107-1113.
Bilby, R. A., and J. W. Ward. 1989. Changes in characteristics and function of woody debris with
increasing size of streams in western Washington. Transactions of the American Fisheries Society
118:368-378.
Bilby, R. E., and P. A. Bisson. 1998. Function and distribution of large woody debris. Pages 324-346 in
River ecology and management: Lessons from the Pacific coastal ecoregion. R. J. Naiman and R. E.
Bilby, editors. Springer-Verlag, New York, NY.
Bischel, H. N., J. E. Lawrence, B. J. Halaburka, M. H. Plumlee, A. S. Bawazir, J. P. King, J. E. McCray, V. H.
Resh, and R. G. Luthy. 2013. Renewing urban streams with recycled water for streamflow
Stream and Wetland Connectivity: January 2015
A Review and Synthesis
-------
augmentation: Hydrologic, water quality, and ecosystem services management. Environmental
Engineering Science 30:455-479.
Blanchong, J. A., M. D. Samuel, and G. Mack. 2006. Multi-species patterns of avian cholera mortality in
Nebraska's Rainwater Basin. Journal of Wildlife Diseases 42:81-91.
Blann, K., J. F. Nerbonne, and B. Vondracek. 2002. Relationship of riparian buffer type to water
temperature in the driftless area ecoregion of Minnesota. North American Journal of Fisheries
Management 22:441-451.
Blann, K. L, J. L. Anderson, G. R. Sands, and B. Vondracek. 2009. Effects of agricultural drainage on
aquatic ecosystems: A review. Critical Reviews in Environmental Science and Technology 39:909-
1001.
Bodamer, B. L., and J. M. Bossenbroek. 2008. Wetlands as barriers: Effects of vegetated waterways on
downstream dispersal of zebra mussels. Freshwater Biology 53:2051-2060.
Boelter, D., and E. S. Verry. 1977. Peatland and water in the northern lake states. General Technical
Report NC-31, U.S. Department of Agriculture, U.S. Forest Service, North Central Forest Experiment
Station, St. Paul, MN.
Bogan, M. T., and K. S. Boersma. 2012. Aerial dispersal of aquatic invertebrates along and away from
arid-land streams. Freshwater Science 31:1131-1144.
Bogan, M. T., and D. A. Lytle. 2007. Seasonal flow variation allows 'time-sharing' by disparate aquatic
insect communities in montane desert streams. Freshwater Biology 52:290-304.
Bohlke, J., R. Antweiler, J. Harvey, A. Laursen, L. Smith, R. Smith, and M. Voytek. 2009. Multi-scale
measurements and modeling of denitrification in streams with varying flow and nitrate
concentration in the upper Mississippi River basin, USA. Biogeochemistry 93:117-141.
Bohonak, A. J., and D. G. Jenkins. 2003. Ecological and evolutionary significance of dispersal by
freshwater invertebrates. Ecology Letters 6:783-796.
Boltz, J. M., and R. R. J. Stauffer. 1989. Fish assemblages of Pennsylvania wetlands. Pages 158-170 in
Wetland ecology and conservation: Emphasis in Pennsylvania. S. K. Majumdar, editor. The
Pennslyvania Academy of Sciences, Lafayette College, Easton, PA.
Boon, P. 1.1991. Bacterial assemblages in rivers and billabongs of southeastern Australia. Microbial
Ecology 22:27-52.
Boon, P. I. 2006. Biogeochemistry and bacterial ecology of hydrologically dynamic wetlands. Pages 115-
176 in Ecology of freshwater and estuarine wetlands. D. P. Batzer and R. R. Sharitz, editors.
University of California Press, Berkeley, CA.
Booth, D. B. 1990. Stream-channel incision following drainage-basin urbanization. Journal of the
American Water Resources Association 26:407-417.
Booth, D. B., D. Hartley, and R. Jackson. 2002. Forest cover, impervious-surface area, and the mitigation
of stormwater impacts. Journal of the American Water Resources Association 38:835-845.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis
-------
Bormann, F. H., and G. E. Likens. 1979. Catastrophic disturbance and the steady-state in northern
hardwood forests. American Scientist 67:660-669.
Born, S. M., S. A. Smith, and D. A. Stephenson. 1979. Hydrogeology of glacial-terrain lakes, with
management and planning applications. Journal of Hydrology 43:7-43.
Borselli, L, P. Cassi, and D. Torri. 2008. Prolegomena to sediment and flow connectivity in the landscape:
A CIS and field numerical assessment. CATENA 75:268-277.
Boschilia, S. M., E. F. Oliveira, and S. M. Thomaz. 2008. Do aquatic macrophytes co-occur randomly? An
analysis of null models in a tropical floodplain. Oecologia 156:203-214.
Boss, S. M., and J. S. Richardson. 2002. Effects of food and cover on the growth, survival, and movement
of cutthroat trout (Oncorhynchus clarki) in coastal streams. Canadian Journal of Fisheries and
Aquatic Sciences 59:1044-1053.
Boto, K. G., and W. H. Patrick. 1979. Role of wetlands in the removal of suspended sediments. Pages 479-
489 in Wetland functions and values: The state of our understanding. Proceedings of National
Symposium on Wetlands. P. E. Greeson, J. R. Clark, and J. E. Clark, editors. American Water Resources
Association, Minneapolis, MN.
Boudell, J. A., and J. C. Stromberg. 2008. Flood pulsing and metacommunity dynamics in a desert riparian
ecosystem. Journal of Vegetation Science 19:373-380.
Boulton, A. 2000. The subsurface macrofauna. Pages 337-361 in Streams and Ground Waters. J. B. Jones
and P. J. Mulholland, editors. Academic Press, San Diego, CA.
Boulton, A. J., S. Findlay, P. Marmonier, E. H. Stanley, and H. M. Valett. 1998. The functional significance of
the hyporheric zone in streams and rivers. Annual Review of Ecology and Systematics 29:59-81.
Bracken, L. J., and J. Croke. 2007. The concept of hydrological connectivity and its contribution to
understanding runoff-dominated geomorphic systems. Hydrological Processes 21:1749-1763.
Bracken, L. J., J. Wainwright, G. A. Ali, D. Tetzlaff, M. W. Smith, S. M. Reaney, and A. G. Roy. 2013. Concepts
of hydrological connectivity: Research approaches, pathways and future agendas. Earth-Science
Reviews 119:17-34.
Bradford, M. J., J. A. Grout, and S. Moodie. 2001. Ecology of juvenile Chinook salmon in a small non-natal
stream of the Yukon River drainage and the role of ice conditions on their distribution and survival.
Canadian Journal of Zoology 79:2043-2054.
Brahana, J. V., and E. F. Hollyday. 1988. Dry stream reaches in carbonate terranes: Surface indicators of
ground-water reservoirs. Water Resources Bulletin 24:577-580.
Bramblett, R. G., M. D. Bryant, B. E. Wright, and R. G. White. 2002. Seasonal use of small tributary and
main-stem habitats by juvenile steelhead, coho salmon, and Dolly Varden in a southeastern Alaska
drainage basin. Transactions of the American Fisheries Society 131:498-506.
Branfireun, B. 2004. Does microtopography influence subsurface pore-water chemistry? Implications for
the study of methylmercury in peatlands. Wetlands 24:207-211.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis
-------
Branfireun, B., N. Roulet, C. Kelly, and J. Rudd. 1999. In situ sulphate stimulation of mercury methylation
in a boreal peatland: Toward a link between acid rain and methylmercury contamination in remote
environments. Global Biogeochemical Cycles 13:743-750.
Branfireun, B. A., A. Heyes, and N. T. Roulet. 1996. The hydrology and methylmercury dynamics of a
Precambrian shield headwater peatland. Water Resources Research 32:1785-1794.
Branfireun, B. A., and N. T. Roulet. 1998. The baseflow and storm flow hydrology of a Precambrian shield
headwater peatland. Hydrological Processes 12:57-72.
Branfireun, B. A., and N. T. Roulet. 2002. The boreal catchment hydrological cascade: Controls on the fate
and transport of methylmercury. Hydrology and Earth System Sciences 6:785-794.
Brinson, M. M. 1993. A hydrogeomorphic classification for wetlands. Technical Report WRP-DE-4, U.S.
Army Corps of Engineers, Waterways Experiment Station, Wetlands Research Program, Vicksburg,
MS.
Brittain, J. E., and T. J. Eikeland. 1988. Invertebrate drift: A review. Hydrobiologia 166:77-93.
Brooks, B. W., T. M. Riley, and R. D. Taylor. 2006. Water quality of effluent-dominated ecosystems:
Ecotoxicological, hydrological, and management considerations. Hydrobiologia 556:365-379.
Brooks, P. D., P. A. Haas, and A. K. Huth. 2007. Seasonal variability in the concentration and flux of
organic matter and inorganic nitrogen in a semiarid catchment, San Pedro River, Arizona. Journal of
Geophysical Research: Biogeosciences 112:G03S04.
Brooks, P. D., and M. M. Lemon. 2007. Spatial variability in dissolved organic matter and inorganic
nitrogen concentrations in a semiarid stream, San Pedro River, Arizona. Journal of Geophysical
Research: Biogeosciences 112:G03S05.
Brooks, R. J., P. J. Wigington, D. L. Phillips, R. Comeleo, and R. Coulombe. 2012. Willamette River basin
surface water isoscape (8180 and 82H): Temporal changes of source water within the river.
Ecosphere 3:39.
Brooks, R. P., and T. L. Serfass. 2013. Wetland-riparian wildlife of the Mid-Atlantic Region: An overview.
Pages 259-268 in Mid-Atlantic freshwater wetlands: Advances in wetlands science, management,
policy, and practice. R. P. Brooks and D. H. Wardrop, editors. Springer, New York, NY.
Brooks, R. R., J. A. McCleave, and E. K. Schofield. 1977. Cobalt and nickel uptake by the Nyssaceae. Taxon
26:197-201.
Brooks, R. T., and E. A. Colburn. 2011. Extent and channel morphology of unmapped headwater stream
segments of the Quabbin watershed, Massachusetts. Journal of the American Water Resources
Association 47:158-168.
Brosofske, K. D., J. Q. Chen, R. J. Naiman, and J. F. Franklin. 1997. Harvesting effects on microclimatic
gradients from small streams to uplands in western Washington. Ecological Applications 7:1188-
1200.
Brown, J. H., and A. Kodric-Brown. 1977. Turnover rates in insular biogeography: Effect of immigration
on extinction. Ecology 58:445-449.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis
-------
Brown, T. G., and G. F. Hartman. 1988. Contribution of seasonally flooded lands and minor tributaries to
the production of coho salmon in Carnation Creek, British Columbia. Transactions of the American
Fisheries Society 117:546-551.
Brummer, C.}., T. B. Abbe, J. R. Sampson, and D. R. Montgomery. 2006. Influence of vertical channel
change associated with wood accumulations on delineating channel migration zones, Washington,
USA. Geomorphology 80:295-309.
Brunet, N. N., and C. J. Westbrook. 2012. Wetland drainage in the Canadian prairies: Nutrient, salt and
bacteria characteristics. Agriculture, Ecosystems & Environment 146:1-12.
Brunke, M., and T. 0. M. Gonser. 1997. The ecological significance of exchange processes between rivers
and groundwater. Freshwater Biology 37:1-33.
Buffington, J. M., and D. Tonina. 2009. Hyporheic exchange in mountain rivers II: Effects of channel
morphology on mechanics, scales, and rates of exchange. Geography Compass 3:1038-1062.
Bukaveckas, P. A. 2009. Rivers. Pages 721-732 in Encyclopedia of inland waters. G. E. Likens, editor.
Elsevier, Oxford, UK.
Bull, W. B. 1991. Geomorphic responses to climate change. Oxford University Press, New York, NY.
Bull, W. B., and K. M. Scott. 1974. Impact of mining gravel from urban stream beds in the southwestern
United States. Geology 2:171-174.
Bullock, A., and M. Acreman. 2003. The role of wetlands in the hydrological cycle. Hydrology and Earth
System Sciences 7:358-389.
Bunn, S. E., and A. H. Arthington. 2002. Basic principles and ecological consequences of altered flow
regimes for aquatic biodiversity. Environmental Management 30:492-507.
Bunn, S. E., M. C. Thorns, S. K. Hamilton, and S. J. Capon. 2006. Flow variability in dryland rivers: Boom,
bust and the bits in between. River Research and Applications 22:179-186.
Burchsted, D., M. Daniels, R. Thorson, and J. Vokoun. 2010. The river discontinuum: Applying beaver
modifications to baseline conditions for restoration of forested headwaters. BioScience 60:908-922.
Burkart, M. R., W. W. Simpkins, P. J. Squillace, and M. Helmke. 1999. Tributary stream infiltration as a
source of herbicides in an alluvial aquifer. Journal of Environmental Quality 28:69-74.
Burkholder, B. K., G. E. Grant, R. Haggerty, T. Khangaonkar, and P. J. Wampler. 2008. Influence of
hyporheic flow and geomorphology on temperature of a large, gravel-bed river, Clackamas River,
Oregon, USA. Hydrological Processes 22:941-953.
Burns, D. A. 1996. The effects of liming an Adirondack lake watershed on downstream water chemistry.
Biogeochemistry 32:339-362.
Burt, T. P. 1997. The hydrological role of buffer zones within the drainage basin system. Pages 21-32 in
Buffer zones: Their processes and potential in water protection. Proceedings of the International
Conference on Buffer Zones, September 1996. N. E. Haycock, T. P. Burt, K. W. T. Goulding, and G.
Pinay, editors. Quest Environmental, Hertfordshire, UK.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Burt, T. P., G. Pinay, F. E. Matheson, N. E. Haycock, A. Butturini, J.-C. Clement, S. Danielescu, D. J. Dowrick,
M. M. Hefting, A. Hillbricht-Ilkowska, and V. Maitre. 2002. Water table fluctuations in the riparian
zone: Comparative results from a pan-European experiment. Journal of Hydrology 265:129-148.
Butterfield, B. P. 2005. Rana sphenocephala, Southern leopard frog. Pages 586-587 in Amphibian
declines: The conservation status of United States species. M. Lannoo, editor. University of California
Press, Berkeley, CA.
Buttle, J. 2006. Mapping first-order controls on streamflow from drainage basins: The T3 template.
Hydrological Processes 20:3415-3422.
Cairns, M. A., J. L. Ebersole, J. P. Baker, and P. J. Wigington. 2005. Influence of summer stream
temperatures on black spot infestation of juvenile coho salmon in the Oregon Coast Range.
Transactions of the American Fisheries Society 134:1471-1479.
Caissie, D. 2006. The thermal regime of rivers: A review. Freshwater Biology 51:1389-1406.
Calabrese, J. M., and W. F. Pagan. 2004. A comparison-shopper's guide to connectivity metrics. Frontiers
in Ecology and the Environment 2:529-536.
Callahan, M. K., M. C. Rains, J. C. Bellino, C. M. Walker, S. J. Baird, D. F. Whigham, and R. S. King. 2015.
Controls on temperature in salmonid-bearing headwater streams in two common hydrogeologic
settings, Kenai Peninsula, Alaska. Journal of the American Water Resources Association: doi:
10.1111/jawr.l2235.
Cantone, J., and A. Schmidt. 2011. Dispersion mechanisms and the effect of parameter uncertainty on
hydrologic response in urban catchments. Water Resources Research 47:W05503.
Capers, R. S., R. Selsky, and G. J. Bugbee. 2010. The relative importance of local conditions and regional
processes in structuring aquatic plant communities. Freshwater Biology 55:952-966.
Carey, R., and K. Migliaccio. 2009. Contribution of wastewater treatment plant effluents to nutrient
dynamics in aquatic systems: A review. Environmental Management 44:205-217.
Carlyle, G. C., and A. R. Hill. 2001. Groundwater phosphate dynamics in a river riparian zone: Effects of
hydrologic flowpaths, lithology, and redox chemistry. Journal of Hydrology 247:151-168.
Caruso, B. S., and J. Haynes. 2011. Biophysical-regulatory classification and profiling of streams across
management units and ecoregions. Journal of the American Water Resources Association 47:386-
407.
Castro, N. M., and G. M. Hornberger. 1991. Surface-surface water interactions in an alluviated mountain
stream channel. Water Resources Research 27:1613-1621.
Cayan, D. R., and D. H. Peterson. 1989. The influence of North Pacific atmospheric circulation on
streamflow in the West. Geophysical Monographs 55:375-397.
Chacon, N., N. Dezzeo, M. Rangel, and S. Flores. 2008. Seasonal changes in soil phosphorus dynamics and
root mass along a flooded tropical forest gradient in the lower Orinoco River, Venezuela.
Biogeochemistry 87:157-168.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Chahinian, N., C. Bancon-Montigny, A. Caro, P. Got, J. L. Perrin, D. Rosain, C. Rodier, B. Picot, and M. G.
Tournoud. 2012. The role of river sediments in contamination storage downstream of a waste water
treatment plant in low flow conditions: Organotins, faecal indicator bacteria and nutrients.
Estuarine, Coastal and Shelf Science 114:70-81.
Chapra, S. C. 1996. Surface water quality modeling. McGraw-Hill, New York, NY.
Chaput-Bardy, A., C. Fleurant, C. Lemaire, and J. Secondi. 2009. Modelling the effect of in-stream and
overland dispersal on gene flow in river networks. Ecological Modelling 220:3589-3598.
Chartin, C., 0. Evrard, Y. Onda, J. Patin, I. Lefevre, C. Ottle, S. Ayrault, H. Lepage, and P. Bonte. 2013.
Tracking the early dispersion of contaminated sediment along rivers draining the Fukushima
radioactive pollution plume. Anthropocene 1:23-34.
Chase, E., J. Hunting, C. Staley, and V. J. Harwood. 2012. Microbial source tracking to identify human and
ruminant sources of faecal pollution in an ephemeral Florida river. Journal of Applied Microbiology
113:1396-1406.
Cheesman, A. W., E. J. Dunne, B. L. Turner, and K. R. Reddy. 2010. Soil phosphorus forms in hydrologically
isolated wetlands and surrounding pasture uplands. Journal of Environmental Quality 39:1517-
1525.
Chen, X., and X. Chen. 2003. Stream water infiltration, bank storage, and storage zone changes due to
stream-stage fluctuations. Journal of Hydrology 280:246-264.
Cheng, J. D. 1988. Subsurface stormflows in the highly permeable forested watersheds of southwestern
British Columbia. Journal of Contaminant Hydrology 3:171-191.
Chick, J. H., R. J. Cosgriff, and L. S. Gittinger. 2003. Fish as potential dispersal agents for floodplain plants:
First evidence in North America. Canadian Journal of Fisheries and Aquatic Sciences 60:1437-1439.
Chin, A., and K. J. Gregory. 2001. Urbanization and adjustment of ephemeral stream channels. Annals of
the Association of American Geographers 91:595-608.
Chipps, S. R., D. E. Hubbard, K. B. Werline, N. J. Haugerud, K. A. Powell, J. Thompson, and T. Johnson.
2006. Association between wetland disturbance and biological attributes in floodplain wetlands.
Wetlands 26:297-508.
Church, M. 2002. Geomorphic thresholds in riverine landscapes. Freshwater Biology 47:541-557.
Church, M. 2006. Bed material transport and the morphology of alluvial river channels. Annual Review
of Earth and Planetary Sciences 34:325-354.
Cirmo, C. P., C. T. Driscoll, and K. Bowes. 2000. Chemical fluxes from sediments in two Adirondack
wetlands: Effects of an acid-neutralization experiment. Soil Science Society of America Journal
64:790-799.
Clair, T. A., T. L. Pollock, and J. M. Ehrman. 1994. Exports of carbon and nitrogen from river basins in
Canada's Atlantic Provinces. Global Biogeochemical Cycles 8:441-450.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Clark, W. R. 2000. Ecology of muskrats in prairie wetlands. Pages 287-313 in Prairie wetland ecology:
The contribution of the marsh ecology research program. H. R. Murkin, A. G. van der Valk, and W. R.
Clark, editors. Iowa State University Press, Ames, IA.
Claxton, A.}., P. D. Bates, and H. L. Cloke. 2003. Mixing of hillslope, river, and alluvial ground waters in
lowland floodplains. Ground Water 41:926-936.
Clobert,}., J.-F. Le Galliard, J. Cote, S. Meylan, and M. Massot. 2009. Informed dispersal, heterogeneity in
animal dispersal syndromes and the dynamics of spatially structured populations. Ecology Letters
12:197-209.
Goes, A. L., and D. R. Pool. 2005. Ephemeral-stream channel and basin-floor infiltration and recharge in
the Sierra Vista subwatershed of the Upper San Pedro Basin, southeastern Arizona. USGS Open-File
Report 2005-1023, U.S. Department of the Interior, U.S. Geological Survey, Washington, DC.
Cohen, M., J. Paris, and M. Clark. 2007. P-sorption capacity estimation in southeastern USA wetland soils
using visible/near infrared (VNIR) reflectance spectroscopy. Wetlands 27:1098-1111.
Collen, P., and R. J. Gibson. 2001. The general ecology of beavers (Castor spp.), as related to their
influence on stream ecosystems and riparian habitats, and the subsequent effects on fish - a review.
Reviews in Fish Biology and Fisheries 10:439-461.
Collier, M., R. J. Webb, and J. C. Schmidt. 1996. Dams and rivers: A primer on the downstream effects of
dams. USGS Circular 1126, U.S. Department of the Interior, U.S. Geological Survey, Reston, VA.
Collins, B. D., D. R. Montgomery, K. L. Fetherston, and T. B. Abbe. 2012. The floodplain large-wood cycle
hypothesis: A mechanism for the physical and biotic structuring of temperature forested alluvial
valleys in the North Pacific coastal ecoregion. Geomorphology 139-140:460-470.
Colvin, M. E., and C. M. Moffitt. 2009. Evaluation of irrigation canal networks to assess stream
connectivity in a watershed. River Research and Applications 25:486-496.
Colvin, R., G. R. Giannico, J. Li, K. L. Boyer, and W. J. Gerth. 2009. Fish use of intermittent watercourses
draining agricultural lands in the Upper Willamette River Valley, Oregon. Transactions of the
American Fisheries Society 138:1302-1313.
Comer, P., K. Goodin, A. Tomaino, G. Hammerson, G. Kittel, S. Menard, C. Nordman, M. Pyne, M. Reid, L.
Sneddon, and K. Snow. 2005. Biodiversity values of geographically isolated wetlands in the United
States. NatureServe, Arlington, VA.
Compton, J. E., M. R. Church, S. T. Larned, and W. E. Hogsett. 2003. Nitrogen export from forested
watersheds in the Oregon Coast Range: The role of N2-fixing red alder. Ecosystems 6:773-785.
Constantz, J. 1998. Interaction between stream temperature, streamflow, and groundwater exchanges in
alpine streams. Water Resources Research 34:1609-1615.
Constantz, J., A. E. Stewart, R. Niswonger, and L. Sarma. 2002. Analysis of temperature profiles for
investigating stream losses beneath ephemeral channels. Water Resources Research 38:1316.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Cook, B.}., and F. R. Hauer. 2007. Effects of hydrologic connectivity on water chemistry, soils, and
vegetation structure and function in an intermontane depressional wetland landscape. Wetlands
27:719-738.
Cook, N., F. J. Rahel, and W. A. Hubert. 2010. Persistence of Colorado River cutthroat trout populations in
isolated headwater streams of Wyoming. Transactions of the American Fisheries Society 139:1500-
1510.
Cooper, A., J. W. Gilliam, R. B. Daniels, and W. P. Robarge. 1987. Riparian areas as filters for agricultural
sediment. Soil Science Society of America Journal 51:416-420.
Copp, G. H. 1989. The habitat diversity and fish reproductive function of floodplain ecosystems.
Environmental Biology of Fishes 26:1-27.
Corenblit, D., E. Tabacchi, J. Steiger, and A. M. Gurnell. 2007. Reciprocal interactions and adjustments
between fluvial landforms and vegetation dynamics in river corridors: A review of complementary
approaches. Earth-Science Reviews 84:56-86.
Corti, R., and T. Datry. 2012. Invertebrates and sestonic matter in an advancing wetted front travelling
down a dry river bed (Albarine, France). Freshwater Science 31:1187-1201.
Corti, R., T. Datry, L. Drummond, and S. T. Larned. 2011. Natural variation in immersion and emersion
affects breakdown and invertebrate colonization of leaf litter in a temporary river. Aquatic Sciences
73:537-550.
Costelloe, J. F., A. Shields, R. B. Grayson, and T. A. McMahon. 2007. Determining loss characteristics of
arid zone river waterbodies. River Research and Applications 23:715-731.
Cote, D., D. Kehler, C. Bourne, and Y. Wiersma. 2009. A new measure of longitudinal connectivity for
stream networks. Landscape Ecology 24:101-113.
Coulthard, T. J. 2005. Effects of vegetation on braided stream pattern and dynamics. Water Resources
Research 41:W04003.
Coutant, C. C. 1999. Perspectives on temperature in the Pacific Northwest's fresh waters. ORNL/TM-
1999/44, Oak Ridge National Laboratory Environmental Sciences Division Publication 4849,
Springfield, VA.
Covich, A. P., M. A. Palmer, and T. A. Crowl. 1999. The role of benthic invertebrate species in freshwater
ecosystems: Zoobenthic species influence energy flows and nutrient cycling. BioScience 49:119-127.
Cowardin, L. M., V. Carter, F. C. Golet, and E. T. LaRoe. 1979. Classification of wetlands and deepwater
habitats of the United States. U.S. Department of the Interior, U.S. Fish and Wildlife Service, Office of
Biological Services, Washington, DC.
Craft, C., and W. Casey. 2000. Sediment and nutrient accumulation in floodplain and depressional
freshwater wetlands of Georgia, USA. Wetlands 20:323-332.
Craft, C. B., and C. Chiang. 2002. Forms and amounts of soil nitrogen and phosphorus across a longleaf
pine-depressional wetland landscape. Soil Science Society of America Journal 66:1713-1721.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Creed, I. F., and F. D. Beall. 2009. Distributed topographic indicators for predicting nitrogen export from
headwater catchments. Water Resources Research 45:W10407.
Creed, I. F., S. E. Sanford, F. D. Beall, L. A. Molot, and P. J. Dillon. 2003. Cryptic wetlands: Integrating
hidden wetlands in regression models of the export of dissolved organic carbon from forested
landscapes. Hydrological Processes 17:3629-3648.
Cristea, N. C., and S. J. Burges. 2009. Use of thermal infrared imagery to complement monitoring and
modeling of spatial stream temperatures. Journal of Hydrologic Engineering 14:1080-1090.
Cross, F. B., and R. E. Moss. 1987. Historic changes in fish communities and aquatic habitats in plains
streams of Kansas. Pages 155-165 in Community and evolutionary ecology of North American
stream fishes. W. J. Matthews and D. C. Heins, editors. University of Oklahoma Press, Norman, OK.
Cuffney, T. F., J. B. Wallace, and G. J. Luthgart. 1990. Experimental evidence quantifying the role of
benthic invertebrates in organic matter dynamics of headwater streams. Freshwater Biology
23:281-299.
Cummins, K. W., C. E. Gushing, and G. W. Minshall. 2006. Introduction: An overview of stream
ecosystems. Pages 1-8 in River and stream ecosystems of the world. C. E. Gushing, K. W. Cummins,
and G. W. Minshall, editors. University of California Press, Berkley, CA.
Cummins, K. W., and M. J. Klug. 1979. Feeding ecology of stream invertebrates. Annual Review of Ecology
and Systematics 10:147-172.
Cummins, K. W., R. C. Petersen, F. 0. Howard, J. C. Wuycheck, and V. I. Holt. 1973. Utilization of leaf litter
by stream detritivores. Ecology 54:336-345.
Cummins, K. W., G. L. Spengler, G. M. Ward, R. M. Speaker, R. W. Ovink, D. C. Mahan, and R. L. Mattingly.
1980. Processing of confined and naturally entrained leaf litter in a woodland stream ecosystem.
Limnology and Oceanography 25:952-957.
Cummins, K. W., M. A. Wilzbach, D. M. Gates, J. B. Perry, and W. B. Taliferro. 1989. Shredders and riparian
vegetation - Leaf litter that falls into streams influences communities of stream invertebrates.
BioScience 39:24-30.
Curriero, F. C., J. A. Patz, J. B. Rose, and S. Lele. 2001. The association between extreme precipitation and
waterborne disease outbreaks in the United States, 1948-1994. American Journal of Public Health
91:1194-1199.
Curry, R. A., C. Brady, D. L. G. Noakes, and R. G. Danzmann. 1997. Use of small streams by young brook
trout spawned in a lake. Transactions of the American Fisheries Society 126:77-83.
Dabney, S. M., L. D. Meyer, W. C. Harmon, C. V. Alonso, and G. R. Foster. 1995. Depositional patterns of
sediment trapped by grass hedges. Transactions of the American Society of Agricultural Engineers
38:1719-1729.
Dahl, T. E. 1990. Wetlands Losses in the United States 1780's to 1980's. U.S. Department of the Interior,
U.S. Fish and Wildlife Service, Washington, DC.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Dahm, C. N., N. B. Grimm, P. Marmonier, H. M. Valett, and P. Vervier. 1998. Nutrient dynamics at the
interface between surface waters and groundwaters. Freshwater Biology 40:427-451.
Daniels, R. B., and J. G. Gilliam. 1996. Sediment and chemical load reduction by grass and riparian filters.
Soil Science Society of America Journal 60:246-251.
Banks, H. V. 2007. How aquatic insects live in cold climates. Canadian Entomologist 139:443-471.
Darracq, A., and G. Destouni. 2005. In-stream nitrogen attenuation: Model-aggregation effects and
implications for coastal nitrogen impacts. Environmental Science & Technology 39:3716-3722.
Darracq, A., and G. Destouni. 2007. Physical versus biogeochemical interpretations of nitrogen and
phosphorus attenuation in streams and its dependence on stream characteristics. Global
Biogeochemical Cycles 21:GB3003.
David, A., C. Bancon-Montigny, C. Salles, C. Rodier, and M.-G. Tournoud. 2012. Contamination of riverbed
sediments by hazardous substances in the Mediterranean context: Influence of hydrological
conditions. Journal of Hydrology 468-469:76-84.
Davis, C. B., J. L. Baker, A. G. van der Valk, and C. E. Beer. 1981. Prairie pothole marshes as traps for
nitrogen and phosphorus in agricultural runoff. Pages 153-163 in Selected proceedings of the
Midwest conference on wetland values and management, June 17-19,1981, St. Paul, MN. B.
Richardson, editor. The Freshwater Society, St Paul, MN.
Davis, J. H., S. M. Griffith, and P. J. Wigington. 2011. Surface and ground water nitrogen dynamics in a
well-drained riparian forest within a poorly-drained agricultural landscape. Journal of
Environmental Quality 40:505-516.
Day, D. G. 1978. Drainage density changes during rainfall. Earth Surface Processes and Landforms 3:319-
326.
de Vries, J. J. 1995. Seasonal expansion and contraction of stream networks in shallow groundwater
systems. Journal of Hydrology 170:15-26.
Deiner, K., J. C. Garza, R. Coey, and D. J. Girman. 2007. Population structure and genetic diversity of trout
(Oncorhynchus mykiss) above and below natural and man-made barriers in the Russian River,
California. Conservation Genetics 8:437-454.
Demars, B. 0. L., J. Russell Manson, J. S. Olafsson, G. M. Gi'slason, R. Gudmundsdottir, G. U. Y. Woodward, J.
Reiss, D. E. Pichler, J. J. Rasmussen, and N. Friberg. 2011. Temperature and the metabolic balance of
streams. Freshwater Biology 56:1106-1121.
Detenbeck, N. E., C. A. Johnston, and G. J. Niemi. 1993. Wetland effects on lake water quality in the
Minneapolis/St. Paul metropolitan area. Landscape Ecology 8:39-61.
Devito, K., I. Creed, T. Can, C. Mendoza, R. Petrone, U. Silins, and B. Smerdon. 2005. A framework for
broad-scale classification of hydrologic response units on the Boreal Plain: Is topography the last
thing to consider? Hydrological Processes 19:1705-1714.
Devito, K. J., A. R. Hill, and N. Roulet. 1996. Groundwater-surface water interactions in headwater
forested wetlands of the Canadian Shield. Journal of Hydrology 181:127-147.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Bias, M. S., T. Oberdorff, B. Hugueny, F. Leprieur, C. Jezequel, J.-F. Cornu, S. Brosse, G. Grenouillet, and P.
A. Tedesco. 2014. Global imprint of historical connectivity on freshwater fish biodiversity. Ecology
Letters 17:1130-1140.
Dickinson, J. E., J. R. Kennedy, D. R. Pool, J. T. Cordova, J. T. Parker, J. P. Macy, and B. Thomas. 2010.
Hydrogeologic framework of the middle San Pedro watershed, southeastern Arizona. USGS Scientific
Investigations Report 2010-5126, prepared in cooperation with the Arizona Department of Water
Resources, U.S. Department of the Interior, U.S. Geological Survey, Reston, VA.
Dierberg, F. E., and P. L. Brezonik. 1984. Nitrogen and phosphorus mass balances in a cypress dome
receiving wastewater. Pages 112-118 in Cypress swamps. K. C. Ewel and H. T. Odum, editors.
University Press of Florida, Gainesville, FL.
Dieter, D., D. von Schiller, E. M. Garci'a-Roger, M. M. Sanchez-Montoya, R. Gomez, J. Mora-Gomez, F.
Sangiorgio, J. Gelbrecht, and K. Tockner. 2011. Preconditioning effects of intermittent stream flow on
leaf litter decomposition. Aquatic Sciences 73:599-609.
Dietrich, W. E., and T. Dunne. 1993. The channel head. Pages 175-219 in Channel network hydrology. K.
Beven and M. J. Kirby, editors. John Wiley & Sons, New York, NY.
Dillaha, T. A., and S. P. Inamdar. 1997. Buffer zones as sediment traps or sources. Pages 33-42 in Buffer
zones: Their processes and potential in water protection. Proceedings of the International
Conference on Buffer Zones, September 1996. N. E. Haycock, T. P. Burt, K. W. T. Goulding, and G.
Pinay, editors. Quest Environmental, Hertfordshire, UK.
Dillaha, T. A., R. B. Reneau, S. Mostaghimi, and D. Lee. 1989. Vegetative filter strips for agricultural
nonpoint source pollution control. Transactions of the American Society of Agricultural Engineers
32:513-519.
Dillon, P. J., and L. A. Molot. 1997. Effects of landscape form on export of dissolved organic carbon, iron,
and phosphorus from forested stream catchments. Water Resources Research 33:2591-2600.
Dosskey, M. G. 2001. Toward quantifying water pollution abatement in response to installing buffers on
crop land. Environmental Management 28:577-598.
Dou, P., B. Cui, T. Xie, D. Dong, and B. Gu. 2015. Macrobenthos diversity response to hydrological
connectivity gradient. Wetlands: doi: 10.1007/sl3157-13014-10580-13158.
Doyle, M. W., E. H. Stanley, and J. M. Harbor. 2003. Hydrogeomorphic controls on phosphorus retention
in streams. Water Resources Research 39:1147.
Driscoll, C. T., V. Blette, C. Yan, C. L. Schofield, R. Munson, and J. Holsapple. 1995. The role of dissolved
organic carbon in the chemistry and bioavailability of mercury in remote Adirondack lakes. Water
Air and Soil Pollution 80:499-508.
Dugdale, S. J., N. E. Bergeron, and A. St-Hilaire. 2013. Temporal variability of thermal refuges and water
temperature patterns in an Atlantic salmon river. Remote Sensing of Environment 136:358-373.
Duncan, S. H., R. E. Bilby, J. W. Ward, and J. T. Heffner. 1987. Transport of road-surface sediment through
ephemeral stream channel. Water Resources Bulletin 23:113-119.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Dunkerley, D. L. 1992. Channel geometry, bed material, and inferred flow conditions in ephemeral
stream systems, Barrier Range, western N.S.W. Australia. Hydrological Processes 6:417-433.
Dunne, E.}., K. R. Reddy, and M. K. Clark. 2006. Phosphorus release and retention by soils of natural
isolated wetlands. International Journal of Environment and Pollution 28:496-516.
Dunne, E.}.,}. Smith, D. B. Perkins, M. W. Clark, J. W. Jawitz, and K. R. Reddy. 2007. Phosphorus storages
in historically isolated wetland ecosystems and surrounding pasture uplands. Ecological
Engineering 31:16-28.
Dunne, T., and R. D. Black. 1970. Partial area contributions to storm runoff in a small New England
watershed. Water Resources Research 6:1296-1311.
Dunne, T., and L. B. Leopold. 1978. Water in environmental planning. W.H. Freeman and Co., San
Francisco, CA.
Eaton, L. S., B. A. Morgan, R. C. Kochel, and A. D. Howard. 2003. Role of debris flows in long-term
landscape denudation in the central Appalachians of Virginia. Geology 31:339-342.
Eberle, L. C., and J. A. Stanford. 2010. Importance and seasonal availability of terrestrial invertebrates as
prey for juvenile salmonids in floodplain spring brooks of the Kol River (Kamchatka, Russian
Federation). River Research and Applications 26:682-694.
Ebersole, J. L., M. E. Colvin, P. J. Wigington, S. G. Leibowitz, J. P. Baker, M. R. Church, J. E. Compton, B. A.
Miller, M. A. Cairns, B. P. Hansen, and H. R. LaVigne. 2009. Modeling stream network-scale variation
in coho salmon overwinter survival and smolt size. Transactions of the American Fisheries Society
138:564-580.
Ebersole, J. L., W. J. Liss, and C. A. Frissell. 2003. Cold water patches in warm streams: Physicochemical
characteristics and the influence of shading. Journal of the American Water Resources Association
39:355-368.
Ebersole, J. L., P. J. Wigington, J. P. Baker, M. A. Cairns, M. R. Church, B. P. Hansen, B. A. Miller, H. R.
LaVigne, B. W. Compton, and S. G. Leibowitz. 2006. Juvenile coho salmon growth and survival across
stream network seasonal habitats. Transactions of the American Fisheries Society. Transactions of
the American Fisheries Society 135:1681-1697.
Ebersole, J. L., P. J. Wigington, S. G. Leibowitz, R. L. Comeleo, and J. V. Sickle. 2015. Predicting the
occurrence of cold-water patches at intermittent and ephemeral tributary confluences with warm
rivers. Freshwater Science: doi:10.1086/678127.
Eckhardt, B. W., and T. R. Moore. 1990. Controls on dissolved organic carbon concentrations in streams
of southern Quebec. Canadian Journal of Fisheries and Aquatic Sciences 47:1537-1544.
Edwards, E. D., and A. D. Huryn. 1995. Annual contribution of terrestrial invertebrates to a New Zealand
trout stream. New Zealand Journal of Marine and Freshwater Research 29:467-477.
Elliott, J. M. 1971. Distances travelled by drifting invertebrates in a Lake District stream. Oecologia
6:350-379.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Elliott, J. M. 2003. A comparative study of the dispersal of 10 species of stream invertebrates. Freshwater
Biology 48:1652-1668.
Elmore, A.}., and S. S. Kaushal. 2008. Disappearing headwaters: Patterns of stream burial due to
urbanization. Frontiers in Ecology and the Environment 6:308-312.
Elwood, J. W., J. D. Newbold, R. V. O'Neill, and W. Van Winkle. 1983. Resource spiralling: An operational
paradigm for analyzing lotic ecosystems. Pages 3-23 in Dynamics of lotic ecosystems. T. D. Fontaine
and S. M. Bartell, editors. Ann Arbor Science Publishers, Ann Arbor, MI.
Ensign, S. H., and M. W. Doyle. 2006. Nutrient spiraling in streams and river networks. Journal of
Geophysical Research: Biogeosciences 111:G04009.
Epperson, B. K., B. H. McRae, K. Scribner, S. A. Cushman, M. S. Rosenberg, M. J. Fortin, P. M. A. James, M.
Murphy, S. Manel, P. Legendre, and M. R. T. Dale. 2010. Utility of computer simulations in landscape
genetics. Molecular Ecology 19:3549-3564.
Erman, D. C., and V. M. Hawthorne. 1976. The quantitative importance of an intermittent stream in the
spawning of rainbow trout. Transactions of the American Fisheries Society 105:675-681.
Eros, T., J. Olden, R. Schick, D. Schmera, and M.-J. Fortin. 2012. Characterizing connectivity relationships
in freshwaters using patch-based graphs. Landscape Ecology 27:303-317.
Eshelman, K. N., and H. F. Hemond. 1985. The role of organic acids in the acid-base status of surface
waters at Bickford Watershed, Massachusetts. Water Resources Research 21:1503-1510.
Euliss, N. H., J. W. Labaugh, L. H. Fredrickson, D. M. Mushet, M. R. K. Laubhan, G. A. Swanson, T. C. Winter,
D. 0. Rosenberry, and R. D. Nelson. 2004. The wetland continuum: A conceptual framework for
interpreting biological studies. Wetlands 24:448-458.
Ewel, K. C., and H. T. Odum. 1984. Cypress swamps. University Presses of Florida, Gainesville, FL.
Pagan, W. F. 2002. Connectivity, fragmentation, and extinction risk in dendritic metapopulations.
Ecology 83:3243-3249.
Falke, J. A., K. R. Bestgen, and K. D. Fausch. 2010. Streamflow reductions and habitat drying affect
growth, survival, and recruitment of brassy minnow across a Great Plains landscape. Transactions of
the American Fisheries Society 139:1566-1583.
Falke, J. A., and K. B. Gido. 2006. Effects of reservoir connectivity on stream fish assemblages in the Great
Plains. Canadian Journal of Fisheries and Aquatic Sciences 63:480-493.
Farmer, A. H., and A. H. Parent. 1997. Effects of the landscape on shorebird movements at spring
migration stopovers. The Condor 99:698-707.
Farrar, D., L. C. Alexander, L. L. Yuan, and J. Gerritsen. 2014. Regional observational studies: Addressing
confounding. Pages 187-202 in Ecological causal assessment. S. B. Norton, S. M. Cormier, and G. W.
Suter, editors. CRC Press, London, UK.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Fausch, K. D., and K. R. Bestgen. 1997. Ecology of fishes indigenous to the central and southwestern
Great Plains. Pages 131-166 in Ecology and conservation of Great Plains vertebrates. F. L. Knopf and
F. B. Samson, editors. Springer-Verlag, New York, NY.
Fausch, K. D., B. E. Rieman, J. B. Dunham, M. K. Young, and D. P. Peterson. 2009. Invasion versus isolation:
Trade-offs in managing native salmonids with barriers to upstream movement. Conservation
Biology 23:859-870.
Fausch, K. D., C. E. Torgersen, C. V. Baxter, and H. W. Li. 2002. Landscapes to riverscapes: Bridging the
gap between research and conservation of stream fishes. BioScience 52:483-498.
Fekete, B. M., C. J. Vorb'smarty, and W. Grabs. 2002. High-resolution fields of global runoff combining
observed river discharge and simulated water balances. Global Biogeochemical Cycles 16:15-11-15-
10.
Fellman, J. B., K. C. Petrone, and P. F. Grierson. 2013. Leaf litter age, chemical quality, and
photodegradation control the fate of leachate dissolved organic matter in a dryland river. Journal of
Arid Environments 89:30-37.
Feminella, J. W. 1996. Comparison of benthic macroinvertebrate assemblages in small streams along a
gradient of flow permanence. Journal of the North American Benthological Society 15:651-669.
Feminella, J. W., and C. P. Hawkins. 1995. Interactions between stream herbivores and periphyton: A
quantitative analysis of past experiments. Journal of the North American Benthological Society
14:465-509.
Per, T., and Z. Hroudova. 2008. Detecting dispersal of Nuphar lutea in river corridors using microsatellite
markers. Freshwater Biology 53:1409-1422.
Ferguson, C., A. M. d. R. Husman, N. Altavilla, D. Deere, and N. Ashbolt. 2003. Fate and transport of
surface water pathogens in watersheds. Critical Reviews in Environmental Science and Technology
33:299-361.
Ferguson, R., and T. Hoey. 2008. Effects of tributaries on main-channel geomorphology. Pages 183-208
in River confluences, tributaries and the fluvial network. S. P. Rice, A. G. Roy, and B. L. Rhoads,
editors. John Wiley & Sons, Chichester, UK.
Ferguson, R. I., J. R. Cudden, T. Hoey, and S. P. Rice. 2006. River system discontinuities due to lateral
inputs: Generic styles and controls. Earth Surface Processes and Landforms 31:1149-1166.
Fernald, A. F., D. Landers, and P. J. Wigington. 2006. Water quality changes along hyporheic flow paths
between a large gravel bar river and off-channel alcoves in Oregon, USA.. River Research and
Applications 22:1111-1114.
Fernald, A. F., P. J. Wigington, and D. Landers. 2001. Transient storage and hyporheic flow along the
Willamette River, Oregon: Field measurements and model estimates. Water Resources Research
37:1681-1694.
Ferone, J. M., and K. J. Devito. 2004. Shallow groundwater-surface water interactions in pond-peatland
complexes along a Boreal Plains topographic gradient. Journal of Hydrology 292:75-95.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Figuerola,}., and A. J. Green. 2002. Dispersal of aquatic organisms by waterbirds: A review of past
research and priorities for future studies. Freshwater Biology 47:483-494.
Figuerola,}., A. J. Green, and T. C. Michot. 2005. Invertebrate eggs can fly: Evidence of waterfowl-
mediated gene flow in aquatic invertebrates. The American Naturalist 165:274-280.
Finkelstein, P. L, and L. E. Truppi. 1991. Spatial distribution of precipitation seasonality in the United
States. Journal of Climate 4:373-385.
Fisher, S. G., L. J. Gray, N. B. Grimm, and D. E. Busch. 1982. Temporal succession in a desert stream
ecosystem following flash flooding. Ecological Monographs 52:93-110.
Fisher, S. G., and G. E. Likens. 1973. Energy flow in Bear Brook, New Hampshire: An integrative approach
to stream ecosystem metabolism. Ecological Monographs 43:421-439.
Fisher, S. G., J. Welter, J. Schade, and J. Henry. 2001. Landscape challenges to ecosystem thinking:
Creative flood and drought in the American Southwest Scientia Marina 65 (Supplement 2): 181-
192.
Fisher, S. J., and D. W. Willis. 2000. Seasonal dynamics of aquatic fauna and habitat parameters in a
perched upper Missouri River wetland. Wetlands 20:470-478.
Fleming, S. W., P. H. Whitfield, R. D. Moore, and E. J. Quilty. 2007. Regime-dependent streamflow
sensitivities to Pacific climate modes cross the Georgia-Pugettransboundary ecoregion.
Hydrological Processes 21:3264-3287.
Florsheim, J. L., J. F. Mount, and L. T. Rutten. 2001. Effect of baselevel change on floodplain and fan
sediment storage and ephemeral tributary channel morphology, Navarro River, California. Earth
Surface Processes and Landforms 26:219-232.
Folk, M. J., and T. C. Tacha. 1990. Sandhill crane roost site characteristics in the North Platte River valley.
The Journal of Wildlife Management 54:480-486.
Ford, J., and B. L. Bedford. 1987. The hydrology of Alaskan wetlands, USA: A review. Arctic and Alpine
Research 19:209-229.
Franklin, S., J. Kupfer, S. R. Pezeshki, R. Gentry, and R. D. Smith. 2009. Complex effects of channelization
and levee construction on western Tennessee floodplain forest function. Wetlands 29:451-464.
Fraser, D. F., J. F. Gilliam, and T. Yiphoi. 1995. Predation as an agent of population fragmentation in a
tropical watershed. Ecology 76:1461-1472.
Fraser, D. J., L. K. Weir, L. Bernatchez, M. M. Hansen, and E. B. Taylor. 2011. Extent and scale of local
adaptation in salmonid fishes: Review and meta-analysis. Heredity 106:404-420.
Freeman, M. C., C. M. Pringle, and C. R. Jackson. 2007. Hydrologic connectivity and the contribution of
stream headwaters to ecological integrity at regional scales. Journal of the American Water
Resources Association 43:5-14.
Freeze, A. R., and J. A. Cherry. 1979. Groundwater. Prentice-Hall, Englewood Cliffs, NJ.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Freeze, R. A. 1971. Three-dimensional, transient, saturated-unsaturated flow in a groundwater basin.
Water Resources Research 7:347-366.
Frisch, D., A. J. Green, and J. Figuerola. 2007. High dispersal capacity of a broad spectrum of aquatic
invertebrates via waterbirds. Aquatic Sciences 69:568-574.
Frisch, D., and S. T. Threlkeld. 2005. Flood-mediated dispersal versus hatching: Early recolonisation
strategies of copepods in floodplain ponds. Freshwater Biology 50:323-330.
Frissell, C. A., W. J. Liss, C. E. Warren, and M. D. Hurley. 1986. A hierarchical framework for stream
habitat classification: Viewing streams in a watershed context. Environmental Management 10:199-
214.
Fritz, K. M., and W. K. Dodds. 2002. Macroinvertebrate assemblage structure across a tallgrass prairie
stream landscape. Archiv fur Hydrobiologie 154:79-102.
Fritz, K. M., and W. K. Dodds. 2004. Resistance and resilience of macroinvertebrate assemblages to
drying and flood in a tallgrass prairie stream system. Hydrobiologia 527:99-112.
Fritz, K. M., J. W. Feminella, C. Colson, B. G. Lockaby, R. Governo, and R. B. Rummer. 2006a. Biomass and
decay rates of roots and detritus in sediments of intermittent coastal plain streams. Hydrobiologia
556:265-277.
Fritz, K. M., J. M. Glime, J. Hribljan, and J. L. Greenwood. 2009. Can bryophytes be used to characterize
hydrologic permanence in forested headwater streams? Ecological Indicators 9:681-692.
Fritz, K. M., B. R. Johnson, and D. M. Walters. 2006b. Field operations manual for assessing the hydrologic
permanence and ecological condition of headwater streams. EPA/600/R-06/126, U.S.
Environmental Protection Agency, Office of Research and Development, National Exposure Research
Laboratory, Washington, DC.
Fritz, K. M., B. R. Johnson, and D. M. Walters. 2008. Physical indicators of hydrologic permanence in
forested headwater streams. Journal of the North American Benthological Society 27:690-704.
Fronhofer, E. A., A. Kubisch, F. M. Hilker, T. Hovestadt, and H. J. Poethke. 2012. Why are metapopulations
so rare? Ecology 93:1967-1978.
Fuller, C. C., and J. W. Harvey. 2000. Reactive uptake of trace metals in the hyporheic zone of a mining-
contaminated stream, Final Creek, Arizona. Environmental Science & Technology 34:1150-1155.
Fullerton, A. H., K. M. Burnett, E. A. Steel, R. L. Flitcroft, G. R. Pess, B. E. Feist, C. E. Torgersen, D. J. Miller,
and B. L. Sanderson. 2010. Hydrological connectivity for riverine fish: Measurement challenges and
research opportunities. Freshwater Biology 55:2215-2237.
Gahl, M. K., A. J. K. Calhoun, and R. Graves. 2009. Facultative use of seasonal pools by American bullfrogs
(Rana catesbeiand). Wetlands 29:697-703.
Galat, D. L., L. H. Fredrickson, D. D. Humburg, K. J. Bataille, J. R. Bodie, J. Dohrenwend, G. T. Gelwicks, J. E.
Havel, D. L. Helmers, J. B. Hooker, J. R. Jones, M. F. Knowlton, J. Kubisiak, J. Mazourek, A. C. McColpin,
R. B. Renken, and R. D. Semlitsch. 1998. Flooding to restore connectivity of regulated, large-river
wetlands. BioScience 48:721-733.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Galatowitsch, S. M., and A. G. van der Valk. 1996. The vegetation of restored and natural prairie wetlands.
Ecological Applications 6:102-112.
Galloway, M. E., and B. A. Branfireun. 2004. Hydrological and biogeochemical controls on mercury fate
and transport in a southern Ontario forested wetland. The Science of Total Environment 325:239-
254.
Galpern, P., M. Manseau, and A. Fall. 2011. Patch-based graphs of landscape connectivity: A guide to
construction, analysis and application for conservation. Biological Conservation 144:44-55.
Galster, J. C. 2007. Natural and anthropogenic influences on the scaling of discharge with drainage area
for multiple watersheds. Geosphere 3:260-271.
Galster, J. C. 2009. Testing the linear relationship between peak annual river discharge and drainage
area using long-term USGS river gauging records. Pages 159-171 in Management and restoration of
fluvial systems with broad historical changes and human impacts. L. A. James, S. L. Rathburn, and G.
R. Whittecar, editors. Geological Society of America, Boulder, CO.
Galster, J. C., F. J. Pazzaglia, B. R. Hargreaves, D. P. Morris, S. C. Peters, and R. N. Weisman. 2006. Effects of
urbanization on watershed hydrology: The scaling of discharge with drainage area. Geology 34:713-
716.
Gamble, D. E., J. Grody, J. J. Mack, and M. Micacchion. 2007. An ecological and functional assessment of
urban wetlands in central Ohio. Columbus, Ohio. Ohio EPA Technical Report WET/ 2007-3B, Ohio
Environmental Protection Agency, Wetland Ecology Group, Division of Surface Water, Columbus,
OH.
Ganio, L. M., C. E. Torgersen, and R. E. Gresswell. 2005. A geostatistical approach for describing spatial
pattern in stream networks. Frontiers in Ecology and the Environment 3:138-144.
Gannett, M. W., J. K. E. Lite, D. S. Morgan, and C. A. Collins. 2001. Ground-water hydrology of the Upper
Deschutes Basin, Oregon. USGS Water-Resources Investigations Report 00-4162, U.S. Department of
the Interior, U.S. Geological Survey, Portland, OR.
Gardner, B., and P. J. Sullivan. 2004. Spatial and temporal stream temperature prediction: Modeling
nonstationary temporal covariance structures. Water Resources Research 40:1-9.
Gardner, K. K., and B. L. McGlynn. 2009. Seasonality in spatial variability and influence of land use/land
cover and watershed characteristics on stream water nitrate concentrations in a developing
watershed in the Rocky Mountain West. Water Resources Research 45:W08411.
Gasith, A., and V. H. Resh. 1999. Streams in Mediterranean climate regions: Abiotic influences and biotic
responses to predictable seasonal events. Annual Review of Ecology and Systematics 30:51-81.
Gehrke, G. E., J. D. Blum, and M. Marvin-DiPasquale. 2011. Sources of mercury to San Francisco Bay
surface sediment as revealed by mercury stable isotopes. Geochimica et Cosmochimica Acta 75:691-
705.
Gergel, S. E. 2005. Spatial and non-spatial factors: When do they affect landscape indicators of watershed
loading? Landscape Ecology 20:177-189.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Gergel, S. E., M. G. Turner, and T. K. Kratz. 1999. Dissolved organic carbon as an indicator of the scale of
watershed influence on lakes and rivers. Ecological Applications 9:1377-1390.
Germanoski, D., and D. F. Ritter. 1988. Tributary response to local base level lowering below a dam.
Regulated Rivers: Research & Management 2:11-24.
Gibbons, J. W. 2003. Terrestrial habitat: A vital component for herpetofauna of isolated wetlands.
Wetlands 23:630-635.
Gibbons, J. W., and R. D. Semlitsch. 1991. Guide to amphibians and reptiles of the Savannah River Site.
University of Georgia Press, Athens, GA.
Gibbons, J. W., C. T. Winne, D. E. Scott, J. D. Willson, X. Glaudas, K. M. Andrews, B. D. Todd, L. A. Fedewa, L.
Wilkinson, R. N. Tsaliagos, S. J. Harper, J. L. Greene, T. D. Tuberville, B. S. Metts, M. E. Dorcast, J. P.
Nestor, C. A. Young, T. Akre, R. N. Reed, K. A. Buhlmann, J. Norman, D. A. Croshaw, C. Hagen, and B. B.
Rothermel. 2006. Remarkable amphibian biomass and abundance in an isolated wetland:
Implications for wetland conservation. Conservation Biology 20:1457-1465.
Gilvear, D., and R. Bryant. 2003. Analysis of aerial photography and other remotely sensed data. Pages
135-170 in Tools in fluvial geomorphology. G. M. Kondolf and H. Piegay, editors. John Wiley & Sons,
Chichester, UK.
Gleason, R. A., B. A. Tangen, M. K. Laubhan, K. E. Kermes, and N. H. Euliss. 2007. Estimating water storage
capacity of existing and potentially restorable wetland depressions in a subbasin of the Red River of
the North. USGS Open-File Report 2007-1159, U.S. Department of the Interior, U.S. Geological
Survey, Reston, VA.
Gleeson, T., L. Marklund, L. Smith, and A. H. Manning. 2011. Classifying the water table at regional to
continental scales. Geophysical Research Letters 38:L05401.
Golden, H. E., and C. D. Knightes. 2011. Simulated watershed mercury and nitrate flux responses to
multiple land cover conversion scenarios. Environmental Toxicology and Chemistry 30:773-786.
Golden, H. E., C. D. Knightes, P. A. Conrads, T. D. Feaster, G. M. Davis, S. T. Benedict, and P. M. Bradley.
2013. Climate change and watershed mercury export: A multiple projection and model analysis.
Environmental Toxicology and Chemistry 32:2165-2174.
Golden, H. E., C. R. Lane, D. M. Amatya, K. W. Bandilla, H. Raanan Kiperwas, C. D. Knightes, and H. Ssegane.
2014. Hydrologic connectivity between geographically isolated wetlands and surface water systems:
A review of select modeling methods. Environmental Modelling & Software 53:190-206.
Gomez-Uchida, D., T. W. Knight, and D. E. Ruzzante. 2009. Interaction of landscape and life history
attributes on genetic diversity, neutral divergence and gene flow in a pristine community of
salmonids. Molecular Ecology 18:4854-4869.
Gomi, T., and R. C. Sidle. 2003. Bed load transport in managed steep-gradient headwater streams of
southeastern Alaska. Water Resources Research 39:1336.
Gomi, T., R. C. Sidle, and J. S. Richardson. 2002. Understanding processes and downstream linkages of
headwater systems. BioScience 52:905-916.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Gonod, L. V., J. Chadoeuf, and C. Chenu. 2006. Spatial distribution of microbial 2,4-dichlorophenoxy acetic
acid mineralization from field to microhabitat scales. Soil Science Society of America Journal 70:64-
70.
Gooderham, J. P. R., L. A. Barmuta, and P. E. Davis. 2007. Upstream heterogeneous zones: Small stream
systems structured by a lack of competence? Journal of the North American Benthological Society
26:365-374.
Goodrich, D. C., L. J. Lane, R. M. Shillito, S. N. Miller, K. H. Syed, and D. A. Woolhiser. 1997. Linearity of
basin response as a function of scale in a semiarid watershed. Water Resources Research 33:2951-
2965.
Goodrich, D. C., D. G. Williams, C. L. Unkrich, J. F. Hogan, R. L. Scott, K. R. Hultine, D. R. Pool, A. L. Goes, and
S. Miller. 2004. Comparison of methods to estimate ephemeral channel recharge, Walnut Gulch, San
Pedro River basin, Arizona. Pages 77-99 in Recharge and vadose zone processes: Alluvial basins of
the southwestern United States. F. M. Phillips, J. F. Hogan, and B. Scanlon, editors. American
Geophysical Union, Washington, DC.
Gooseff, M. N., K. E. Bencala, and S. M. Wondzell. 2008. Solute transport along stream and river networks.
Pages 395-417 in River confluences, tributaries, and the fluvial network. S. P. Rice, A. G. Roy, and B.
L. Rhoads, editors. John Wiley & Sons, Chichester, UK.
Gorham, E., J. K. Underwood, F. B. Martin, and J. G. Ogden. 1986. Natural and anthropogenic causes of
lake acidification. Nature 324:451-453.
Gorman, 0. T. 1986. Assemblage organization of stream fishes: The effect of rivers on adventitious
streams. The American Naturalist 128:611-616.
Graf, W. L. 1994. Plutonium and the Rio Grande: Environmental change and contamination in the nuclear
age. Oxford University Press, New York, NY.
Graf, W. L., S. L. Clark, M. T. Kammerer, T. Lehman, K. Randall, and R. Schroeder. 1991. Geomorphology of
heavy metals in the sediments of Queen Creek, Arizona, USA. CATENA 18:567-582.
Granado, D. C., and R. Henry. 2014. Phytoplankton community response to hydrologic variations in
oxbow lakes with different levels of connection to a tropical river. Hydrobiologia 721:223-238.
Grant, E. H. C., W. H. Lowe, and W. F. Pagan. 2007. Living in the branches: Population dynamics and
ecological processes in dendritic networks. Ecology Letters 10:165-175.
Grant, E. H. C., J. D. Nichols, W. H. Lowe, and W. F. Pagan. 2010. Use of multiple dispersal pathways
facilitates amphibian persistence in stream networks. Proceedings of the National Academy of
Sciences of the United States of America 107:6936-6940.
Gray, L. J. 1993. Response of insectivorous birds to emerging aquatic insects in riparian habitats of a
tallgrass prairie stream. American Midland Naturalist 129:288-300.
Greathouse, E. A., C. M. Pringle, W. H. McDowell, and J. G. Holmquist. 2006. Indirect upstream effects of
dams: Consequences of migratory consumer extirpation in Puerto Rico. Ecological Applications
16:339-352.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Green, A.}., and J. Figuerola. 2005. Recent advances in the study of long-distance dispersal of aquatic
invertebrates via birds. Diversity and Distributions 11:149-156.
Green, D. M. 2005. Bufo americanus, American toad. Pages 692-704 in Amphibian declines: The
conservation status of United States species. M. Lannoo, editor. University of California Press,
Berkeley, CA.
Gregory, K. J. 1976. Drainage networks and climate. Pages 289-315 in Geomorphology and climate. E.
Derbyshire, editor. John Wiley & Sons, London, UK.
Gregory, K. J. 2006. The human role in changing river channels. Geomorphology 79:172-191.
Gregory, K. J., and D. E. Walling. 1968. The variation of drainage density within a catchment. Bulletin of
the International Association of Scientific Hydrology 13:61-68.
Gregory, S. V., F. J. Swanson, W. A. McKee, and K. W. Cummings. 1991. An ecosystem perspective of
riparian zones: Focus on links between land and water. BioScience 41:540-551.
Grenouillet, G., D. Pont, and C. Herisse. 2004. Within-basin fish assemblage structure: The relative
influence of habitat versus stream spatial position on local species richness. Canadian Journal of
Fisheries and Aquatic Sciences 61:93-102.
Grigal, D. F. 2002. Inputs and outputs of mercury from terrestrial watersheds: A review. Environmental
Reviews 10:1-39.
Grimshaw, D. L, and J. Lewin. 1980. Source identification for suspended sediments. Journal of Hydrology
47:151-162.
Gupta, V. K., E. Waymire, and C. T. Wang. 1980. A representation of an instantaneous unit hydrograph
from geomorphology. Water Resources Research 16:855-862.
Gurnell, A., K. Thompson, J. Goodson, and H. Moggridge. 2008. Propagule deposition along river margins:
Linking hydrology and ecology. Journal of Ecology 96:553-565.
Gurnell, A. M. 2003. Wood storage and mobility. Pages 75-91 in Ecology and management of wood in
world rivers. S. V. Gregory, K. L. Boyer, and A. M. Gurnell, editors. American Fisheries Society
Symposium 37, Bethesda, MD.
Gurnell, A. M. 2007. Analogies between mineral sediment and vegetative particle dynamics in fluvial
systems. Geomorphology 89:9-22.
Gurnell, A. M., K. J. Gregory, and G. E. Petts. 1995. The role of coarse woody debris in forest aquatic
habitats: Implications for management. Aquatic Conservation: Marine and Freshwater Ecosystems
5:143-166.
Gurnell, A. M., H. Piegay, F. J. Swanson, and S. V. Gregory. 2002. Large wood and fluvial processes.
Freshwater Biology 47:601-619.
Guy, T. J., R. E. Gresswell, and M. A. Banks. 2008. Landscape-scale evaluation of genetic structure among
barrier-isolated populations of coastal cutthroat trout, Oncorhynchus clarkii clarkii. Canadian Journal
of Fisheries and Aquatic Sciences 65:1749-1762.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Haag, K. H., and W. R. Pfeiffer. 2012. Flooded area and plant zonation in isolated wetlands in well fields
in the Northern Tampa Bay Region, Florida, following reductions in groundwater-withdrawal rates.
USGS Scientific Investigations Report 2012-5039. U.S. Department of the Interior, U.S. Geological
Survey, Washington, DC.
Haan, C. T., and H. P. Johnson. 1968. Hydraulic model of runoff from depressional areas. Part I: General
considerations. Transactions of the American Society of Agricultural Engineers 11:364-367.
Haig, S. M., D. W. Mehlman, and L. W. Oring. 1998. Avian movements and wetland connectivity in
landscape conservation. Conservation Biology 12:749-758.
Haitjema, H. M., and S. Mitchell-Bruker. 2005. Are water tables a subdued replica of the topography?
Ground Water 43:781-786.
Hall, B. R., D. J. Raynal, and D. J. Leopold. 2001a. Environmental influences on plant species composition
in ground-water seeps in the Catskill Mountains of New York. Wetlands 21:125-134.
Hall, C. J., A. Jordaan, and M. G. Frisk. 2011. The historic influence of dams on diadromous fish habitat
with a focus on river herring and hydrologic longitudinal connectivity. Landscape Ecology 26:95-
107.
Hall, R. 0., M. A. Baker, C. D. Arp, and B. J. Koch. 2009. Hydrologic control of nitrogen removal, storage,
and export in a mountain stream. Limnology and Oceanography 54:2128-2142.
Hall, R. 0., G. E. Likens, and H. M. Malcom. 2001b. Trophic basis of invertebrate production in 2 streams
at the Hubbard Brook Experimental Forest. Journal of the North American Benthological Society
20:432-447.
Hall, R. 0., and J. L. Meyer. 1998. The trophic significance of bacteria in a detritus-based stream food web.
Ecology 79:1995-2012.
Hall, R. 0., J. B. Wallace, and S. L. Eggert. 2000. Organic matter flow in stream food webs with reduced
detrital resource base. Ecology 81:3445-3463.
Hamilton, S. K., S. E. Bunn, M. C. Thorns, and J. C. Marshall. 2005. Persistence of aquatic refugia between
flow pulses in a dryland river system (Cooper Creek, Australia). Limnology and Oceanography
50:743-754.
Hammersmark, C. T., M. C. Rains, and J. F. Mount. 2008. Quantifying the hydrological effects of stream
restoration in a montane meadow environment. River Research and Applications 24:735-753.
Hanfling, B., and D. Weetman. 2006. Concordant genetic estimators of migration reveal
anthropogenically enhanced source-sink population structure in the river sculpin, Coitus gobio.
Genetics 173:1487-1501.
Hansen, W. F. 2001. Identifying stream types and management implications. Forest Ecology and
Management 143:39-46.
Hanski, I. 1999. Metapopulation ecology. Oxford University Press, Oxford, UK.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Hanson, M. A., K. D. Zimmer, M. G. Butler, B. A. Tangen, B. R. Herwig, and N. H. Euliss. 2005. Biotic
interactions as determinants of ecosystem structure in prairie wetlands: An example using fish.
Wetlands 25:764-775.
Haramis, G. M. 1990. Breeding ecology of the wood duck: A review. Pages 45-60 in Proceedings of the
1988 North American Wood Duck Symposium. L. H. Fredrickson, G. V. Burger, S. P. Havera, D.A.
Graber, R.E. Kirby, and T.S. Taylor, editor., St. Louis, MO.
Harmon, M. E., J. F. Franklin, F. J. Swanson, P. Sollins, S. V. Gregory, J. D. Lattin, N. H. Anderson, S. P. Cline,
N. G. Aumen, J. R. Sedell, G. W. Lienkaemper, K. Cromack, and K. W. Cummins. 1986. Ecology of
coarse woody debris in temperature ecosystems. Advances in Ecological Research 15:133-302.
Harrington, G. A., P. G. Cook, and A. L. Herczeg. 2002. Spatial and temporal variability of ground water
recharge in central Australia: A tracer approach. Ground Water 40:518-528.
Harvey, B. C., R. J. Nakamoto, and J. L. White. 2006. Reduced streamflow lowers dry-season growth of
rainbow trout in a small stream. Transactions of the American Fisheries Society 135:998-1005.
Harvey, J. W., J. D. Drummond, R. L. Martin, L. E. McPhillips, A. I. Packman, D. J. Jerolmack, S. H. Stonedahl,
A. F. Aubeneau, A. H. Sawyer, L. G. Larsen, and C. R. Tobias. 2012. Hydrogeomorphology of the
hyporheic zone: Stream solute and fine particle interactions with a dynamic streambed. Journal of
Geophysical Research: Biogeosciences 117:GOON11.
Harvey, J. W., and C. C. Fuller. 1998. Effect of enhanced manganese oxidation in the hyporheic zone on
basin-scale geochemical mass balance. Water Resources Research 34:623-636.
Hassan, M. A. 1990. Observations of desert flood bores. Earth Surface Processes and Landforms 15:481-
485.
Hassan, M. A., A. S. Gottesfeld, D. R. Montgomery, J. F. Tunnicliffe, G. K. C. Clarke, G. Wynn, H. Jones-Cox, R.
Poirier, E. Maclsaac, H. Herunter, and S. J. Macdonald. 2008. Salmon-driven bed load transport and
bed morphology in mountain streams. Geophysical Research Letters 35:L04405.
Hastings, A., and S. Harrison. 1994. Metapopulation dynamics and genetics. Annual Review of Ecology
and Systematics 25:167-188.
Haukos, D. A., M. R. Miller, D. L. Orthmeyer, J. Y. Takekawa, J. P. Fleskes, M. L. Casazza, W. M. Perry, and J.
A. Moon. 2006. Spring migration of northern pintails from Texas and New Mexico, USA. Waterbirds
29:127-136.
Hawkins, C. P., and J. R. Sedell. 1981. Longitudinal and seasonal-changes in functional-organization of
macroinvertebrate communities in four Oregon streams. Ecology 62:387-397.
Hecnar, S. J., and R. T. McLoskey. 1996. Regional dynamics and the status of amphibians. Ecology
77:2091-2097.
Hedin, L. 0., J. J. Armesto, and A. H. Johnson. 1995. Patterns of nutrient loss from unpolluted old-growth
temperate forests: Evaluation of biogeochemical theory. Ecology 76:493-509.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Heiler, G., T. Hein, F. Schiemer, and G. Bornette. 1995. Hydrological connectivity and flood pulses as the
central aspects of a river-floodplain system. Regulated Rivers Research and Management 11:351-
361.
Heimann, D. C., and M. J. Roell. 2000. Sediment loads and accumulation in a small riparian wetland
system in northern Missouri. Wetlands 20:219-231.
Heine, R. A., C. L. Lant, and R. R. Sengupta. 2004. Development and comparison of approaches for
automated mapping of stream channel networks. Annals of the Association of American
Geographers 94:477-490.
Helfield, J. M., and R. J. Naiman. 2006. Keystone interactions: Salmon and bear in riparian forests of
Alaska. Ecosystems 9:167-180.
Helton, A. M., G. C. Poole, J. L. Meyer, W. M. Wollheim, B. J. Peterson, P. J. Mulholland, E. S. Bernhardt, J. A.
Stanford, C. Arango, L. R. Ashkenas, L. W. Cooper, W. K. Dodds, S. V. Gregory, R. 0. Hall, S. K. Hamilton,
S. L. Johnson, W. H. McDowell, J. D. Potter, J. L. Tank, S. M. Thomas, H. M. Valett, J. R. Webster, and L.
Zeglin. 2011. Thinking outside the channel: Modeling nitrogen cycling in networked river
ecosystems. Frontiers in Ecology and the Environment 9:229-238.
Hemond, H. F. 1980. Biogeochemistry of Thoreau's Bog, Concord, Massachusetts. Ecological Monographs
50:507-526.
Hemond, H. F. 1983. The nitrogen budget of Thoreau's Bog. Ecology 64:99-109.
Hendricks, R. 2005. Siren lacertina Linnaeus, 1766, Greater Siren. Pages 912-914 in Amphibian declines:
The conservation status of United States species. M. Lannoo, editor. University of California Press,
Berkeley, CA.
Henning, J. A., R. E. Gresswell, and I. A. Fleming. 2007. Use of seasonal freshwater wetlands by fishes in a
temperate river floodplain. Journal of Fish Biology 71:476-492.
Hermoso, V., M. J. Kennard, and S. Linke. 2012. Integrating multidirectional connectivity requirements in
systematic conservation planning for freshwater systems. Diversity and Distributions 18:448-458.
Hershey, A. E., J. Pastor, B. J. Peterson, and G. W. Kling. 1993. Stable isotopes resolve the drift paradox for
baetis mayflies in an arctic river. Ecology 74:2315-2325.
Herwig, B. R., K. D. Zimmer, M. A. Hanson, M. L. Konsti, J. A. Younk, R. W. Wright, S. R. Vaughn, and M. D.
Haustein. 2010. Factors influencing fish distributions in shallow lakes in prairie and prairie-
parkland regions of Minnesota, USA. Wetlands 30:609-619.
Hess, G. R. 1996. Linking extinction to connectivity and habitat destruction in metapopulation models.
The American Naturalist 148:226-236.
Hester, E. T., and M. W. Doyle. 2008. In-stream geomorphic structures as drivers of hyporheic exchange.
Water Resources Research 44:W03417.
Hester, E. T., M. W. Doyle, and G. C. Poole. 2009. The influence of in-stream structures on summer water
temperatures via induced hyporheic exchange. Limnology and Oceanography 54:355.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Hester, E. T., and M. N. Gooseff. 2010. Moving beyond the banks: Hyporheic restoration is foundation to
restoring ecological services and functions of streams. Environmental Science & Technology
44:1521-1525.
Hewlett, J. D. 1982. Principles of forest hydrology. University of Georgia Press, Athens, GA.
Hewlett, J. D., G. B. Cunningham, and C. A. Troendle. 1977. Predicting stormflow and peakflow from small
basin in humid areas by the R-index method. Water Resources Bulletin 13:231-253.
Hewlett, J. D., and A. R. Hibbert. 1967. Factors affecting the response of small watersheds to precipitation
in humid areas. Pages 275-290 in International symposium on forest hydrology. W. S. Sopper and H.
W. Hull, editors. Pergamon Press, New York, NY.
Heyes, A., T. R. Moore, J. W. M. Rudd, and J. J. Dugoua. 2000. Methyl mercury in pristine and impounded
boreal peatlands, Experimental Lakes Area, Ontario, Canada. Canadian Journal of Fisheries and
Aquatic Sciences 57:2211-2222.
Hill, A. R., C. F. Labadia, and K. Sanmugadas. 1998. Hyporheic zone hydrology and nitrogen dynamics in
relation to the streambed topography of a N-rich stream. Biogeochemistry 42:285-310.
Hill, A. R., and D. J. Lymburner. 1998. Hyporheic zone chemistry and stream-subsurface exchange in two
groundwater-fed streams. Canadian Journal of Fisheries and Aquatic Sciences 55:495-506.
Hill, W. R., and A. W. Knight. 1988. Nutrient and light limitations of algae in two northern California
streams. Journal of Phycology 24:125-132.
Hitt, N. P., and P. L. Angermeier. 2008. Evidence for fish dispersal from spatial analysis of stream
network topology. Journal of the North American Benthological Society 27:304-320.
Hitt, N. P., C. A. Frissell, C. C. Muhlfeld, and F. W. Allendorf. 2003. Spread of hybridization between native
westslope cutthroat trout, Oncorhynchus clarki lewisi, and nonnative rainbow trout, Oncorhynchus
mykiss. Canadian Journal of Fisheries and Aquatic Sciences 60:1440-1451.
Hladyz, S., S. C. Watkins, K. L. Whitworth, and D. S. Baldwin. 2011. Flows and hypoxic blackwater events
in managed ephemeral river channels. Journal of Hydrology 401:117-125.
Hobbie, J. E., and R. G. Wetzel. 1992. Microbial control of dissolved organic carbon in lakes: Research for
the future. Hydrobiologia 229:169-180.
Hoffmann, C. C., C. Kjaergaard, J. Uusi-Kamppa, H. C. Hansen, and B. Kronvang. 2009. Phosphorus
retention in riparian buffers: Review of their efficiency. Journal of Environmental Quality 38:1942-
1955.
Holloway, J. M., R. A. Dahlgren, B. Hansen, and W. H. Casey. 1998. Contribution of bedrock nitrogen to
high nitrate concentrations in stream water. Nature 395:785-788.
Holmstrup, M., A.-M. Bindesb01, G. J. Oostingh, A. Duschl, V. Scheil, H.-R. Kohler, S. Loureiro, A. M. V. M.
Scares, A. L. G. Ferreira, C. Kienle, A. Gerhardt, R. Laskowski, P. E. Kramarz, M. Bayley, C. Svendsen,
and D. J. Spurgeon. 2010. Interactions between effects of environmental chemicals and natural
stressors: A review. Science of The Total Environment 408:3746-3762.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Hooke, J. 2003. Coarse sediment connectivity in river channel systems: A conceptual framework and
methodology. Geomorphology 56:79-94.
Hope, D., M. F. Billet, and M. S. Cresser. 1994. A review of the export of carbon in river water: Fluxes and
processes. Environmental Pollution 84:301-324.
Hornberger, G. M., J. P. Raffensperger, and P. L. Wilberg. 1998. Elements of physical hydrology. Johns
Hopkins University Press, Baltimore, MD.
Hornberger, M. I., S. N. Luoma, M. L. Johnson, and M. Holyoak. 2009. Influence of remediation in a mine-
impacted river: Metal trends over large spatial and temporal scales. Ecological Applications
19:1522-1535.
Horton, R. E. 1945. Erosional development of streams and their drainage basins: Hydrophysical
approach to quantitative morphology. Bulletin of the Geological Society of America 56:275-370.
Horwitz, R. J. 1978. Temporal variability patterns and the distributional patterns of stream fishes.
Ecological Monographs 48:307-321.
Howarth, R. W., J. R. Fruci, and D. Sherman. 1991. Inputs of sediment and carbon to an estuarine
ecosystem: Influence of land use. Ecological Applications 1:27-39.
Huang, C., Y. Peng, M. Lang, I.-Y. Yeo, and G. McCarty. 2014. Wetland inundation mapping and change
monitoring using Landsat and airborne LiDAR data. Remote Sensing of Environment 141:231-242.
Hubbard, D. E., and R. L. Linder. 1986. Spring runoff retention in prairie pothole wetlands. Journal of Soil
and Water Conservation 41:122-125.
Hudy, M., J. A. Coombs, K. H. Nislow, and B. H. Letcher. 2010. Dispersal and within-stream spatial
population structure of brook trout revealed by pedigree reconstruction analysis. Transactions of
the American Fisheries Society 139:1276-1287.
Huggenberger, P., E. Hoehn, R. Beschta, and W. Woessner. 1998. Abiotic aspects of channels and
floodplains in riparian ecology. Freshwater Biology 40:407-425.
Hughes, D. A., and J. Lewin. 1982. A small-scale flood plain. Sedimentology 29:891-895.
Hughes, D. A., and K. Sami. 1992. Transmission losses to alluvium and associated moisture dynamics in a
semiarid ephemeral channel system in southern Africa. Hydrological Processes 6:45-53.
Hughes, J. M., D. J. Schmidt, and D. S. Finn. 2009. Genes in streams: Using DNA to understand the
movement of freshwater fauna and their riverine habitat. BioScience 59:573-583.
Humphries, P., H. Keckeis, and B. Finlayson. 2015. The river wave concept: Integrating river ecosystem
models. BioScience: doi: 10.1093/biosci/biull30.
Humphries, S., and G. D. Ruxton. 2002. Is there really a drift paradox? Journal of Animal Ecology 71:151-
154.
Hunsinger, T. W., and M. J. Lannoo. 2005. Notophthalmus viridescens, eastern newt. Pages 912-914 in
Amphibian declines: The conservation status of United States species. M. Lannoo, editor. University
of California Press, Berkeley, CA.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Hunt, R.}., M. Strand, and J. F. Walker. 2006. Measuring groundwater-surface water interaction and its
effect on wetland stream benthic productivity, Trout Lake watershed, northern Wisconsin, USA.
Journal of Hydrology 320:370-384.
Hunter, M. A., T. Quinn, and M. P. Hayes. 2005. Low flow spatial characteristics in forested headwater
channels of southwest Washington. Journal of the American Water Resources Association 41:503-
516.
Hupp, C. R. 2000. Hydrology, geomorphology and vegetation of Coastal Plain rivers in the south-eastern
USA. Hydrological Processes 14:2991-3010.
Hupp, C. R., and W. R. Osterkamp. 1996. Riparian vegetation and fluvial geomorphic processes.
Geomorphology 14:277-295.
Hupp, C. R., W. R. Osterkamp, and A. D. Howard. 1995. Biogeomorphology, terrestrial and freshwater
systems. Elsevier Science, Amsterdam, The Netherlands.
Hupp, C. R., A. R. Pierce, and G. B. Noe. 2009. Floodplain geomorphic processes and environmental
impacts of human alteration along Coastal Plain rivers, USA. Wetlands 29:413-429.
Hurley, J. P., J. M. Benoit, C. L. Babiarz, M. M. Shafer, A. W. Andren, J. R. Sullivan, R. Hammond, and D. A.
Webb. 1995. Influences of watershed characteristics on mercury levels in Wisconsin rivers.
Environmental Science & Technology 29:1867-1875.
Huryn, A. D., and K. E. Gibbs. 1999. Riparian sedge meadows in Maine. A macroinvertebrate community
structured by river-floodplain interaction. Pages 363-382 in Invertebrates in freshwater wetlands of
North America: Ecology and management. D. P. Batzer, R. B. Rader, and S. A. Wissinger, editors. John
Wiley & Sons, New York, NY.
Huryn, A. D., K. A. Slavik, R. L. Lowe, S. M. Parker, D. S. Anderson, and B. J. Peterson. 2005. Landscape
heterogeneity and the biodiversity of arctic stream communities: A habitat template analysis.
Canadian Journal of Fisheries and Aquatic Sciences 62:1905-1919.
Ijjasz-Vasquez, E. J., R. L. Bras, and I. Rodriguez-Iturbe. 1993. Hack's relation and optimal channel
networks: The elongation of river basins as a consequence of energy minimization. Geophysical
Research Letters 20:1583-1586.
Ilg, C., F. Dziock, F. Foeckler, K. Follner, M. Gerisch, J. Glaeser, A. Rink, A. Schanowski, M. Scholz, 0.
Deichner, and K. Henle. 2008. Long-term reactions of plants and macroinvertebrates to extreme
floods in floodplain grasslands. Ecology 89:2392-2398.
Ivey, C. T., and J. H. Richards. 2001. Genetic diversity of everglades sawgrass, Cladium jamaicense
(Cyperaceae). International Journal of Plant Sciences 162:817-825.
Izbicki, J. A. 2007. Physical and temporal isolation of mountain headwater streams in the western
Mojave Desert, southern California. Journal of the American Water Resources Association 43:26-40.
Jackson, C. R., and C. M. Pringle. 2010. Ecological benefits of reduced hydrologic connectivity in
intensively developed landscapes. BioScience 60:37-46.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Jacobson, L. M., M. B. David, and L. E. Drinkwater. 2011. A spatial analysis of phosphorus in the
Mississippi river basin. Journal of Environmental Quality 40:931-941.
Jacques, J., and D. L. Lorenz. 1988. Techniques for estimating the magnitude and frequency of floods of
ungauged streams in Minnesota. USGS Water-Resources Investigations Report 84-4170, U.S.
Department of the Interior, U.S. Geological Survey, Washington, DC.
James, A. L., and N. T. Roulet. 2007. Investigating hydrologic connectivity and its association with
threshold change in runoff response in a temperate forested watershed. Hydrological Processes
21:3391-3408.
James, E. R., M. Manga, T. P. Rose, and G. B. Hudson. 2000. The use of temperature and the isotopes of 0,
H, C, and noble gases to determine the pattern and spatial extent of groundwater flow. Journal of
Hydrology 237:100-112.
Jansson, R., U. Zinko, D. M. Merritt, and C. Nilsson. 2005. Hydrochory increases riparian plant species
richness: A comparison between a free-flowing and a regulated river. Journal of Ecology 93:1094-
1103.
Jeffres, C. A., J. J. Opperman, and P. B. Moyle. 2008. Ephemeral floodplain habitats provide best growth
conditions for juvenile Chinook salmon in a California river. Environmental Biology of Fishes
83:449-458.
Jencso, K. G., and B. L. McGlynn. 2011. Hierarchical controls on runoff generation: Topographically
driven hydrologic connectivity, geology, and vegetation. Water Resources Research 47:W11527.
Jencso, K. G., B. L. McGlynn, M. N. Gooseff, S. M. Wondzell, K. E. Bencala, and L. A. Marshall. 2009.
Hydrologic connectivity between landscapes and streams: Transferring reach- and plot-scale
understanding to the catchment scale. Water Resources Research 45:W04428.
Jenkins, K. M., and A. J. Boulton. 2003. Connectivity in a dryland river: Short-term aquatic
microinvertebrate recruitment following floodplain inundation. Ecology 84:2708-2723.
Johnson, A. 1.1967. Specific yield— Compilation of specific yields for various materials. USGS Water-
Supply Paper 1662-D, U.S. Department of the Interior, U.S. Geological Survey, Washington, DC.
Johnson, B. R., A. Haas, and K. M. Fritz. 2010. Use of spatially explicit physicochemical data to measure
downstream impacts of headwater stream disturbance. Water Resources Research 46:W09526.
Johnson, L. B., G. E. Host, J. H. Olker, and C. Richards. 2006. Landscape- and reach-scale predictors of
large wood abundance in low-gradient streams. Pages 151-173 in Landscape influences on stream
habitats and biological assemblages. R. M. Hughes, L. Wang, and P. W. Seelbach, editors. American
Fisheries Society Symposium 48, Bethesda, MD.
Johnson, P. T. J., J. D. Olden, and M. J. Vander Zanden. 2008. Dam invaders: Impoundments facilitate
biological invasions into freshwaters. Frontiers in Ecology and the Environment 6:357-363.
Johnson, S. A., and R. B. Owen. 2005. Amphiuma means Garden, 1821, Two-toed amphiuma. Pages 642-
646 in Amphibian declines: The conservation status of United States species. M. Lannoo, editor.
University of California Press, Berkeley, CA.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Johnson, S. L, and J. A. Jones. 2000. Stream temperature responses to forest harvest and debris flows in
western Cascades, Oregon. Canadian Journal of Fisheries and Aquatic Sciences 57:30-39.
Johnston, C., N. Detenbeck, and G. Niemi. 1990. The cumulative effect of wetlands on stream water
quality and quantity. A landscape approach. Biogeochemistry 10:105-141.
Johnston, C. A. 1991. Sediment and nutrient retention by freshwater wetlands: Effects on surface water
quality. Critical Reviews in Environmental Control 21:491-565.
Johnston, C. A. 1993. Material fluxes across wetland ecotones in Northern landscapes. Ecological
Applications 3:424-440.
Johnston, C. A., and R. J. Naiman. 1990a. Aquatic patch creation in relation to beaver population trends.
Ecology 71:1617-1621.
Johnston, C. A., and R. J. Naiman. 1990b. The use of a geographic information system to analyze long-
term landscape alteration by beaver. Landscape Ecology 4:5-19.
Johnston, C. A., B. A. Shmagin, P. C. Frost, C. Cherrier, J. H. Larson, G. A. Lamberti, and S. D. Bridgham.
2008. Wetland types and wetland maps differ in ability to predict dissolved organic carbon
concentrations in streams. Sci Total Environ 404:326-334.
Jones, J. B., and P. J. Mulholland. 2000. Streams and ground waters. Academic Press, San Diego, CA.
Jordan, T. E., M. P. Andrews, R. P. Szuch, D. F. Whigham, D. E. Weller, and A. D. Jacobs. 2007. Comparing
functional assessments of wetlands to measurements of soil characteristics and nitrogen processing.
Wetlands 27:479-497.
Joyal, L. A., M. McCollough, and M. L. Hunter. 2001. Landscape ecology approaches to wetland species
conservation: A case study of two turtle species in southern Maine. Conservation Biology 15:1755-
1762.
Jude, D. J., and J. Pappas. 1992. Fish utilization of Great Lakes coastal wetlands. Journal of Great Lakes
Research 18:651-672.
Julian, J. T., G. Rocco, M. M. Turner, and R. P. Brooks. 2013. Assessing wetland-riparian amphibian and
reptile communities of the Mid-Atlantic Region. Pages 313-337 in Mid-Atlantic freshwater wetlands:
Advances in wetlands science, management, policy, and practice. R. P. Brooks and D. H. Wardrop,
editors. Springer, New York, NY.
Junk, W. J., P. B. Bayley, and R. E. Sparks. 1989. The flood pulse concept in river-floodplain systems. Pages
110-127 in Proceedings of the international large river symposium, Canadian Special Publication of
Fisheries and Aquatic Sciences 106. D. P. Dodge, editor., Ottawa, Canada.
Kadlec, R. H., and S. D. Wallace. 2009. Treatment wetlands. 2nd edition. CRC Press, Boca Raton, FL.
Kanno, Y., B. H. Letcher, J. A. Coombs, K. H. Nislow, and A. R. Whiteley. 2014. Linking movement and
reproductive history of brook trout to assess habitat connectivity in a heterogeneous stream
network. Freshwater Biology 59:142-154.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Kao, C. M., W. J. Y., K. F. Chen, H. Y. Lee, and M. J. Wu. 2002. Non-point source pesticide removal by a
mountainous wetland. Water Science and Technology 46:199-206.
Kaplan, L. A., and T. L. Bott. 1982. Diel fluctuations of DOC generated by algae in a piedmont stream.
Limnology and Oceanography 27:1091-1100.
Kaplan, L. A., T. N. Wiegner, J. D. Newbold, P. H. Ostrom, and H. Gandhi. 2008. Untangling the complex
issue of dissolved organic carbon uptake: A stable isotope approach. Freshwater Biology 53:855-
864.
Karr, J. R., L. A. Toth, and D. R. Dudley. 1985. Fish communities of Midwestern rivers: A history of
degradation. BioScience 35:90-95.
Karwan, D. L., and J. E. Saiers. 2012. Hyporheic exchange and streambed filtration of suspended particles.
Water Resources Research 48:W01519.
Katano, I., J. N. Negishi, T. Minagawa, H. Doi, Y. Kawaguchi, and Y. Kayaba. 2009. Longitudinal
macroinvertebrate organization over contrasting discontinuities: Effects of a dam and a tributary.
Journal of the North American Benthological Society 28:331-351.
Kaushal, S., and K. Belt 2012. The urban watershed continuum: Evolving spatial and temporal
dimensions. Urban Ecosystems 15:409-435.
Kaushik, N. K, and H. B. N. Hynes. 1971. The fate of dead leaves that fall into streams. Archivfur
Hydrobiologie 68:465-515.
Kawaguchi, Y., and S. Nakano. 2001. Contribution of terrestrial invertebrates to the annual resource
budget for salmonids in forest and grassland reaches of a headwater stream. Freshwater Biology
46:303-316.
Kehew, A. E., R. N. Passero, R. V. Krishnamurthy, C. K. Lovett, M. A. Betts, and B. A. Dayharsh. 1998.
Hydrogeochemical interaction between a wetland and an unconfined glacial drift aquifer,
southwestern Michigan. Ground Water 36:849-856.
Keitt, T., D. L. Urban, and B. T. Milne. 1997. Detecting critical scales in fragmented landscapes. Ecology
and Society 1:4.
Keller, E. A., A. MacDonald, and T. Tally. 1981. Streams in the coastal redwood environment: The role of
large organic debris. Pages 161-176 in Proceedings of a symposium on watershed rehabilitation in
Redwood National Park and other Pacific coastal areas; 1981 August 25-28 in Arcata, CA. R. N. Coats,
editor. Center for Natural Resources Studies of JMI, Inc., Sacramento, CA.
Keller, E. A., and F. J. Swanson. 1979. Effects of large organic material on channel form and fluvial
processes. Earth Surface Processes and Landforms 4:361-380.
Kiffney, P. M., C. M. Greene, J. E. Hall, and J. R. Davis. 2006. Tributary streams create spatial
discontinuities in habitat, biological productivity, and diversity in mainstem rivers. Canadian Journal
of Fisheries and Aquatic Sciences 63:2518-2530.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Kiffney, P. M., J. S. Richardson, and M. C. Feller. 2000. Fluvial and epilithic organic matter dynamics in
headwater streams of southwestern British Columbia, Canada. Archiv fur Hydrobiologie 149:109-
129.
Kimball, B. A., E. Callender, and E. V. Axtmann. 1995. Effects of colloids on metal transport in a river
receiving acid mine drainage, Upper Arkansas River, Colorado, USA. Applied Geochemistry 10:285-
306.
King, A. J., P. Humphries, and P. S. Lake. 2003. Fish recruitment on floodplains: The roles of patterns of
flooding and life history characteristics. Canadian Journal of Fisheries and Aquatic Sciences 60:773-
786.
Kish, G. R., C. E. Stringer, M. T. Stewart, M. C. Rains, and A. E. Torres. 2010. A geochemical mass-balance
method for base-flow separation, upper Hillsborough River watershed, west-central Florida, 2003-
2005 and 2009. USGS Scientific Investigations Report 2010-5092, U.S. Department of the Interior,
U.S. Geological Survey, Washington, DC.
Knight, K. W., R. C. Schultz, C. M. Mabry, and T. M. Isenhart. 2010. Ability of remnant riparian forests,
with and without grass filters, to buffer concentrated surface runoff. Journal of the American Water
Resources Association 46:311-322.
Knightes, C. D., E. M. Sunderland, M. Craig Barber, J. M. Johnston, and R. B. Ambrose. 2009. Application of
ecosystem-scale fate and bioaccumulation models to predict fish mercury response times to changes
in atmospheric deposition. Environmental Toxicology and Chemistry 28:881-893.
Knighton, A. D. 1980. Longitudinal changes in size and sorting of stream-bed material in four English
rivers. Geological Society of America Bulletin 91:55-62.
Knighton, A. D. 1998. Fluvial forms and processes: A new perspective. Hodder Arnold, London, UK.
Knispel, S., and E. Castella. 2003. Disruption of a longitudinal pattern in environmental factors and
benthic fauna by a glacial tributary. Freshwater Biology 48:604-618.
Knudby, C., and J. Carrera. 2005. On the relationship between indicators of geostatistical, flow and
transport connectivity. Advances in Water Resources 28:405-421.
Knutson, M. G., J. R. Sauer, D. A. Olsen, M. J. Mossman, L. M. Hemesath, and M. J. Lannoo. 1999. Effects of
landscape composition and wetland fragmentation on frog and toad abundance and species richness
in Iowa and Wisconsin, U.S.A. Conservation Biology 13:1437-1446.
Kondolf, G. M. 1997. Hungry water: Effects of dams and gravel mining on river channels. Environmental
Management 21:533-551.
Kondolf, G. M., A. J. Boulton, S. 0. O'Daniel, G. C. Poole, F. J. Rahel, E. H. Stanley, E. Wohl, A. Bang, J.
Carlston, C. Cristoni, H. Huber, S. Koljonen, P. Louhi, and K. Nakamura. 2006. Process-based
ecological river restoration: Visualizing three-dimensional connectivity and dynamic vectors to
recover lost linkages. Ecology and Society 11:5.
Koprivnjak, J.-F., and T. R. Moore. 1992. Sources, sinks, and fluxes of dissolved organic carbon in
subarctic fen catchments. Arctic and Alpine Research 24:204-210.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Kortelainen, P. 1993. Content of total organic carbon in Finnish lakes and its relationship to catchment
characteristics. Canadian Journal of Fisheries and Aquatic Sciences 50:1477-1483.
Krapu, G. L, D. E. Facey, E. K. Fritzell, and D. H. Johnson. 1984. Habitat use by migrant Sandhill cranes in
Nebraska. The Journal of Wildlife Management 48:407-417.
Kriimmel, E., R. Macdonald, L. Kimpe, I. Gregory-Eaves, M. Demers, J. Smol, B. Finney, and J. Blais. 2003.
Aquatic ecology: Delivery of pollutants by spawning salmon. Nature 425:255-256.
Krutz, L. J., T. J. Gentry, S. A. Senseman, I. L. Pepper, and D. P. Tierney. 2006. Mineralization of atrazine,
metolachlor and their respective metabolites in vegetated filter strips and cultivated soil. Pest
Management Science 62:505-514.
Labbe, T. R., and K. D. Fausch. 2000. Dynamics of intermittent stream habitat regulate persistence of a
threatened fish at multiple scales. Ecological Applications 10:1774-1791.
LaGrange, T. G., and J. J. Dinsmore. 1989. Habitat use by mallards during spring migration through
central Iowa. Journal of Wildlife Management 53:1076-1081.
Lake, P. S. 2000. Disturbance, patchiness, and diversity in streams. Journal of the North American
Benthological Society 19:573-592.
Lamberti, G. A., and V. H. Resh. 1987. Seasonal patterns of suspended bacteria and algae in 2 northern
California streams. Archivfur Hydrobiologie 110:45-57.
Lamoureux, V. S., and D. M. Madison. 1999. Overwintering habitats of radio-implanted green frogs, Rana
clamitans. Journal of Herpetology 33:430-435.
Lampkin, A. J., and M. R. Sommerfeld. 1986. Impact of acid mine-drainage from abandoned spoils on the
chemistry of an intermittent-stream in the arid southwest. Hydrobiologia 139:135-142.
Lancaster, S. T., S. K. Hayes, and G. E. Grant. 2003. Effects of wood on debris flow runout in small
mountain watersheds. Water Resources Research 39:1168.
Lande, R., and S. Shannon. 1996. The role of genetic variation in adaptation and population persistence
in a changing environment. Evolution 50:434-437.
Lane, C. R., and E. D'Amico. 2010. Calculating the ecosystem service of water storage in isolated wetlands
using LiDAR in north central Florida, USA. Wetlands 30:967-977.
Lane, E. W. 1955. The importance of fluvial morphology in hydraulic engineering. Journal of the
Hydraulics Division American Society of Civil Engineers 81:1-17.
Lane, S. N., S. M. Reaney, and A. L. Heathwaite. 2009. Representation of landscape hydrological
connectivity using a topographically driven surface flow index. Water Resources Research
45:W08423.
Lang, M., 0. McDonough, G. McCarty, R. Oesterling, and B. Wilen. 2012. Enhanced detection of wetland-
stream connectivity using LiDAR. Wetlands 32:461-473.
Lange, J. 2005. Dynamics of transmission losses in a large arid stream channel. Journal of Hydrology
306:112-126.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Langston, M. A., and D. M. Kent. 1997. Fish recruitment to a constructed wetland. Journal of Freshwater
Ecology 12:123-129.
Larned, S. T., D. B. Arscott, J. Schmidt, and J. C. Diettrich. 2010. A framework of analyzing longitudinal and
temporal variation in river flow and developing flow-ecology relationships. Journal of the American
Water Resources Association 46:541-553.
Larned, S. T., D. M. Hicks, J. Schmidt, A. J. H. Davey, K. Dey, M. Scarsbrook, D. B. Arscott, and R. A. Woods.
2008. The Selwyn River of New Zealand: A benchmark system for alluvial plain rivers. River
Research and Applications 24:1-21.
Laronne, J. B., and I. Reid. 1993. Very high rates of bedload sediment transport by ephemeral desert
rivers. Nature 366:148-150.
Larsen, L. G., J. Choi, M. K. Nungesser, and J. W. Harvey. 2012. Directional connectivity in hydrology and
ecology. Ecological Applications 22:2204-2220.
Lautz, L. K., and R. M. Fanelli. 2008. Seasonal biogeochemical hotspots in the streambed around
restoration structures. Biogeochemistry 91:85-104.
Lavery, H. J. 1972. The grey teal at saline drought-refuges in north Queensland. Wildfowl 23:56-63.
Lawrence, J. E., M. E. Skold, F. A. Hussain, D. R. Silverman, V. H. Resh, D. L. Sedlak, R. G. Luthy, and J. E.
McCray. 2013. Hyporheic zone in urban streams: A review and opportunities for enhancing water
quality and improving aquatic habitat by active management. Environmental Engineering Science
30:480-501.
Le, P. V. V., and P. Kumar. 2014. Power law scaling of topographic depressions and their hydrologic
connectivity. Geophysical Research Letters 41:1553-1559.
LeBlanc, R. T., R. D. Brown, and J. E. FitzGibbon. 1997. Modeling the Effects of Land Use Change on the
Water Temperature in Unregulated Urban Streams. Journal of Environmental Management 49:445-
469.
Lecerf, A., D. Patfield, A. Boiche, M. P. Riipinen, E. Chauvet, and M. Dobson. 2007. Stream ecosystems
respond to riparian invasion by Japanese knotweed [Fallopia japonicd). Canadian Journal of
Fisheries and Aquatic Sciences 64:1273-1283.
Lee, M. A. B., K. L. Snyder, P. Valentine-Darby, S. J. Miller, and K. J. Ponzio. 2005. Dormant season
prescribed fire as a management tool for the control ofSalix caroliniana Michx. in a floodplain
marsh. Wetlands Ecology and Management 13:479-487.
Lehman, P. W., T. Sommer, and L. Rivard. 2008. The influence of floodplain habitat on the quantity and
quality of riverine phytoplankton carbon produced during the flood season in San Francisco Estuary.
Aquatic Ecology 42:363-378.
Leibowitz, S. G. 2003. Isolated wetlands and their functions: An ecological perspective. Wetlands
23:517-531.
Leibowitz, S. G., and K. C. Vining. 2003. Temporal connectivity in a prairie pothole complex. Wetlands
23:13-25.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Leibowitz, S. G., P. J. Wigington, M. C. Rains, and D. M. Downing. 2008. Non-navigable streams and
adjacent wetlands: Addressing science needs following the Supreme Court's Rapanos decision.
Frontiers in Ecology and the Environment 6:364-371.
Leigh, C., F. Sheldon, R. T. Kingsford, and A. H. Arthington. 2010. Sequential floods drive 'booms' and
wetland persistence in dryland rivers: A synthesis. Marine and Freshwater Research 61:896-908.
Leopold, L. B. 1968. Hydrology for urban planning. A guidebook on the hydrologic effects of urban land
use. USGS Circular 554, U.S. Department of the Interior, U.S. Geological Survey, Washington, DC.
Leopold, L. B. 1994. A view of the river. Harvard University Press, Cambridge, MA.
Leopold, L. B., and T. Maddock. 1953. The hydraulic geometry of stream channels and some
physiographic implications. USGS Professional Paper 252, U.S. Department of the Interior, U.S.
Geological Service, Washington, DC.
Leopold, L. B., M. G. Wolman, and J. P. Miller. 1964. Fluvial processes in geomorphology. W.H. Freeman
and Co., San Francisco, CA.
Lerner, D. N. 1986. Leaking pipes recharge ground water. Ground Water 24:654-662.
Letcher, B. H., K. H. Nislow, J. A. Coombs, M. J. O'Donnell, and T. L. Dubreuil. 2007. Population response to
habitat fragmentation in a stream-dwelling brook trout population. PLOS ONE 2:ell39.
Levick, L., J. Fonseca, D. Goodrich, M. Hernandez, D. Semmens, R. Leidy, M. Scianni, P. Guertin, M. Tluczek,
and W. Kepner. 2008. The ecological and hydrological significance of ephemeral and intermittent
streams in the arid and semi-arid American Southwest. EPA/600/R-08/134 and ARS/233046, U.S.
Environmental Protection Agency, Office of Research and Development and U.S. Department of
Agriculture/Agricultural Research Service. Southwest Watershed Research Center, Washington, DC.
Lewis, M. A. 1977. Aquatic inhabitants of a mine waste stream in Arizona. Research Note RM-349, U.S.
Department of Agriculture, U.S. Forest Service, Fort Collins, CO.
Lewis, M. A., and R. Burraychak. 1979. Impact of copper mining on a desert intermittent stream. Journal
of Arizona-Nevada Academy of Science 14:22-29.
Lexartza-Artza, L, and J. Wainwright. 2009. Hydrological connectivity: Linking concepts with practical
implications. CATENA 79:146-152.
Lide, R. F., V. G. Meentemeyer, J. E. Pinder, and L. M. Beatty. 1995. Hydrology of a Carolina bay located on
the upper Coastal Plain of western South Carolina. Wetlands 15:47-57.
Ligon, F. K., W. E. Dietrich, and W. J. Trush. 1995. Downstream ecological effects of dams. BioScience
45:183-192.
Limm, M. P., and M. P. Marchetti. 2009. Juvenile Chinook salmon (Oncorhynchus tshawytscha) growth in
off-channel and main-channel habitats on the Sacramento River, CA using otolith increment widths.
Environmental Biology of Fishes 85:141-151.
Lindsay, J. B., I. F. Creed, and F. D. Beall. 2004. Drainage basin morphometrics for depressional
landscapes. Water Resources Research 4:W09307.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Linqvist, 0., K. Johansson, M. Aastrup, A. Andersson, L. Bringmark, G. Hovsenius, L. Hakanson, A.
Iverfeldt, M. Meili, and B. Timm. 1991. Mercury in the Swedish environment— recent research on
causes, consequences, and remedial measures. Water Air and Soil Pollution 55:xi-xiii.
Lloyd, 0. B., and W. L. Lyke. 1995. Ground water atlas of the United States, segment 10. USGS Hydrologic
Investigations Atlas 730-K, U.S. Department of the Interior, U.S. Geological Survey, Reston, VA.
Loheide, S. P., and S. M. Gorelick. 2006. Quantifying stream-aquifer interactions through the analysis of
remotely sensed thermographic profiles and in situ temperature histories. Environmental Science &
Technology 40:3336-3341.
Lorenz, D. L., C. A. Sanocki, and M. J. Kocian. 2010. Techniques for estimating the magnitude and
frequency of peak flows on small streams in Minnesota based on through water year 2005. USGS
Scientific Investigations Report 2009-5250, U.S. Department of the Interior, U.S. Geological Survey,
in cooperation with the Minnesota Department of Transportation and the Minnesota Pollution
Control Agency, Reston, VA.
Lowrance, R., L. S. Altier, J. D. Newbold, R. R. Schnabel, P. M. Groffman, J. M. Denver, D. L. Correll, J. W.
Gilliam, J. L. Robinson, R. B. Brinsfield, K. W. Staver, W. Lucas, and A. H. Todd. 1997. Water quality
functions of riparian forest buffers in Chesapeake Bay watersheds. Environmental Management
21:687-712.
Luttrell, G. R., A. A. Echelle, W. L. Fisher, and D. J. Eisenhour. 1999. Declining status of two species of the
Macrhybopsis aestivalis complex (Teleostei: Cyprinidae) in the Arkansas River basin and related
effects of reservoirs as barriers to dispersal. Copeia 1999:981-989.
Lytle, D. A., and N. L. Poff. 2004. Adaptation to natural flow regimes. Trends in Ecology and Evolution
19:94-100
MacDonald, L. H., and D. Coe. 2007. Influence of headwater streams on downstream reaches in forested
areas. Forest Science 53:148-168.
Magana, H. A. 2013. Flood pulse trophic dynamics of larval fishes in a restored arid-land, river
floodplain, middle Rio Grande, Los Lunas, New Mexico. Reviews in Fish Biology and Fisheries
23:507-521.
Magette, W. L., R. B. Brinsfield, R. E. Palmer, and J. D. Wood. 1989. Nutrient and sediment removal by
vegetated filter strips. Transactions of the American Society of Agricultural Engineers 32:663-667.
Malard, F., K. Tockner, M. Dole-Olivier, and J. V. Ward. 2002. A landscape perspective on surface-
subsurface hydrological exchanges in river corridors. Freshwater Biology 47:621-640.
Malard, F., K. Tockner, and J. V. Ward. 1999. Shifting dominance of subcatchment water sources and flow
paths in a glacial floodplain, Val Roseg, Switzerland. Arctic, Antarctic, and Alpine Research 31:135-
150.
Malard, F., U. Uehlinger, R. Zah, and K. Tockner. 2006. Flood-pulse and riverscape dynamics in a braided
glacier river. Ecology 87:704-716.
Malmqvist, B. 2002. Aquatic invertebrates in riverine landscapes. Freshwater Biology 47:679-694.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Malvadkar, U., F. Scatena, and M. Leon. 2015. A comparison of connectivity metrics on watersheds and
implications for water management. River Research and Applications: doi: 10.1002/rra.2730.
Marcus, W. A. 1987. Copper dispersion in ephemeral stream sediments. Earth Surface Processes and
Landforms 12:217-228.
Marker, A. F. H., and R. J. M. Gunn. 1977. Benthic algae of some streams in southern England. III. Seasonal
variations in chlorophyll a in the seston. Journal of Ecology 65:223-234.
Martin, S. L, and P. A. Soranno. 2006. Lake landscape position: Relationships to hydrologic connectivity
and landscape features. Limnology and Oceanography 51:801-814.
Marton, J. M., M. S. Fennessy, and C. B. Craft. 2014. USDA conservation practices increase carbon storage
and water quality improvement functions: An example from Ohio. Restoration Ecology 22:117-124.
Matheney, M. P., and C. F. Rabeni. 1995. Patterns of movement and habitat use by northern hogsuckers
in an Ozark stream. Transactions of the American Fisheries Society 124:886-897.
Matheney, R. K., and P. J. Gerla. 1996. Environmental isotopic evidence for the origins of ground and
surface water in a prairie discharge wetland. Wetlands 16:109-120.
Matthews, W. J., and E. Marsh-Matthews. 2007. Extirpation of red shiner in direct tributaries of Lake
Texoma (Oklahoma-Texas): A cautionary case history from a fragmented river-reservoir system.
Transactions of the American Fisheries Society 136:1041-1062.
Matthews, W. J., and H. W. Robinson. 1998. Influence of drainage connectivity, drainage area and
regional species richness on fishes of the interior highlands in Arkansas. American Midland
Naturalist 139:1-19.
May, C. L, and R. E. Gresswell. 2003. Processes and rates of sediment and wood accumulation in
headwater streams of the Oregon Coast Range, USA. Earth Surface Processes and Landforms 28:409-
424.
May, C. L., and D. C. Lee. 2004. The relationship among in-channel sediment storage, pool depth, and
summer survival of juvenile salmonids in Oregon Coast Range streams. North American Journal of
Fisheries Management 24:761-774.
Mayer, P. M., S. K. Reynolds, M. D. McCutchen, and T. J. Canfield. 2007. Meta-analysis of nitrogen removal
in riparian buffers. Journal of Environmental Quality 36:1172-1180.
Mayer, T. D., and S. W. Naman. 2011. Streamflow response to climate as influenced by geology and
elevation. Journal of the American Water Resources Association 47:724-738.
McCauley, L. A., and M. J. Anteau. 2014. Generating nested wetland catchments with readily-available
digital elevation data may improve evaluations of land-use change on wetlands. Wetlands 34:1123-
1132.
McClain, M. E., E. W. Boyer, C. L. Dent, S. E. Gergel, N. B. Grimm, P. M. Groffman, S. C. Hart, J. W. Harvey, C.
A. Johnston, E. Mayorga, W. H. McDowell, and G. Pinay. 2003. Biogeochemical hot spots and hot
moments at the interface of terrestrial and aquatic ecosystems. Ecosystems 6:301-312.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
McDonough, 0. T., M. W. Lang, J. D. Hosen, and M. A. Palmer. 2015. Surface hydrologic connectivity
between Delmarva bay wetlands and nearby streams along a gradient of agricultural alteration.
Wetlands: doi 10.1007/sl3157-13014-10591-13155.
McEachern, P., E. E. Prepas, and D. S. Chanasyk. 2006. Landscape control of water chemistry in northern
boreal streams of Alberta. Journal of Hydrology 323:303-324.
McGlynn, B. L, and J. J. McDonnell. 2003. Quantifying the relative contributions of riparian and hillslope
zones to catchment runoff. Water Resources Research 39:1310.
McGlynn, B. L., J. J. McDonnell, J. Seibert, and C. Kendall. 2004. Scale effects on headwater catchment
runoff timing, flow sources, and groundwater-streamflow relations. Water Resources Research
40:W07504.
McGuire, K. J., J. J. McDonnell, M. Weiler, C. Kendall, B. L. McGlynn, J. M. Welker, and J. Seibert. 2005. The
role of topography on catchment-scale water residence time. Water Resources Research
41:W05002.
McLaughlin, D. L., and M. J. Cohen. 2013. Realizing ecosystem services: Wetland hydrologic function
along a gradient of ecosystem condition. Ecological Applications 23:1619-1631.
McLaughlin, D. L., D. A. Kaplan, and M. J. Cohen. 2014. A significant nexus: Geographically isolated
wetlands influence landscape hydrology. Water Resources Research 50:7153-7166.
McLaughlin, J. W., E. B. W. Calhoon, M. R. Gale, M. F. Jurgensen, and C. C. Trettin. 2011. Biogeochemical
cycling and chemical fluxes in a managed northern forested wetland, Michigan, USA. Forest Ecology
and Management 261:649-661.
Meador, M. R. 1996. Water transfer projects and the role of fisheries biologists. Fisheries 21:18-23.
Meixner, T. A., A. K. Huth, P. D. Brooks, M. H. Conklin, N. B. Grimm, R. C. Bales, P. A. Haas, and J. R. Petti.
2007. Influence of shifting flow paths on nitrogen concentrations during monsoon floods, San Pedro
River, Arizona. Journal of Geophysical Research: Biogeosciences 112:G03S03.
Merriam, J. L., W. H. McDowell, J. L. Tank, W. M. Wollheim, C. L. Crenshaw, and S. L. Johnson. 2002.
Characterizing nitrogen dynamics, retention and transport in a tropical rainforest stream using an in
situ 15N addition. Freshwater Biology 47:143-160.
Mertes, L. A. K. 1997. Documentation and significance of the perirheic zone. Water Resources Research
33:1749-1762.
Mertes, L. A. K., D. L. Daniel, J. M. Melack, B. Nelson, A. Martinelli, and B. R. Forsberg. 1995. Spatial
patterns of hydrology, geomorphology, and vegetation on the floodplain of the Amazon River in
Brazil from a remote sensing perspective. Geomorphology 13:215-232.
Meyboom, P. 1964. Three observations on streamflow depletion by phreatophytes. Journal of Hydrology
2:248-261.
Meyer, A., N. Kaschek, and E. I. Meyer. 2004. The effect of low flow and stream drying on the distribution
and relative abundance of the alien amphipod, Echinogammarus berilloni (Catta, 1878) in a karstic
stream system (Westphalia, Germany). Crustaceana 77:909-922.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Meyer,]. L. 1994. The microbial loop in flowing waters. Microbial Ecology 28:195-199.
Meyer, J. L., and G. E. Likens. 1979. Transport and transformation of phosphorus in a forest stream
ecosystem. Ecology 60:1255-1269.
Meyer, J. L., D. L. Strayer, J. B. Wallace, S. L. Eggert, G. S. Helfman, and N. E. Leonard. 2007. The
contribution of headwater streams to biodiversity in river networks. Journal of the American Water
Resources Association 43:86-103.
Meyer, J. L., and J. B. Wallace. 2001. Lost linkages and lotic ecology: Rediscovering small streams. Pages
295-317 in Ecology: Achievement and challenge. M. C. Press, N. J. Huntly, and S. Levin, editors.
Blackwell Science, Oxford, UK.
Meyer, J. L., J. B. Wallace, and S. L. Eggert. 1998. Leaf litter as a source of dissolved organic carbon in
streams. Ecosystems 1:240-249.
Meyer, L. D., S. M. Dabney, and W. C. Harmon. 1995. Sediment-trapping effectiveness of stiff-grass
hedges. Transactions of the American Society of Agricultural Engineers 38:809-815.
Michalzik, B., K. Kalbitz, J. H. Park, S. Solinger, and E. Matzner. 2001. Fluxes and concentrations of
dissolved organic carbon and nitrogen - a synthesis for temperate forests. Biogeochemistry 52:173-
205.
Middleton, B. 2000. Hydrochory, seed banks, and regeneration dynamics along the landscape
boundaries of a forested wetland. Plant Ecology 146:169-184.
Middleton, S., and D. Liittschwager. 1994. Witness: Endangered species of North America. Chronicle
Books, San Francisco, CA.
Mierle, G., and R. Ingram. 1991. The role of humic substances in the mobilization of mercury from
watersheds. Water Air and Soil Pollution 56:349-357.
Miller, E. L., R. S. Beasley, and E. R. Lawson. 1988. Forest harvest and site preparation effects on
stoonflow and peakflow of ephemeral streams in the Ouachita Mountains. Journal of Environmental
Quality 17:212-218.
Min, J.-H., D. Perkins, and J. Jawitz. 2010. Wetland-groundwater interactions in subtropical depressional
wetlands. Wetlands 30:997-1006.
Minshall, G. W. 1967. Role of allochthonous detritus in the tropic structure of a woodland springbrook
community. Ecology 48:139-149.
Minshall, G. W., R. C. Petersen, T. L. Bott, C. E. Gushing, K. W. Cummins, R. L. Vannote, and J. R. Sedell.
1992. Stream ecosystem dynamics of the Salmon River, Idaho: An 8th-order system. Journal of the
North American Benthological Society 11:111-137.
Minshall, G. W., R. C. Petersen, K. W. Cummins, T. L. Bott, J. R. Sedell, C. E. Gushing, and R. L. Vannote.
1983. Interbiome comparison of stream ecosystem dynamics. Ecological Monographs 53:1-25.
Mitchell, C. A., T. W. Custer, and P. J. Zwank. 1992. Redhead duck behavior on lower Laguna Madre and
adjacent ponds of southern Texas. The Southwestern Naturalist 37:65-72.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Mitchell, C. P.}., P. W. C. Paton, and C. J. Raithel. 2007. The importance of vernal pools to reptiles, birds
and mammals. Pages 169-192 in Science and conservation of vernal pools in northeastern North
America. A. J. K. Calhoun and P. G. deMaynadier, editors. CRC Press, Boca Raton, FL.
Mitsch, W. J., J. K. Cronk, X. Wu, R. W. Nairn, and D. L. Hey. 1995. Phosphorus retention in constructed
freshwater riparian marshes. Ecological Applications 5:830-845.
Mitsch, W. J., and J. G. Gosselink. 2007. Wetlands. 4th edition. John Wiley & Sons Inc., Hoboken, NJ.
Mitsch, W. J., J. G. Gosselink, C. Anderson, J., and L. Zhang. 2009. Wetland ecosystems. 1st edition. John
Wiley & Sons, Hoboken, NJ.
Miyazono, S., J. N. Aycock, L. E. Miranda, and T. E. Tietjen. 2010. Assemblage patterns offish functional
groups relative to habitat connectivity and conditions in floodplain lakes. Ecology of Freshwater Fish
19:578-585.
Mock, S. J. 1971. A classification channel links in stream networks. Water Resources Research 7:1558-
1566.
Moilanen, A., and I. Hanski. 1998. Metapopulation dynamics: Effects of habitat quality and landscape
structure. Ecology 79:2503-2515.
Moilanen, A., and M. Nieminen. 2002. Simple connectivity measures in spatial ecology. Ecology 83:1131-
1145.
Moll, D. 1990. Population sizes and foraging ecology in a tropical freshwater stream turtle community.
Journal of Herpetology 24:48-53.
Montgomery, D. R. 1994. Road surface drainage, channel initiation, and slope instability. Water
Resources Research 30:1925-1932.
Montgomery, D. R. 1999. Process domains and the river continuum. Journal of the American Water
Resources Association 35:397-410.
Montgomery, D. R., B. D. Collins, J. M. Buffington, and T. B. Abbe. 2003. Geomorphic effects of wood in
rivers. Pages 21-47 in Ecology and management of wood in world rivers. S. V. Gregory, K. L. Boyer,
and A. M. Gurnell, editors. American Fisheries Society Symposium 37, Bethesda, MD.
Moore, R. D., D. L. Spittlehouse, and A. Story. 2005. Riparian microclimate and stream temperature
response to forest harvesting: A review. Journal of the American Water Resources Association
41:813-834.
Moraghan, J. T. 1993. Loss and assimilation of 15N-nitrate added to a North Dakota cattail marsh.
Aquatic Botany 46:225-234.
Moreno-Mateos, D., U. Mander, F. A. Comin, C. Pedrocchi, and E. Uuemaa. 2008. Relationships between
landscape pattern, wetland characteristics, and water quality in agricultural catchments. Journal of
Environmental Quality 37:2170-2180.
Morisawa, M. 1957. Accuracy of determination of stream lengths from topographic maps. Transactions
of the American Geophysical Union 38:86-88.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Morita, K., and S. Yamamoto. 2002. Effects of habitat fragmentation by damming on the persistence of
stream-dwelling charr populations. Conservation Biology 16:1318-1323.
Morley, T. R., A. S. Reeve, and A. J. K. Calhoun. 2011. The role of headwater wetlands in altering
streamflow and chemistry in a Maine, USA catchment. Journal of the American Water Resources
Association 47:337-349.
Morrissey, C. A., I. L. Pollet, S. J. Ormerod, and J. E. Elliott. 2011. American dippers indicate contaminant
biotransport by Pacific salmon. Environmental Science & Technology 46:1153-1162.
Morrissey, M. B., and D. T. de Kerckhove. 2009. The maintenance of genetic variation due to asymmetric
gene flow in dendritic metapopulations. The American Naturalist 174:875-889.
Mudd, E., J. Entry, R. Lowrance, and H. Walker. 1995. The influence of management of riparian areas on
movement of C-14 labelled atrazine through undisturbed soil cores. Pages 201-210 in Versatility of
wetlands in the agricultural landscape. K. L. Campbell, editor. American Society of Agricultural
Engineers, St. Joseph, MI.
Mulholland, P. J., A. M. Helton, G. C. Poole, R. 0. Hall, S. K. Hamilton, B. J. Peterson, J. L. Tank, L. R.
Ashkenas, L. W. Cooper, C. N. Dahm, W. K. Dodds, S. E. G. Findlay, S. V. Gregory, N. B. Grimm, S. L.
Johnson, W. H. McDowell, J. L. Meyer, H. M. Valett, J. R. Webster, C. P. Arango, J. J. Beaulieu, M. J.
Bernot, A. J. Burgin, C. L. Crenshaw, L. T. Johnson, B. R. Niederlehner, J. M. O'Brien, J. D. Potter, R. W.
Sheibley, D. J. Sobota, and S. M. Thomas. 2008. Stream denitrification across biomes and its response
to anthropogenic nitrate loading. Nature 452:202-205.
Mulholland, P. J., and W. R. Hill. 1997. Seasonal patterns in stream water nutrient and dissolved organic
carbon concentrations: Separating catchment flow path and in-stream effects. Water Resources
Research 33:1297-1306.
Mulholland, P. J., and E. J. Kuenzler. 1979. Organic carbon export from upland and forested wetland
watersheds. Limnology and Oceanography 24:960-966.
Mulholland, P. J., E. R. Marzolf, S. P. Hendricks, R. V. Wilkerson, and A. K. Baybayan. 1995. Longitudinal
patterns of nutrient cycling and periphyton characteristics in streams: A test of upstream-
downstream linkage. Journal of the North American Benthological Society 14:357-370.
Mulholland, P. J., R. A. Minear, and J. W. Elwood. 1988. Production of soluble, high molecular weight
phosphorus and its subsequent uptake by stream detritus. Verhandlungen der Internationale
Vereinigungfur Theoretische und Angewandte Limnologie 23:1190-1197.
Mulholland, P. J., H. M. Valett, J. R. Webster, S. A. Thomas, L. W. Cooper, S. K. Hamilton, and B. J. Peterson.
2004. Stream denitrification and total nitrate uptake rates measured using a field15N tracer addition
approach. Limnology and Oceanography 49:809-820.
Mulhouse, J. M., and S. M. Galatowitsch. 2003. Revegetation of prairie pothole wetlands in the mid-
continental US: Twelve years post-reflooding. Plant Ecology 169:143-159.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Mullen, L. B., H. Arthur Woods, M. K. Schwartz, A. J. Sepulveda, and W. H. Lowe. 2010. Scale-dependent
genetic structure of the Idaho giant salamander [Dicamptodon aterrimus) in stream networks.
Molecular Ecology 19:898-909.
Miiller, K. 1982. The colonization cycle of insects. Oecologia 53:202-207.
Murkin, H. R., and P. J. Caldwell. 2000. Avian use of prairie wetlands. Pages 249-286 in Prairie wetland
ecology: The contribution of the marsh ecology research program. H. R. Murkin, A. G. van der Valk,
and W. R. Clark, editors. Iowa State University Press, Ames, IA.
Nadeau, T. L., and M. C. Rains. 2007. Hydrological connectivity between headwater streams and
downstream waters: How science can inform policy. Journal of the American Water Resources
Association 43:118-133.
Naiman, R. J. 1982. Characteristics of sediment and organic carbon export from pristine boreal forest
watersheds. Canadian Journal of Fisheries and Aquatic Sciences 39:1699-1718.
Naiman, R. J., and R. E. Bilby. 1998. River ecology and management: Lessons from the Pacific coastal
ecoregion Springer-Verlag, New York, NY.
Naiman, R. J., and H. Decamps. 1997. The ecology of interfaces: Riparian zones. Annual Review of Ecology
and Systematics 28:621-658.
Naiman, R. J., H. Decamps, and M. E. McClain. 2005. Riparia: Ecology, conservation, and management of
streamside communities. Elsevier Academic Press, Burlington, MA.
Naiman, R. J., C. A. Johnston, and J. C. Kelley. 1988. Alteration of North-American streams by beaver.
BioScience 38:753-762.
Naiman, R. J., G. Pinay, C. A. Johnston, and J. Pastor. 1994. Beaver influences on the long-term
biogeochemical characteristics of boreal forest drainage networks. Ecology 75:905-921.
Naiman, R. J., and J. R. Sedell. 1979. Characterization of particulate organic matter transported by some
Cascade Mountain streams. Journal of the Fisheries Research Board of Canada 36:17-31.
Nakamura, F., and F. J. Swanson. 1993. Effects of coarse woody debris on morphology and sediment
storage of a mountain stream system in western Oregon. Earth Surface Processes and Landforms
18:43-61.
Nakano, S., H. Miyasaka, and N. Kuhara. 1999. Terrestrial-aquatic linkages: Riparian arthropod inputs
alter trophic cascades in a stream food web. Ecology 80:2435-2441.
Nakano, S., and M. Murakami. 2001. Reciprocal subsidies: Dynamic interdependence between terrestrial
and aquatic food webs. Proceedings of the National Academy of Sciences 98:166-170.
Nanson, G. C., and J. C. Croke. 1992. A genetic classification of floodplains. Geomorphology 4:459-486.
National Research Council. 2002. Riparian areas: Functions and strategies for management. The National
Academies Press, Washington, DC.
Nedeau, E. J., R. W. Merritt, and M. G. Kaufman. 2003. The effect of an industrial effluent on an urban
stream benthic community: Water quality vs. habitat quality. Environmental Pollution 123:1-13.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Neubauer, E., S. J. Kb'hler, F. von der Kammer, H. Laudon, and T. Hofmann. 2013. Effect of pH and stream
order on iron and arsenic speciation in boreal catchments. Environmental Science & Technology
47:7120-7128.
Newbold, J. D. 1992. Cycles and spirals of nutrients. Pages 379-408 in The rivers handbook. Volume 1.
Hydrological and ecological principles. P. Calow and G. E. Petts, editors. Blackwell Scientific, Oxford,
UK.
Newbold, J. D., J. W. Elwood, R. V. Oneill, and A. L. Sheldon. 1983a. Phosphorus dynamics in a woodland
stream ecosystem: A study of nutrient spiralling. Ecology 64:1249-1265.
Newbold, J. D., J. W. Elwood, R. V. Oneill, and W. Vanwinkle. 1981. Measuring nutrient spiralling in
streams. Canadian Journal of Fisheries and Aquatic Sciences 38:860-863.
Newbold, J. D., J. W. Elwood, M. S. Schulze, R. W. Stark, and J. C. Barmeier. 1983b. Continuous ammonium
enrichment of a woodland stream: Uptake kinetics, leaf decomposition, and nitrification. Freshwater
Biology 13:193-204.
Newbold, J. D., S. Herbert, B. W. Sweeney, P. Kiry, and S. J. Alberts. 2010. Water quality functions of a 15-
year-old riparian forest buffer system. Journal of the American Water Resources Association
46:299-310.
Newbold, J. D., P. J. Mulholland, J. W. Elwood, and R. V. Oneill. 1982a. Organic carbon spiralling in stream
ecosystems. Oikos 38:266-272.
Newbold, J. D., R. V. O'Neill, J. W. Elwood, and W. Van Winkle. 1982b. Nutrient spiraling in streams:
Implications for nutrient limitation and invertebrate activity. The American Naturalist 120:628-665.
Newbold, J. D., S. A. Thomas, G. W. Minshall, C. E. Gushing, and T. Georgian. 2005. Deposition, benthic
residence, and resuspension of fine organic particles in a mountain stream. Limnology and
Oceanography 50:1571-1580.
Newman, D. G., and C. R. Griffin. 1994. Wetland use by river otters in Massachusetts. Journal of Wildlife
Management 58:18-23.
NHD. 2008. National Hydrography Dataset. U.S. Geological Survey.
Nielsen, J. L. 1992. Microhabitat-specific foraging behavior, diet, and growth of juvenile coho salmon.
Transactions of the American Fisheries Society 121:617-634.
Niemuth, N. D., B. Wangler, and R. E. Reynolds. 2010. Spatial and temporal variation in wet area of
wetlands in the Prairie Pothole Region of North Dakota and South Dakota. Wetlands 30:1053-1064.
Nilsson, C., R. L. Brown, R. Jansson, and D. M. Merritt. 2010. The role of hydrochory in structuring
riparian and wetland vegetation. Biological Reviews 85:837-858.
Nixon, S. W., and V. J. Lee. 1986. Wetlands and water quality: A regional review of recent research in the
United States on the role of freshwater and saltwater wetlands as sources, sinks, and transformers
of nitrogen, phosphorus, and various heavy metals. Technical Report Y-86-2, U.S. Army Corp of
Engineers, Waterways Experiment Station, Vicksburg, MS.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Noe, G. B., and C. R. Hupp. 2005. Carbon, nitrogen, and phosphorus accumulation in floodplains of
Atlantic Coastal Plain rivers, USA. Ecological Applications 15:1178-1190.
Noe, G. B., and C. R. Hupp. 2007. Seasonal variation in nutrient retention during inundation of a short-
hydroperid floodplain. River Research and Applications 23:1088-1101.
Norton, J. B., J. A. Sandor, C. S. White, and V. Laahty. 2007. Organic matter transformations through
arroyos and alluvial fan soils within a Native American agroecosystem. Soil Science Society of
America Journal 71:829-835.
Novitzki, R. P. 1979. Hydrologic characteristics of Wisconsin's wetlands and their influence on floods.
Pages 377-388 in Wetland functions and values: The status of our understanding. Proceedings of the
national symposium on wetlands. P. Greeson, J. Clark, and J. E. Clark, editors. American Water
Resources Association, Minneapolis MN.
Nuff, J. C., and G. P. Asner. 2001. Dissolved organic carbon in terrestrial ecosystems: Synthesis and a
model. Ecosystems 4:29-48.
Nyberg, L, 0. Calles, and L. Greenberg. 2008. Impact of short-term regulation on hyporheic water quality
in a boreal river. River Research and Applications 24:407-419.
O'Brien, J. M., S. K. Hamilton, L. E. Kinsman-Costello, J. T. Lennon, and N. E. Ostrom. 2012. Nitrogen
transformations in a through-flow wetland revealed using whole-ecosystem pulsed 15N additions.
Limnology and Oceanography 57:221-234.
O'Driscoll, M. A., and D. R. DeWalle. 2010. Seeps regulate stream nitrate concentration in a forested
Appalachian catchment. Journal of Environmental Quality 39:420-431.
Obermann, M., K.-H. Rosenwinkel, and M.-G. Tournoud. 2009. Investigation of first flushes in a medium-
sized Mediterranean catchment. Journal of Hydrology 373:405-415.
OMB. 2004. Final information quality bulletin for peer review. (Memorandum M-05-03), Office of
Management and Budget, Washington, DC.
Orr, C. H., K. I. Predick, E. H. Stanley, and K. L. Rogers. 2014. Spatial autocorrelation of denitrification in a
restored and a natural floodplain. Wetlands 34:89-100.
Osborne, L. L., and M. Wiley. 1992. Influence of tributary position on the structure of warmwater fish
communities. Canadian Journal of Fisheries and Aquatic Sciences 49:671-681.
Osterkamp, W., L. Lane, and C. Savard. 1994. Recharge estimates using a geomorphic/distributed-
parameter simulation approach, Amargosa River Basin. Journal of the American Water Resources
Association 30:493-507.
Osterkamp, W. R. 2008. Annotated definitions of selected geomorphic terms and related terms of
hydrology, sedimentology, soil science and ecology USGS Open File Report 2008-1217, U.S.
Department of the Interior, U.S. Geological Survey, Reston, VA.
Osterkamp, W. R., and C. R. Hupp. 2010. Fluvial processes and vegetation— glimpses of the past, the
present, and perhaps the future. Geomorphology 116:274-285.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Paik, K., and P. Kumar. 2004. Hydraulic geometry and the nonlinearity of the network instantaneous
response. Water Resources Research 40:W03602.
Paillex, A., S. Doledec, E. Castella, and S. Merigoux. 2009. Large river floodplain restoration: Predicting
species richness and trait responses to the restoration of hydrological connectivity. Journal of
Applied Ecology 46:250-258.
Paller, M. H. 1994. Relationships between fish assemblage structure and stream order in South-Carolina
coastal-plain streams. Transactions of the American Fisheries Society 123:150-161.
Palmer, M. A., and C. M. Febria. 2012. The heartbeat of ecosystems. Science 336:1393-1394.
Pannell, J. R., and B. Charlesworth. 2000. Effects of metapopulation processes on measures of genetic
diversity. Philosophical Transactions of the Royal Society of London B-Biological Sciences
355:1851-1864.
Park, D., M. Sullivan, E. Bayne, and G. Scrimgeour. 2008. Landscape-level stream fragmentation caused
by hanging culverts along roads in Alberta's boreal forest. Canadian Journal of Forest Research
38:566-575.
Parker, J. K., D. Mclntyre, and R. T. Noble. 2010. Characterizing fecal contamination in stormwater runoff
in coastal North Carolina, USA. Water Research 44:4186-4194.
Pascual-Hortal, L, and S. Saura. 2006. Comparison and development of new graph-based landscape
connectivity indices: towards the priorization of habitat patches and corridors for conservation.
Landscape Ecology 21:959-967.
Paterson, K. G., and J. L. Schnoor. 1992. Fate of alachlor and atrazine in a riparian zone field site. Water
Environment Research 64:274-283.
Patten, D. T., L. Rouse, and J. C. Stromberg. 2008. Isolated spring wetlands in the Great Basin and Mojave
Deserts, USA: Potential response of vegetation to groundwater withdrawal. Environmental
Management 41:398-413.
Patz, M. J., K. J. Reddy, and Q. D. Skinner. 2006. Trace elements in coalbed methane produced water
interacting with semi-arid ephemeral stream channels. Water Air and Soil Pollution 170:55-67.
Paul, M. J., and J. L. Meyer. 2001. Streams in the urban landscape. Annual Review of Ecology and
Systematics 32:333-365.
Paul, M. J., J. L. Meyer, and C. A. Couch. 2006. Leaf breakdown in streams differing in catchment land use.
Freshwater Biology 51:1684-1695.
Payne, S. M., and W. W. Woessner. 2010. An aquifer classification system and Geographical Information
System-based analysis tool for watershed managers in the western U.S. Journal of the American
Water Resources Association 46:1003-1023.
Pearse, A. T., G. L. Krapu, D. A. Brandt, and P. J. Kinzel. 2010. Changes in agriculture and abundance of
snow geese affect carrying capacity of Sandhill cranes in Nebraska. Journal of Wildlife Management
74:479-488.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Pease, A. A., J. J. Davis, M. S. Edwards, and T. F. Turner. 2006. Habitat and resource use by larval and
juvenile fishes in an arid-land river (Rio Grande, New Mexico). Freshwater Biology 51:475-486.
Perkin, J. S., and K. B. Gido. 2012. Fragmentation alters stream fish community structure in dendritic
ecological networks. Ecol Appl 22:2176-2187.
Person, H. S., E. J. Coil, and R. T. Beall. 1936. Little waters. A study of headwater streams and other little
waters, their use and relations to the land. The Upstream Engineering Conference, The 75th
Congress, 2nd Session, Senate Document No. 198, U.S. Government Printing Office, Washington, DC.
Peterjohn, W. T., and D. L. Correll. 1984. Nutrient dynamics in an agricultural watershed: Observations
on the role of a riparian forest. Ecology 65:1466-1475.
Petersen, R. C., and K. W. Cummins. 1974. Leaf processing in a woodland stream. Freshwater Biology
4:343-368.
Peterson, B. J., W. M. Wollheim, P. J. Mulholland, J. R. Webster, J. L. Meyer, J. L. Tank, E. Marti, W. B.
Bowden, H. M. Valett, A. E. Hershey, W. H. McDowell, W. K. Dodds, S. K. Hamilton, S. Gregory, and D.
D. Morrall. 2001. Control of nitrogen export from watersheds by headwater streams. Science
292:86-90.
Peterson, E. E., D. M. Theobald, and J. M. Ver Hoef. 2007. Geostatistical modelling on stream networks:
Developing valid covariance matrices based on hydrologic distance and stream flow. Freshwater
Biology 52:267-279.
Peterson, E. W., and T. B. Sickbert 2006. Stream water bypass through meander neck, laterally extending
the hyporheic zone. Hydrogeology Journal 14:1443-1451.
Petranka, J. W., and C. T. Holbrook. 2006. Wetland restoration for amphibians: Should local sites be
designed to support metapopulations or patchy populations? Restoration Ecology 14:404-411.
Pezeshki, S. R., S. W. Li, F. D. Shields, and L. T. Martin. 2007. Factors governing survival of black willow
(Salix nigrd) cuttings in a streambank restoration project. Ecological Engineering 29:56-65.
Phillips, P. J., and R. J. Shedlock. 1993. Hydrology and chemistry of groundwater and seasonal ponds in
the Atlantic coastal-plain in Delaware, USA. Journal of Hydrology 141:157-178.
Phillips, R., 0. Beeri, and E. DeKeyser. 2005. Remote wetland assessment for Missouri Coteau prairie
glacial basins. Wetlands 25:335-349.
Phillips, R. W., C. Spence, and J. W. Pomeroy. 2011. Connectivity and runoff dynamics in heterogeneous
basins. Hydrological Processes 25:3061-3075.
Pilliod, D. S., C. R. Peterson, and P. I. Ritson. 2002. Seasonal migration of Columbia spotted frogs [Rana
luteiventris) among complementary resources in a high mountain basin. Canadian Journal of Zoology
80:1849-1862.
Pires, A. M., I. G. Cowx, and M. M. Coelho. 1999. Seasonal changes in fish community structure of
intermittent streams in the middle reaches of the Guadiana basin, Portugal. Journal of Fish Biology
54:235-249.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Platania, S. P., and C. S. Altenbach. 1998. Reproductive strategies and egg types of seven Rio Grande
basin cyprinids. Copeia 1998:559-569.
Poff, N. L. 1992. Why disturbances can be predictable: A perspective on the definition of disturbance in
streams. Journal of the North American Benthological Society 11:86-92.
Poff, N. L., J. D. Allan, M. B. Bain, J. R. Karr, K. L. Prestegaard, B. D. Richter, R. E. Sparks, and J. C.
Stromberg. 1997. The natural flow regime: A paradigm for river conservation and restoration.
BioScience 47:769-784.
Poff, N. L., J. D. Olden, D. M. Merritt, and D. M. Pepin. 2007. Homogenization of regional river dynamics by
dams and global biodiversity implications. Proceedings of the National Academy of Sciences of the
United States of America 104:5732-5737.
Polls, G. A., W. B. Anderson, and R. D. Holt. 1997. Toward an integration of landscape and food web
ecology: The dynamics of spatially subsidized food webs. Annual Review of Ecology and Systematics
28:289-316.
Pollock, M. M., G. R. Pess, T. J. Beechie, and D. R. Montgomery. 2004. The importance of beaver ponds to
coho salmon production in the Stillaguamish River Basin, Washington, USA. North American Journal
of Fisheries Management 24:749-760.
Pollux, B. J. A., N. J. Ouborg, J. M. Van Groenendael, and M. Klaassen. 2007. Consequences of intraspecific
seed-size variation in Sparganium emersum for dispersal by fish. Functional Ecology 21:1084-1091.
Poole, G. C. 2010. Stream hydrogeomorphology as a physical science basis for advances in stream
ecology. Journal of the North American Benthological Society 29:12-25.
Poole, G. C., and C. H. Berman. 2001. An ecological perspective on in-stream temperature: Natural heat
dynamics and mechanisms of human-caused thermal degradation. Environmental Management
27:787-802.
Poole, G. C., J. A. Stanford, S. W. Running, and C. A. Frissell. 2006. Multiscale geomorphic drivers of
groundwater flow paths: Subsurface hydrologic dynamics and hyporheic diversity. Journal of the
North American Benthological Society 25:288-303.
Porvari, P., and M. Verta. 2003. Total and methyl mercury concentrations and fluxes from small boreal
forest catchments in Finland. Environmental Pollution 123:181-191.
Powell, D. M., R. Brazier, A. Parsons, J. Wainwright, and M. Nichols. 2007. Sediment transfer and storage
in dryland headwater streams. Geomorphology 88:152-166.
Power, G., R. S. Brown, and J. G. Imhof. 1999. Groundwater and fish: Insights from northern North
America. Hydrological Processes 13:401-422.
Power, M. E., and W. E. Dietrich. 2002. Food webs in river networks. Ecological Research 17:451-471.
Power, M. E., G. Parker, W. E. Dietrich, and A. Sun. 1995a. How does floodplain width affect floodplain
river ecology? A preliminary exploration using simulations. Geomorphology 13:301-317.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Power, M. E., A. Sun, G. Parker, W. E. Dietrich, and J. T. Wootton. 1995b. Hydraulic food-chain models.
BioScience 45:159-167.
Powers, S. M., R. A. Johnson, and E. H. Stanley. 2012. Nutrient retention and the problem of hydrologic
disconnection in streams and wetlands. Ecosystems 15:435-449.
Pringle, C., G. Blake, A. Covich, K. Buzby, and A. Finley. 1993. Effects of omnivorous shrimp in a montane
tropical stream: Sediment removal, disturbance of sessile invertebrates and enhancement of
understory algal biomass. Oecologia 93:1-11.
Pringle, C. M. 2001. Hydrologic connectivity and the management of biological reserves: A global
perspective. Ecological Applications 11:981-998.
Progar, D. J., and A. R. Moldenke. 2002. Insect production from temporary and perennially flowing
headwater streams in western Oregon. Journal of Freshwater Ecology 17:391-407.
Pyzoha, J. E., T. J. Callahan, G. Sun, C. C. Trettin, and M. Miwa. 2008. A conceptual hydrologic model for a
forested Carolina bay depressional wetland on the Coastal Plain of South Carolina, USA. Hydrological
Processes 22:2689-2698.
Quinn, P. F., K. J. Beven, and R. Lamb. 1995. The in(a/tan/(3) index: How to calculate it and how to use it
within the topmodel framework. Hydrological Processes 9:161-182.
Quinton, W. L, M. Hayashi, and A. Pietroniro. 2003. Connectivity and storage functions of channel fens
and flat bogs in northern basins. Hydrological Processes 17:3665-3684.
Rabalais, N. N., R. E. Turner, and W. J. Wiseman. 2002. Gulf of Mexico hypoxia, a.k.a. "the dead zone".
Annual Review of Ecology and Systematics 33:235-263.
Rader, R. B. 1997. A functional classification of the drift: Traits that influence invertebrate availability to
salmonids. Canadian Journal of Fisheries and Aquatic Sciences 54:1211-1234.
Rae, C., K. Rothley, and S. Dragicevic. 2007. Implications of error and uncertainty for an environmental
planning scenario: A sensitivity analysis of CIS-based variables in a reserve design exercise.
Landscape and Urban Planning 79:210-217.
Rains, M. C. 2011. Water sources and hydrodynamics of closed-basin depressions, Cook Inlet region,
Alaska. Wetlands 31:377-387.
Rains, M. C., R. A. Dahlgren, G. E. Fogg, T. Harter, and R. J. Williamson. 2008. Geological control of physical
and chemical hydrology in California vernal pools. Wetlands 28:347-362.
Rains, M. C., G. E. Fogg, T. Harter, R. A. Dahlgren, and R. J. Williamson. 2006. The role of perched aquifers
in hydrological connectivity and biogeochemical processes in vernal pool landscapes, Central Valley,
California. Hydrological Processes 20:1157-1175.
Randall, G. W., D. R. Huggins, M. P. Russelle, D. J. Fuchs, W. W. Nelson, and J. L. Anderson. 1997. Nitrate
losses through subsurface tile drainage in conservation reserve program, alfalfa, and row crop
systems. Journal of Environmental Quality 26:1240-1247.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Rassam, D. W., C. S. Fellows, R. De Hayr, H. Hunter, and P. Bloesch. 2006. The hydrology of riparian buffer
zones; two case studies in an ephemeral and a perennial stream. Journal of Hydrology 325:308-324.
Rayfield, B., M.-J. Fortin, and A. Fall. 2010. Connectivity for conservation: A framework to classify
network measures. Ecology 92:847-858.
Reaney, S. M. 2008. The use of agent based modelling techniques in hydrology: Determining the spatial
and temporal origin of channel flow in semi-arid catchments. Earth Surface Processes and
Landforms 33:317-327.
Reddy, K. R., and R. D. DeLaune. 2008. Biogeochemistry of wetlands: Science and applications. CRC Press,
Boca Raton, FL.
Reddy, K. R., R. H. Kadlec, E. Flaig, and P. M. Gale. 1999. Phosphorus retention in streams and wetlands: A
review. Critical Reviews in Environmental Science and Technology 29:83-146.
Reeves, G. H., K. M. Burnett, and E. V. McGarry. 2003. Sources of large wood in the main stem of a fourth-
order watershed in coastal Oregon. Canadian Journal of Forest Research 33:1363-1370.
Rehg, K. J., A. I. Packman, and J. Ren. 2005. Effects of suspended sediment characteristics and bed
sediment transport on streambed clogging. Hydrological Processes 19:413-427.
Renard, K. G., and R. V. Keppel. 1966. Hydrographs of ephemeral streams in the southwest. Journal of the
Hydraulics Division Proceedings of the American Society of Civil Engineers 92(HY2):33-52.
Reneau, S. L, P. G. Drakos, D. Katzman, D. V. Malmon, E. V. McDonald, and R. T. Ryti. 2004. Geomorphic
controls on contaminant distribution along an ephemeral stream. Earth Surface Processes and
Landforms 29:1209-1223.
Renwick, W., S. Smith, J. Bartley, and R. Buddemeier. 2005. The role of impoundments in the sediment
budget of the conterminous United States. Geomorphology 71:99-111.
Resh, V. H., A. V. Brown, A. P. Covich, M. E. Gurtz, H. W. Li, G. W. Minshall, S. R. Reice, A. L. Sheldon, J. B.
Wallace, and R. C. Wissmar. 1988. The role of disturbance in stream ecology. Journal of the North
American Benthological Society 7:433-455.
Rhoads, B. L. 1987. Changes in stream channel characteristics at tributary junctions. Physical Geography
8:346-367.
Ribeiro, M. L., K. Blanckaert, A. G. Roy, and A. J. Schleiss. 2012. Flow and sediment dynamics in channel
confluences. Journal of Geophysical Research: Earth Surface 117:F01035.
Rice, S., and M. Church. 1998. Grain size along two gravel-bed rivers: Statistical variation, spatial pattern
and sedimentary links. Earth Surface Processes and Landforms 23:345-363.
Rice, S. P., M. T. Greenwood, and C. B. Joyce. 2001. Tributaries, sediment sources, and the longitudinal
organization of macroinvertebrate fauna along river systems. Canadian Journal of Fisheries and
Aquatic Sciences 58:824-840.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Rice, S. P., P. M. Kiffney, C. M. Greene, and G. R. Press. 2008. The ecological importance of tributaries and
confluences. Pages 209-242 in River Confluences, Tributaries and the Fluvial Network. S. P. Rice, A.
G. Roy, and B. L. Rhoads, editors. John Wiley & Sons, Chichester, UK.
Richardson, D. C., L. A. Kaplan, J. D. Newbold, and A. K. Aufdenkampe. 2009. Temporal dynamics of
seston: A recurring nighttime peak and seasonal shifts in composition in a stream ecosystem.
Limnology and Oceanography 54:344-354.
Richardson, J. L., L. P. Wilding, and R. B. Daniels. 1992. Recharge and discharge of groundwater in aquic
conditions illustrated with flownet analysis. Geoderma 53:65-78.
Richardson, J. S., R. J. Naiman, F. J. Swanson, and D. E. Hibbs. 2005. Riparian communities associated with
Pacific North west headwater streams: Assemblages, processes, and uniqueness. Journal of the
American Water Resources Association 41:935-947.
Rieke-Zapp, D. H., and M. H. Nichols. 2011. Headcut retreat in a semiarid watershed in the southwestern
United States since 1935. CATENA 87:1-10.
Rinaldi, M., B. Wyzga, and N. Surian. 2005. Sediment mining in alluvial channels: Physical effects and
management perspectives. River Research and Applications 21:805-828.
Rinaldo, A., A. Marani, and R. Rigon. 1991. Geomorphological dispersion. Water Resources Research
27:513-525.
Roach, W. J., J. B. Heffernan, N. B. Grimm, J. R. Arrowsmith, C. Eisinger, and T. Rychener. 2008.
Unintended consequences of urbanization for aquatic ecosystems: A case study from the Arizona
desert. BioScience 58:715-727.
Robinson, C. T., K. Tockner, and J. V. Ward. 2002. The fauna of dynamic riverine landscapes. Freshwater
Biology 47:661-677.
Rodriguez-Iturbe, I., R. Muneepeerakul, E. Bertuzzo, S. A. Levin, and A. Rinaldo. 2009. River networks as
ecological corridors: A complex systems perspective for integrating hydrologic, geomorphologic,
and ecologic dynamics. Water Resources Research 45:W01413.
Rodriguez-Iturbe, L, and J. B. Valdes. 1979. The geomorphologic structure of hydrologic response. Water
Resources Research 15:1409-1420.
Roe, J. H., B. A. Kingsbury, and N. R. Herbert. 2004. Comparative water snake ecology: Conservation of
mobile animals that use temporally dynamic resources. Biological Conservation 118:79-89.
Rooney, R. C., C. Carli, and S. Bayley. 2013. River connectivity affects submerged and floating aquatic
vegetation in floodplain wetlands. Wetlands 33:1165-1177.
Rorabaugh, J. C. 2005. Rana pipiens, northern leopard frog. Pages 570-577 in Amphibian declines: The
conservation status of United States species. M. Lannoo, editor. University of California Press,
Berkeley, CA.
Rorabaugh, J. C., and M. J. Lannoo. 2005. Pseudacris regilla Baird and Girard, 1852(b), Pacific treefrog.
Pages 478-484 in Amphibian declines: The conservation status of United States species. M. Lannoo,
editor. University of California Press, Berkeley, CA.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Rosado,}., M. Morals, and K. Tockner. 2015. Mass dispersal of terrestrial organisms during first flush
events in a temporary stream. River Research and Applications: doi: 10.1002/rra.2791.
Roscher, J. P. 1967. Alga dispersal by muskrat intestinal contents. Transactions of the American
Microscopical Society 86:497-498.
Rose, S. 2007. The effects of urbanization on the hydrochemistry of base flow within the Chattahoochee
River Basin (Georgia, USA). Journal of Hydrology 341:42-54.
Rosenberry, D. 0., and T. C. Winter. 1997. Dynamics of water-table fluctuations in an upland between
two prairie-pothole wetlands in North Dakota. Journal of Hydrology 191:266-289.
Rosenfeld, J. S., and E. Raeburn. 2009. Effects of habitat and internal prey subsidies on juvenile coho
salmon growth: Implications for stream productive capacity. Ecology of Freshwater Fish 18:572-
584.
Roth, N. E., J. D. Allan, and D. L. Erickson. 1996. Landscape influences on stream biotic integrity assessed
at multiple spatial scales. Landscape Ecology 11:141-156.
Roulet, N. T., and M.-K. Woo. 1986. Hydrology of a wetland in the continuous permafrost region. Journal
of Hydrology 89:73-91.
Rovansek, R. J., L. D. Hinzman, and D. L. Kane. 1996. Hydrology of a tundra wetland complex on the
Alaskan arctic coastal plain, USA. Arctic and Alpine Research 28:311-317.
Rowan, J. S., S. J. A. Barnes, S. L. Hetherington, B. Lambers, and F. Parsons. 1995. Geomorphology and
pollution: The environmental impacts of lead mining, Leadhills, Scotland. Journal of Geochemical
Exploration 52:57-65.
Roy, A. G., and M. J. Woldenberg. 1986. A model for changes in channel form at a river confluence. Journal
of Geology 94:402-411.
Roy, A. H., A. L. Dybas, K. M. Fritz, and H. R. Lubbers. 2009. Urbanization affects the extent and hydrologic
permanence of headwater streams in a midwestern US metropolitan area. Journal of the North
American Benthological Society 28:911-928.
Royer, T. V., M. B. David, and L. E. Gentry. 2006. Timing of riverine export of nitrate and phosphorus from
agricultural watersheds in Illinois: Implications for reducing nutrient loading to the Mississippi
River. Environmental Science & Technology 40:4126-4131.
Royer, T. V., J. L. Tank, and M. B. David. 2004. Transport and fate of nitrate in headwater agricultural
streams in Illinois. Journal of Environmental Quality 33:1296-1304.
Rutherford, J. C., N. A. Marsh, P. M. Davies, and S. E. Bunn. 2004. Effects of patchy shade on stream water
temperature: How quickly do small streams heat and cool? Marine and Freshwater Research
55:737-748.
Rypel, A. L., D. A. Arrington, and R. H. Findlay. 2008. Mercury in southeastern US riverine fish
populations linked to water body type. Environmental Science & Technology 42:5118-5124.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Saco, P. M., and P. Kumar. 2002. Kinematic dispersion in stream networks 1. Coupling hydraulics and
network geometry. Water Resources Research 38:1244.
Saco, P. M., and P. Kumar. 2008. Hydrologic dispersion in fluvial networks. Pages 307-335 in River
confluences, tributaries and the fluvial network. S. P. Rice, A. G. Roy, and B. L. Rhoads, editors. John
Wiley & Sons, Chichester, UK.
Sandor, J. A., J. B. Norton, J. A. Homburg, D. A. Muenchrath, C. S. White, S. E. Williams, C. I. Havener, and P.
D. Stahl. 2007. Biogeochemical studies of a Native American runoff agroecosystem. Geoarchaeology
22:359-386.
Sass, G. Z., and I. F. Creed. 2008. Characterizing hydrodynamics on boreal landscapes using archived
synthetic aperture radar imagery. Hydrological Processes 22:1687-1699.
Saura, S., and L. Pascual-Hortal. 2007. A new habitat availability index to integrate connectivity in
landscape conservation planning: Comparison with existing indices and application to a case study.
Landscape and Urban Planning 83:91-103.
Sawyer, A. H., M. Bayani Cardenas, and J. Buttles. 2011. Hyporheic exchange due to channel-spanning
logs. Water Resources Research 47:W08502.
Sawyer, A. H., M. Bayani Cardenas, and J. Buttles. 2012. Hyporheic temperature dynamics and heat
exchange near channel-spanning logs. Water Resources Research 48:W01529.
Scanlon, B. R., C. C. Faunt, L. Longuevergne, R. C. Reedy, W. M. Alley, V. L. McGuire, and P. B. McMahon.
2012. Groundwater depletion and sustainability of irrigation in the US High Plains and Central
Valley. Proceedings of the National Academy of Sciences 109:9320-9325.
Scanlon, B. R., R. S. Goldsmith, and J. G. Paine. 1997. Analysis of focused unsaturated flow beneath
fissures in the Chihuahuan Desert, Texas, USA. Journal of Hydrology 203:58-78.
Scanlon, B. R., K. E. Keese, A. L. Flint, L. E. Flint, C. B. Gaye, W. M. Edmunds, and I. Simmers. 2006. Global
synthesis of groundwater recharge in semiarid and arid regions. Hydrological Processes 20:3335-
3370.
Schalk, C. M., and T. M. Luhring. 2010. Vagility of aquatic salamanders: Implications for wetland
connectivity. Journal of Herpetology 44:104-109.
Scheidegger, A. E. 1965. The algebra of stream-order numbers. USGS Professional Paper 525-B, U.S.
Department of the Interior, U.S. Geological Survey, Washington, DC.
Schelker, J., K. Ohman, S. Lb'fgren, and H. Laudon. 2014. Scaling of increased dissolved organic carbon
inputs by forest clear-cutting - What arrives downstream? Journal of Hydrology 508:299-306.
Schemel, L. E., T. R. Sommer, A. B. Muller-Solger, and W. C. Harrell. 2004. Hydrologic variability, water
chemistry, and phytoplankton biomass in a large floodplain of the Sacramento River, CA, USA.
Hydrobiologia 513:129-139.
Schiller, K. E., P. Jindal, N. B. Basu, and M. J. Helmers. 2012. Impact of artificial subsurface drainage on
groundwater travel times and baseflow discharge in an agricultural watershed, Iowa (USA).
Hydrological Processes 26:3092-3100.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Schindler, D. E., P. R. Leavitt, C. S. Brock, S. P. Johnson, and P. D. Quay. 2005. Marine-derived nutrients,
commercial fisheries, and production of salmon and lake algae in Alaska. Ecology 86:3225-3231.
Schindler, D. W., and P. J. Curtis. 1997. The role of DOC in protecting freshwaters subjected to climate
warming and acidification from UV exposure. Biogeochemistry 36:1-8.
Schlesinger, W. H., and C. S. Jones. 1984. The comparative importance of overland runoff and mean
annual rainfall to shrub communities of the Mojave Desert. Botanical Gazette 145:116-124.
Schlosser, I. J. 1987. A conceptual framework for fish communities in small warmwater streams. Pages
17-24 in Community and evolutionary ecology of North American stream fishes. W. J. Matthews and
D. C. Heins, editors. University of Oklahoma Press, Norman, OK.
Schlosser, I. J. 1991. Stream fish ecology: A landscape perspective. BioScience:704-712.
Schmidt, J. C., and P. R. Wilcock. 2008. Metrics for assessing the downstream effects of dams. Water
Resources Research 44:W04404.
Schneider, R. L, and R. R. Sharitz. 1988. Hydrochory and regeneration in a bald cypress water tupelo
swamp forest. Ecology 69:1055-1063.
Schramm, H. L., and M. A. Eggleton. 2006. Applicability of the flood-pulse concept in a temporal
floodplain river ecosystem: Thermal and temporal components. River Research and Applications
22:543-553.
Schrank, A. J., and F. J. Rahel. 2004. Movement patterns in inland cutthroat trout (Oncorhynchus clarki
Utah): Management and conservation implications. Canadian Journal of Fisheries and Aquatic
Sciences 61:1528-1537.
Schrank, S. J., C. S. Guy, M. R. Whiles, and B. L. Brock. 2001. Influence of instream and landscape-level
factors on the distribution of Topeka shiners Notropis topeka in Kansas streams. Copeia 2001:413-
421.
Schroder, B. 2006. Pattern, process, and function in landscape ecology and catchment hydrology— how
can quantitative landscape ecology support predictions in ungauged basins? Hydrology and Earth
System Sciences 10:967-979.
Schumm, S. A. 1956. The evolution of drainage systems and slopes in badlands at Perth Amboy, New
Jersey. Bulletin of the Geological Society of America 67:597-646.
Schwalb, A. N., M. Garvie, and J. D. Ackerman. 2010. Dispersion of freshwater mussel larvae in a lowland
river. Limnology and Oceanography 55:628-638.
Schwalb, A. N., T. J. Morris, N. E. Mandrak, and K. Cottenie. 2013. Distribution of unionid freshwater
mussels depends on the distribution of host fishes on a regional scale. Diversity and Distributions
19:446-454.
Scott, R. L., W. J. Shuttleworth, T. 0. Keefer, and A. W. Warrick. 2000. Modeling multiyear observations of
soil moisture recharge in the semiarid American Southwest. Water Resources Research 36:2233-
2247.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Scrivener, J. C., T. G. Brown, and B. C. Andersen. 1994. Juvenile Chinook salmon (Oncorhynchus
tshawytschd) utilization of Hawks Creek, a small and nonnatal tributary of the Upper Fraser River.
Canadian Journal of Fisheries and Aquatic Sciences 51:1139-1146.
Sear, D. A., C. E. Millington, D. R. Kitts, and R. Jeffries. 2010. Logjam controls on channel: Floodplain
interactions in wooded catchments and their role in the formation of multi-channel patterns.
Geomorphology 116:305-319.
Sedell, J. R., G. H. Reeves, F. R. Hauer, J. A. Stanford, and C. P. Hawkins. 1990. Role of refugia in recovery
from disturbances: Modern fragmented and disconnected river systems. Environmental
Management 14:711-724.
Seedang, S., A. G. Fernald, R. M. Adams, and D. H. Landers. 2008. Economic analysis of water temperature
reduction practices in a large river floodplain: An exploratory study of the Willamette River, Oregon.
River Research and Applications 24:941-959.
Seitzinger, S. P., R. W. Sanders, and R. Styles. 2002. Bioavailability of DON from natural and
anthropogenic sources to estuarine plankton. Limnology and Oceanography 47:353-366.
Selvendiran, P., C. T. Driscoll, J. T. Bushey, and M. R. Montesdeoca. 2008. Wetland influence on mercury
fate and transport in a temperate forested watershed. Environmental Pollution 154:46-55.
Semlitsch, R. D. 2008. Differentiating migration and dispersal processes for pond-breeding amphibians.
The Journal of Wildlife Management 72:260-267.
Semlitsch, R. D., and J. R. Bodie. 2003. Biological criteria for buffer zones around wetlands and riparian
habitats for amphibians and reptiles. Conservation Biology 17:1219-1228.
Senderovich, Y., I. Izhaki, and M. Halpern. 2010. Fish as reservoirs and vectors of Vibrio cholerae. PLOS
ONE5:e8607.
Seo, J. I., F. Nakamura, D. Nakano, H. Ichiyanagi, and K. W. Chun. 2008. Factors controlling the fluvial
export of large woody debris, and its contribution to organic carbon budgets at watershed scales.
Water Resources Research 44:W04428.
Serfass, T. L, M. J. Lovallo, R. P. Brooks, A. H. Hayden, and D. H. Mitcheltree. 1999. Status and distribution
of river otters in Pennsylvania following a reintroduction project. Journal of the Pennsylvania
Academy of Science 73:10-14.
Shaman, J., M. Stieglitz, andD. Burns. 2004. Are big basins just the sum of small catchments?
Hydrological Processes 18:3195-3206.
Shanks, C. E., and G. C. Arthur. 1952. Muskrat movements and population dynamics in Missouri farm
ponds and streams. Journal of Wildlife Management 16:138-148.
Sharitz, R. R. 2003. Carolina bay wetlands: Unique habitats of the southeastern United States. Wetlands
23:550-562.
Sharma, K. D., and J. S. R. Murthy. 1995. Hydrologic routing of flow in arid ephemeral channels. Journal of
Hydraulic Engineering 121:466-471.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Sharpley, A. N., and S. Rekolainen. 1997. Phosphorus in agriculture and its environmental implications.
Pages 1-54 in Phosphorus losses from soil to water. H. Tunney, 0. T. Carton, P. C. Brookes, and A. E.
Johnston, editors. CAB International, Cambridge, UK.
Shaw, D. A., A. Pietroniro, and L. W. Martz. 2013. Topographic analysis for the prairie pothole region of
western Canada. Hydrological Processes 27:3105-3114.
Shaw, D. A., G. Vanderkamp, F. M. Conly, A. Pietroniro, and L. Martz. 2012. The fill-spill hydrology of
prairie wetland complexes during drought and deluge. Hydrological Processes 26:3147-3156.
Shaw, P. A., and R. G. Bryant. 2011. Pans, playas and salt lakes. Pages 373-401 in Arid zone
geomorphology: Process, form and change in drylands. D. S. G. Thomas, editor. John Wiley & Sons,
New York, NY.
Sheaves, M. 2009. Consequences of ecological connectivity: the coastal ecosystem mosaic. Marine
Ecology Progress Series 391:107-115.
Shedlock, R. J., D. A. Wilcox, T. A. Thompson, and D. A. Cohen. 1993. Interactions between ground-water
and wetlands, southern shore of Lake-Michigan, USA. Journal of Hydrology 141:127-155.
Sheldon, A. L. 1988. Conservation of stream fishes: Patterns of diversity, rarity, and risk. Conservation
Biology 2:149-156.
Shentsis, I., and E. Rosenthal. 2003. Recharge of aquifers by flood events in an arid region. Hydrological
Processes 17:695-712.
Shih, J. S., R. B. Alexander, R. A. Smith, E. W. Boyer, G. E. Schwarz, and S. Chung. 2010. An initial
SPARROW model of land use and in-stream controls on total organic carbon in streams of the
conterminous United States. USGS Open File Report 2010-1276, U.S. Department of the Interior, U.S.
Geological Survey, Reston, VA.
Shook, K. R., and J. W. Pomeroy. 2011. Memory effects of depressional storage in northern prairie
hydrology. Hydrological Processes 25:3890-3898.
Shoup, D. E., and D. H. Wahl. 2009. Fish diversity and abundance in relation to interannual and lake-
specific variation in abiotic characteristics of floodplain lakes of the lower Kaskaskia River, Illinois.
Transactions of the American Fisheries Society 138:1076-1092.
Shreve, R. L. 1967. Infinite topologically random channel networks. Journal of Geology 75:178-186.
Simmons, J. A. 2010. Phosphorus removal by sediment in streams contaminated with acid mine
drainage. Water Air and Soil Pollution 209:123-132.
Sjodin, A., W. M. Lewis, and J. F. Saunders. 2001. Analysis of groundwater exchange for a large plains
river in Colorado (USA). Hydrological Processes 15:609-620.
Skagen, S. K., and F. L. Knopf. 1993. Toward conservation of midcontinental shorebird migrations.
Conservation Biology 7:533-541.
Smith, M. W., L. J. Bracken, and N. J. Cox. 2010. Toward a dynamic representation of hydrological
connectivity at the hillslope scale in semiarid areas. Water Resources Research 46:W12540.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Smith, R. D., A. Ammann, C. Bartoldus, and M. M. Brinson. 1995. An approach for assessing wetland
functions using hydrogeomorphic classification, reference wetlands, and functional indices.
Wetlands Research Program Technical Report WRP-DE-9, U.S. Army Corps of Engineers, Waterways
Experiment Station, Wetlands Research Program, Vicksburg, MS.
Smith, T. A., and C. E. Kraft. 2005. Stream fish assemblages in relation to landscape position and local
habitat variables. Transactions of the American Fisheries Society 134:430-440.
Smock, L. A. 1994. Movements of invertebrates between stream channels and forested floodplains.
Journal of the North American Benthological Society 13:524-531.
Smock, L. A., J. E. Gladden, J. L. Riekenberg, L. C. Smith, and C. R. Black. 1992. Lotic macroinvertebrate
production in three dimensions: Channel surface, hyporheic, and floodplain environments. Ecology
73:876-886.
Snell, J. D., and M. Sivapalan. 1994. On geomorphological dispersion in natural catchments and the
geomorphological unit hydrograph. Water Resources Research 30:2311-2323.
Snodgrass, J. W., A. L. Bryan, R. F. Lide, and G. M. Smith. 1996. Factors affecting the occurrence and
structure offish assemblages in isolated wetlands of the upper Coastal Plain, USA. Canadian Journal
of Fisheries and Aquatic Sciences 53:443-454.
Snodgrass, J. W., M. J. Komoroski, A. L. Bryan, and J. Burger. 2000. Relationships among isolated wetland
size, hydroperiod, and amphibian species richness: Implications for wetland regulations.
Conservation Biology 14:414-419.
Sobota, D. J., S. V. Gregory, and J. Van Sickle. 2006. Riparian tree fall directionality and modeling large
wood recruitment to streams. Canadian Journal of Forest Research 36:1243-1254.
Soons, M. B. 2006. Wind dispersal in freshwater wetlands: Knowledge for conservation and restoration.
Applied Vegetation Science 9:271-278.
Soons, M. B., and G. W. Heil. 2002. Reduced colonization capacity in fragmented populations of wind-
dispersed grassland forbs. Journal of Ecology 90:1033-1043.
Spence, C., and R. W. Phillips. 2015. Refining understanding of hydrological connectivity in a boreal
catchment. Hydrological Processes: doi: 10.1002/hyp.10270.
Spinola, R. M., T. L. Serfass, and R. P. Brooks. 2008. Survival and post-release movements of river otters
translocated to western New York. Northeastern Naturalist 15:13-24.
Squillace, P. J. 1996. Observed and simulated movement of bank-storage water. Ground Water 34:121-
134.
SRMGSC. 2007. Management plan of the Pacific and Central Flyways for the Rocky Mountain population
of greater sandhill cranes. [Joint] Subcommittees, Rocky Mountain Population Greater Sandhill
Cranes, Pacific Flyway Study Committee, Central Flyway Webless Migratory Game Bird Tech.
Committee [c/o USFWS, MEMO], Portland, OR.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
St. Louis, V. L, J. W. M. Rudd, C. A. Kelly, K. G. Beaty, N. S. Bloom, and R. J. Flett. 1994. Importance of
wetlands as sources of methyl mercury to boreal forest ecosystems. Canadian Journal of Fisheries
and Aquatic Sciences 51:1065-1076.
St. Louis, V. L., J. W. M. Rudd, C. A. Kelly, K. G. Beaty, R. J. Flett, and N. Roulet. 1996. Production and loss of
methylmercury and loss of total mercury from boreal forest catchments containing different types of
wetlands. Environmental Science & Technology 30:2719-2729.
Staddon, W. J., R. M. Zablotowicz, and M. A. Locke. 2001. Microbiological characteristics of a vegetative
buffer strip soil and degradation and sorption of metolachlor. Soil Science Society of America Journal
65:1136-1142.
Stanford, J. A., and J. V. Ward. 1984. The effects of regulation on the limnology of the Gunnison River: A
North American case history. Pages 467-480 in Regulated rivers. Proceedings of the second
international symposium on regulated streams, Oslo, August 1982. A. Lillehammer and S. J. Salveit,
editors. Universitetsforlaget AS, Oslo, Norway.
Stanford, J. A., and J. V. Ward. 1988. The hyporheic habitat of river ecosystems. Nature 335:64-66.
Stanford, J. A., and J. V. Ward. 1993. An ecosystem perspective of alluvial rivers: Connectivity and the
hyporheic corridor. Journal of the North American Benthological Society 12:48-60.
Stanley, E. H., S. G. Fisher, and N. B. Grimm. 1997. Ecosystem expansion and contraction in streams.
BioScience 47:427-435.
Statzner, B. 2012. Geomorphological implications of engineering bed sediments by lotic animals.
Geomorphology 157-158:49-65.
Stead, T. K., J. M. Schmid-Araya, and A. G. Hildrew. 2005. Secondary production of a stream metazoan
community: Does the meiofauna make a difference? Limnology and Oceanography 50:398-403.
Steward, A. L., D. von Schiller, K. Tockner, J. C. Marshall, and S. E. Bunn. 2012. When the river runs dry:
Human and ecological values of dry riverbeds. Frontiers in Ecology and the Environment 10:202-
209.
Stoddard, J. L., D. V. Peck, A. R. Olsen, D. P. Larsen, J. Van Sickle, C. P. Hawkins, R. M. Hughes, T. R.
Whittier, G. Lomnicky, A. T. Herlihy, P. R. Kaufmann, S. A. Peterson, P. L. Ringold, S. G. Paulsen, and R.
Blair. 2005. Environmental Monitoring and Assessment Program (EMAP) western streams and
rivers statistical summary. EPA 620/R-05/006, U.S. Environmental Protection Agency, Office of
Research and Development, Washington, DC.
Stonedahl, S. H., J. W. Harvey, A. Wb'rman, M. Salehin, and A. I. Packman. 2010. A multiscale model for
integrating hyporheic exchange from ripples to meanders. Water Resources Research 46:W12539.
Strahler, A. N. 1957. Quantitative analysis of watershed geomorphology. American Geophysical Union
Transactions 38:913-920.
Subalusky, A. L., L. A. Fitzgerald, and L. L. Smith. 2009a. Ontogenetic niche shifts in the American
alligator establish functional connectivity between aquatic systems. Biological Conservation
142:1507-1514.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Subalusky, A. L, L. 1. Smith, and L. A. Fitzgerald. 2009b. Detection of American alligators in isolated,
seasonal wetlands. Applied Herpetology 6:199-210.
Sullivan, S. M. P., and M. C. Watzin. 2009. Stream-floodplain connectivity and fish assemblage diversity in
the Champlain Valley, Vermont, U.S.A. Journal of Fish Biology 74:1394-1418.
Sun, G. W., H. Riekerk, and L. V. Korhnak. 1995. Shallow groundwater table dynamics of cypress wetland
pine upland systems in Florida flatwoods. Soil and Crop Science Society of Florida Proceedings
54:66-71.
Suter, G. W., and S. M. Cormier. 2011. Why and how to combine evidence in environmental assessments:
Weighing evidence and building cases. Science of The Total Environment 409:1406-1417.
Suter, G. W., S. B. Norton, and S. M. Cormier. 2002. A methodology for inferring the causes of observed
impairments in aquatic ecosystems. Environmental Toxicology and Chemistry 21:1101-1111.
Swan, C. M., and M. A. Palmer. 2006. Composition of speciose leaf litter alters stream detritivore growth,
feeding activity and leaf breakdown. Oecologia 147:469-478.
Swanson, C. D., and R. W. Bachmann. 1976. Model of algal exports in some Iowa streams. Ecology
57:1076-1080.
Swanson, F.}., S. L. Johnson, S. V. Gregory, and S. A. Acker. 1998. Flood disturbance in a forested
mountain landscape - Interactions of land use and floods. BioScience 48:681-689.
Swimley, T. J., R. P. Brooks, and T. L. Serfass. 1999. Otter and beaver interactions in the Delaware Water
Gap National Recreation Area. Journal of the Pennsylvania Academy of Science 72:97-101.
Tang, C., I. Machida, S. Shindo, A. Kondoh, and Y. Sakura. 2001. Chemical and isotopic methods for
confirming the roles of wadis in regional groundwater recharge in a regional arid environment: A
case study in Al Ain, UAE. Hydrological Processes 15:2195-2202.
Tank, J. L., E. J. Rosi-Marshall, N. A. Griffiths, S. A. Entrekin, and M. L. Stephen. 2010. A review of
allochthonous organic matter dynamics and metabolism in streams. Journal of the North American
Benthological Society 29:118-146.
Tarboton, D. G. 1997. A new method for the determination of flow directions and upslope areas in grid
digital elevation models. Water Resources Research 33:309-319.
Taylor, B. W., A. S. Flecker, and R. 0. Hall. 2006. Loss of a harvested fish species disrupts carbon flow in a
diverse tropical river. Science 313:833-836.
Taylor, M. P., and J. A. Little. 2013. Environmental impactof a major copper mine spill on a river and
floodplain system. Anthropocene 3:36-50.
Temnerud, J., J. Fb'lster, I. Buffam, H. Laudon, M. Erlandsson, and K. Bishop. 2010. Can the distribution of
headwater stream chemistry be predicted from downstream observations? Hydrological Processes
24:2269-2276.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Temnerud,}.,}. Siebert, M. Jansson, and K. Bishop. 2007. Spatial variation in discharge and
concentrations of organic carbon in a catchment network of boreal streams in northern Sweden.
Journal of Hydrology 342:72-87.
Tetzlaff, D., and C. Soulsby. 2008. Sources of baseflow in larger catchments- using tracers to develop a
holistic understanding of runoff generation. Journal of Hydrology 359:287-302.
Thomas, J. R., B. Middleton, and D. J. Gibson. 2006. A landscape perspective of the stream corridor
invasion and habitat characteristics of an exotic (Dioscorea oppositifolia) in a pristine watershed in
Illinois. Biological Invasions 8:1103-1113.
Thorp, J. H., M. C. Thorns, and M. D. Belong. 2006. The riverine ecosystem synthesis: Biocomplexity in
river networks across space and time. River Research and Applications 22:123-147.
Thurman, E. M. 1985. Organic geochemistry of natural waters. Martinus Nijhoff/Dr. W. Junk Publishers,
Boston, MA.
Thurman, E. M., D. A. Goolsby, M. T. Meyer, and D. W. Kolpin. 1991. Herbicides in surface waters of the
midwestern United States: The effect of spring flush. Environmental Science & Technology 25:1794-
1796.
Tihansky, A. B. 1999. Sinkholes, west-central Florida. Pages 121-140 in Land subsidence in the United
States. USGS Circular 1182. D. Galloway, D. R. Jones, and S. E. Ingebritsen, editors. U.S. Department of
the Interior, U.S. Geological Survey, Reston, VA.
Timm, B. C., K. McGarigal, and B. W. Compton. 2007. Timing of large movement events of pond-breeding
amphibians in western Massachusetts, USA. Biological Conservation 136:442-454.
Tiner, R. W. 2003a. Dichotomous keys and mapping codes for wetland landscape position, landform,
water flow path, and waterbody type descriptors. U.S Fish and Wildlife Service, National Wetlands
Inventory Program, Northeast Region, Hadley, MA.
Tiner, R. W. 2003b. Estimated extentof geographically isolated wetlands in selected areas of the United
States. Wetlands 23:636-652.
Tiner, R. W. 2003c. Geographically isolated wetlands of the United States. Wetlands 23:494-516.
Tiner, R. W. 2011. Dichotomous keys and mapping codes for wetland landscape position, landform,
water flow path, and waterbody type descriptors: Version 2.0. U.S. Fish and Wildlife Service,
National Wetlands Inventory Program, Northeast Region, Hadley, MA.
Tischendorf, L, and L. Fahrig. 2000. On the usage and measurement of landscape connectivity. Oikos
90:7-19.
Tockner, K., F. Malard, and J. V. Ward. 2000. An extension of the flood pulse concept. Hydrological
Processes 14:2861-2883.
Tockner, K., M. Pusch, D. Borchardt, and M. S. Lorang. 2010. Multiple stressors in coupled river-
floodplain ecosystems. Freshwater Biology 55 (Suppl. 1):135-151.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Tockner, K., F. Schiemer, C. Baumgartner, G. Kum, E. Weigand, I. Zweimiiller, and J. V. Ward. 1999. The
Danube restoration project: Species diversity patterns across connectivity gradients in the
floodplain system. Regulated Rivers: Research & Management 15:245-258.
Tomer, M. D., C. G. Wilson, T. B. Moorman, K. J. Cole, D. Heer, and T. M. Isenhart. 2010. Source-pathway
separation of multiple contaminants during a rainfall-runoff event in an artificially drained
agricultural watershed. Journal of Environmental Quality 39:882-895.
Torgersen, C. E., R. N. Faux, B. A. Mclntosh, N. J. Poage, and D. J. Norton. 2001. Airborne thermal remote
sensing for water temperature assessment in rivers and streams. Remote Sensing of Environment
76:386-398.
Torgersen, C. E., R. E. Gresswell, D. S. Bateman, and K. M. Burnett. 2008. Spatial identification of tributary
impacts in river networks. Pages 159-181 in River confluences, tributaries and the fluvial network.
S. P. Rice, A. G. Roy, and B. L. Rhoads, editors. John Wiley & Sons, Chichester, UK.
Toth, J. 1963. A theoretical analysis of groundwater flow in small drainage basins. Journal of Geophysical
Research 68:4795-4812.
Toth, L. A., and A. van der Valk. 2012. Predictability of flood pulse driven assembly rules for restoration
of a floodplain plan community. Wetlands Ecology and Management 20:59-75.
Townsend-Small, A., D. E. Pataki, H. Liu, Z. Li, Q. Wu, and B. Thomas. 2013. Increasing summer river
discharge in southern California, USA, linked to urbanization. Geophysical Research Letters
40:4643-4647.
Townsend, C. R. 1989. The patch dynamics concept of stream community ecology. Journal of the North
American Benthological Society 8:36-50.
Trimble, S. W. 1997. Stream channel erosion and change resulting from riparian forests. Geology
25:467-469.
Trimble, S. W. 1999. Decreased rates of alluvial sediment storage in the Coon Creek basin, Wisconsin,
1975-93. Science 285:1244-1246.
Trimble, S. W. 2010. Streams, valleys and floodplains in the sediment cascade. Pages 307-343 in
Sediment cascades: An integrated approach. T. Burt and R. Allison, editors. John Wiley & Sons, West
Sussex, UK.
Triska, F. J., J. H. Duff, R. W. Sheibley, A. P. Jackman, and R. J. Avanzino. 2007. DIN retention-transport
through four hydrologically connected zones in a headwater catchment of the Upper Mississippi
River. Journal of the American Water Resources Association 43:60-71.
Tromp-van Meerveld, H. J., and J. J. McDonnell. 2006. Threshold relations in subsurface stormflow: 2. The
fill and spill hypothesis. Water Resources Research 42:W02411.
Tronstad, L. M., B. P. Tronstad, and A. C. Benke. 2007. Aerial colonization and growth: Rapid invertebrate
responses to temporary aquatic habitats in a river floodplain. Journal of the North American
Benthological Society 26:460-471.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Truscott, A. M., C. Soulsby, S. C. F. Palmer, L. Newell, and P. E. Hulme. 2006. The dispersal characteristics
of the invasive plant Mimulus guttatus and the ecological significance of increased occurrence of
high-flow events. Journal of Ecology 94:1080-1091.
Turton, D.}., C. T. Haan, and E. L. Miller. 1992. Subsurface flow responses of a small forested catchment
in the Ouachita Mountains. Hydrological Processes 6:111-125.
U.S. EPA. 2010. Causal Analysis/Diagnosis Decision Information System (CADDIS). U.S. Environmental
Protection Agency, Office of Research and Development, Washington, DC.
U.S. EPA. 2014. SAB review of the draft EPA report Connectivity of Streams and Wetlands to
Downstream Waters: A Review and Synthesis of the Scientific Evidence. EPA-SAB-15-001, U.S.
Environmental Protection Agency, Washington, DC.
Uden, D. R., M. L. Hellman, D. G. Angeler, and C. R. Allen. 2014. The role of reserves and anthropogenic
habitats for functional connectivity and resilience of ephemeral wetlands. Ecological Applications
24:1569-1582.
Ullrich, S. M., T. W. Tanton, and S. A. Abdrashitova. 2001. Mercury in the aquatic environment: A review
of factors affecting methylation. Critical Reviews in Environmental Science and Technology 31:241-
293.
Urban, D., and T. Keitt. 2001. Landscape connectivity: A graph-theoretic perspective. Ecology 82:1205-
1218.
Urban, N. R., S. E. Bayley, and S. J. Eisenreich. 1989. Export of dissolved organic carbon and acidity from
peatlands. Water Resources Research 25:1619-1628.
USAGE. 1987. Wetlands delineation manual. Technical Report Y-87-1, U.S. Army Corps of Engineers,
Waterways Experiment Station, Wetlands Research Program, Vicksburg, MS.
USAGE. 2009. National Inventory of Dams. U.S. Army Corps of Engineers.
Valett, H. M., M. A. Baker, J. A. Morrice, C. S. Crawford, M. C. Molles, C. N. Dahm, D. L. Moyer, J. R. Thibault,
and L. M. Ellis. 2005. Biogeochemical and metabolic responses to the flood pulse in a semiarid
floodplain. Ecology 86:220-234.
Valett, H. M., S. G. Fisher, and E. H. Stanley. 1990. Physical and chemical characteristics of the hyporheic
zone of a Sonoran Desert stream. Journal of the North American Benthological Society 9:201-215.
Van Looy, K., C. Cavillon, T. Tormos, J. Piffady, P. Landry, and Y. Souchon. 2013. A scale-sensitive
connectivity analysis to identify ecological networks and conservation value in river networks.
Landscape Ecology 28:1239-1249.
Van Meter, K. J., and N. B. Basu. In press. Signatures of human impact: Size distributions and spatial
organization of wetlands in the Prairie Pothole landscape. Ecological Applications:
doi.org/10.1890/1814-0662.1891.
Vannote, R. L., G. W. Minshall, K. W. Cummins, J. R. Sedell, and C. E. Gushing. 1980. The river continuum
concept. Canadian Journal of Fisheries and Aquatic Sciences 37:130-137.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Vanschoenwinkel, B., S. Gielen, M. Seaman, and L. Brendonck. 2009. Wind mediated dispersal of
freshwater invertebrates in a rock pool metacommunity: Differences in dispersal capacities and
modes. Hydrobiologia 635:363-372.
Ver Hoef, J. M., E. Peterson, and D. Theobald. 2006. Spatial statistical models that use flow and stream
distance. Environmental and Ecological Statistics 13:449-464.
Ver Hoef, J. M., and E. E. Peterson. 2010. A moving average approach for spatial statistical models of
stream networks. Journal of the American Statistical Association 105:6-18.
Vidon, P., C. Allan, D. Burns, T. P. Duval, N. Gurwick, S. Inamdar, R. Lowrance, J. Okay, D. Scott, and S.
Sebestyen. 2010. Hot spots and hot moments in riparian zones: Potential for improved water quality
management. Journal of the American Water Resources Association 46:278-298.
Vilizzi, L., B. J. McCarthy, 0. Scholz, C. P. Sharpe, and D. B. Wood. 2013. Managed and natural inundation:
Benefits for conservation of native fish in a semi-arid wetland system. Aquatic Conservation: Marine
and Freshwater Ecosystems 23:37-50.
Villar, C. A., L. de Cabo, P. Vaithiyanathan, and C. Bonetto. 2001. Litter decomposition of emergent
macrophytes in a floodplain marsh of the Lower Parana River. Aquatic Botany 70:105-116.
Vining, K. C. 2002. Simulation of streamflow and wetland storage, Starkweather Coulee subbasin, North
Dakota. Water years 1981-98. U.S.G.S. Water-Resources Investigations Report 02-4113, U.S.
Department of the Interior, U.S. Geological Survey in cooperation with the North Dakota State Water
Commission, Bismarck, ND.
Vivoni, E., R. S. Bowman, R. L. Wyckoff, R. T. Jakubowski, and K. E. Richards. 2006. Analysis of a monsoon
flood event in an ephemeral tributary and its downstream hydrologic effects. Water Resources
Research 42:W03404.
Volkmar, E. C., and R. A. Dahlgren. 2006. Biological oxygen demand dynamics in the lower San Joaquin
River, California. Environmental Science & Technology 40:5653-5660.
Voos, G., and P. M. Groffman. 1996. Relationships between microbial biomass and dissipation of 2,4-D
and dicambain soil. Biology and Fertility of Soils 24:106-110.
Vorb'smarty, C. J., B. M. Fekete, M. Meybeck, and R. B. Lammers. 2000. Global system of rivers: Its role in
organizing continental land mass and defining land-to-ocean linkages. Global Biogeochemical Cycles
14:599-621.
Vrtiska, M. P., and S.Sullivan. 2009. Abundance and distribution of lesser snow and Ross's geese in the
rainwater basin and Central Platte River Valley of Nebraska. Great Plains Research 19:147-155.
Wagener, T., M. Sivapalan, P. Troch, and R. Woods. 2007. Catchment classification and hydrologic
similarity. Geography Compass 1:901-931.
Wainwright, J., L. Turnbull, T. G. Ibrahim, I. Lexartza-Artza, S. F. Thornton, and R. E. Brazier. 2011.
Linking environmental regimes, space and time: Interpretations of structural and functional
connectivity. Geomorphology 126:387-404.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Waits, E. R., M. J. Bagley, M. J. Blum, F. H. McCormick, and J. M. Lazorchak. 2008. Source-sink dynamics
sustain central stonerollers [Campostoma anomaluni) in a heavily urbanized catchment. Freshwater
Biology 53:2061-2075.
Wallace, J. B., T. F. Cuffney, J. R. Webster, G. J. Lugthart, K. Chung, and B. S. Goldowitz. 1991. Export of fine
organic particles from headwater streams: Effects of season, extreme discharges, and invertebrate
manipulation. Limnology and Oceanography 36:670-682.
Wallace, J. B., S. L. Eggert, J. L. Meyer, and J. R. Webster. 1997. Multiple trophic levels of a forest stream
linked to terrestrial litter inputs. Science 277:102-104.
Wallace, J. B., D. H. Ross, and J. L. Meyer. 1982. Seston and dissolved organic carbon dynamics in a
southern Appalachian stream. Ecology 63:824-838.
Wallace, J. B., and J. R. Webster. 1996. The role of macroinvertebrates in stream ecosystem function.
Annual Review of Entomology 41:115-139.
Wallace, J. B., J. R. Webster, S. L. Eggert, J. L. Meyer, and E. R. Siler. 2001. Large woody debris in a
headwater stream: Long-term legacies of forest disturbance. International Review of Hydrobiology
86:501-513.
Wallace, J. B., M. R. Whiles, S. Eggert, T. F. Cuffney, G. H. Lugthart, and K. Chung. 1995. Long-term
dynamics of coarse particulate organic matter in three Appalachian Mountain streams. Journal of the
North American Benthological Society 14:217-232.
Walling, D. E. 1983. The sediment delivery problem. Journal of Hydrology 65:209-237.
Wallis, E., R. Mac Nally, and P. S. Lake. 2008. A Bayesian analysis of physical habitat changes at tributary
confluences in cobble-bed upland streams of the Acheron River basin, Australia. Water Resources
Research 44:W11421.
Wallis, E., R. Mac Nally, and S. Lake. 2009. Do tributaries affect loads and fluxes of particulate organic
matter, inorganic sediment and wood? Patterns in an upland river basin in south-eastern Australia.
Hydrobiologia 636:307-317.
Walsh, C. J., T. D. Fletcher, and M. J. Burns. 2012. Urban stormwater runoff: A new class of environmental
flow problem. PLOS ONE 7:e45814.
Walsh, C. J., A. H. Roy, J. W. Feminella, P. D. Cottingham, P. M. Groffman, and R. P. Morgan. 2005. The
urban stream syndrome: Current knowledge and the search for a cure. Journal of the North
American Benthological Society 24:706-723.
Walters, D. M., K. M. Fritz, and R. R. Otter. 2008. The dark side of subsidies: Adult stream insects export
organic contaminants to riparian predators. Ecological Applications 18:1835-1841.
Walton, R., J. E. Davis, T. H. Martin, and R. S. Chapman. 1996. Hydrology of the black swamp wetlands on
the Cache River, Arkansas. Wetlands 16:279-287.
Wang, X., A. M. Melesse, M. E. McClain, and W.Yang. 2007. Water quality changes as a result of coalbed
methane development in a rocky mountain watershed. Journal of the American Water Resources
Association 43:1383-1399.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Wang, X., S. Shang, Z. Qu, T. Liu, A. M. Melesse, and W. Yang. 2010. Simulated wetland conservation-
restoration effects on water quantity and quality at watershed scale. Journal of Environmental
Management 91:1511-1525.
Waples, R. S. 2010. Spatial-temporal stratifications in natural populations and how they affect
understanding and estimation of effective population size. Molecular Ecology Resources 10:785-
796.
Ward, J. V. 1989. The four-dimensional nature of lotic ecosystems. Journal of the North American
Benthological Society 8:2-8.
Ward, J. V. 1998. Riverine landscapes: Biodiversity patterns, disturbance regimes, and aquatic
conservation. Biological Conservation 83:269-278.
Ward, J. V., C. T. Robinson, and K. Tockner. 2002a. Applicability of ecological theory to riverine
ecosystems. Verhandlungen der Internationale Vereinigung fur Theoretische und Angewandte
Limnologie 28:443-450.
Ward, J. V., and J. A. Stanford. 1983. The serial discontinuity concept of lotic ecosystems. Pages 29-42 in
Dynamics of lotic ecosystems. T. D. Fontaine and S. M. Bartell, editors. Ann Arbor Science, Ann Arbor,
MI.
Ward, J. V., K. Tockner, D. B. Arscott, and C. Claret. 2002b. Riverine landscape diversity. Freshwater
Biology 47:517-539.
Webb, B. W. 1996. Trends in stream and river temperature. Hydrological Processes 10:205-226.
Webb, E. B., L. M. Smith, M. P. Vrtiska, and T. G. Lagrange. 2010. Effects of local and landscape variables
on wetland bird habitat use during migration through the Rainwater Basin. Journal of Wildlife
Management 74:109-119.
Webster, J. R. 1983. The role of benthic macroinvertebrates in detritus dynamics of streams: A computer
simulation. Ecological Monographs 53:383-404.
Webster, J. R. 2007. Spiraling down the river continuum: Stream ecology and the U-shaped curve.
Journal of the North American Benthological Society 26:375-389.
Webster, J. R., E. F. Benfield, T. P. Ehrman, M. A. Schaeffer, J. L. Tank, J. J. Hutchens, and D. J. D'Angelo.
1999. What happens to allochthonous material that falls into streams? A synthesis of new and
published information from Coweeta. Freshwater Biology 41:687-705.
Webster, J. R., and J. L. Meyer. 1997. Stream organic matter budgets. Journal of the North American
Benthological Society 16:3-161.
Webster, J. R., and B. C. Patten. 1979. Effects of watershed perturbation on stream potassium and
calcium dynamics. Ecological Monographs 49:51-72.
Wellborn, G. A., D. K. Skelly, and E. E. Werner. 1996. Mechanisms creating community structure across a
freshwater habitat gradient. Annual Review of Ecology and Systematics 27:337-363.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Weller, M. W. 1964. Distribution and migration of the redhead. The Journal of Wildlife Management:64-
103.
Wemple, B. C., J. A. Jones, and G. E. Grant. 1996. Channel network extension by logging roads in two
basins, Western Cascades, Oregon. Journal of the American Water Resources Association 32:1195-
1207.
Wemple, B. C., F. J. Swanson, and J. A. Jones. 2001. Forest roads and geomorphic process interactions,
Cascade Range, Oregon. Earth Surface Processes and Landforms 26:191-204.
Westbrook, C. J., D. J. Cooper, and B. W. Baker. 2006. Beaver dams and overbank floods influence
groundwater-surface water interactions of a Rocky Mountain riparian area. Water Resources
Research 42:WO6404.
Western, A. W., G. Blb'schl, and R. B. Grayson. 2001. Toward capturing hydrologically significant
connectivity in spatial patterns. Water Resources Research 37:83-97.
Wetzel, R. G. 1992. Gradient-dominated ecosystems: Sources and regulatory functions of dissolved
organic matter in freshwater ecosystems. Hydrobiologia 229:181-198.
Wetzel, R. G., and B. A. Manny. 1972. Decomposition of dissolved organic carbon and nitrogen
compounds from leaves in an experimental hard-water stream. Limnology and Oceanography
17:927-931.
Wharton, C. H., W. M. Kitchens, E. C. Pendleton, and T. W. Sipe. 1982. The ecology of bottomland
hardwood swamps of the Southeast: A community profile. FWS/OBS-81/37, U.S. Department of the
Interior, U.S. Fish and Wildlife Service, Biological Services Program, Washington, DC.
Whigham, D. F., C. Chitterling, and B. Palmer. 1988. Impacts of fresh-water wetlands on water-quality: A
landscape perspective. Environmental Management 12:663-671.
Whigham, D. F., and T. E. Jordan. 2003. Isolated wetlands and water quality. Wetlands 23:541-549.
Whiles, M. R., and W. K. Dodds. 2002. Relationships between stream size, suspended particles, and filter-
feeding macroinvertebrates in a Great Plains drainage network. Journal of Environmental Quality
31:1589-1600.
Whited, D. C., M. S. Lorang, M. J. Harner, R. Hauer, J. S. Kimball, and J. A. Stanford. 2007. Climate,
hydrologic disturbance, and succession: Drivers of floodplain pattern. Ecology 88:940-953.
Whiteley, A. R., K. Hastings, J. K. Wenburg, C. A. Frissell, J. C. Martin, and F. W. Allendorf. 2010. Genetic
variation and effective population size in isolated populations of coastal cutthroat trout.
Conservation Genetics 11:1929-1943.
Whiting, P. J., and M. Pomeranets. 1997. A numerical study of bank storage and its contribution to
streamflow. Journal of Hydrology 202:121-136.
Whitmire, S. L, and S. K. Hamilton. 2008. Rates of anaerobic microbial metabolism in wetlands of
divergent hydrology on a glacial landscape. Wetlands 28:703-714.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Wiens, J. A. 2002. Riverine landscapes: Taking landscape ecology into the water. Freshwater Biology
47:501-515.
Wigington, P.}.,}. L. Ebersole, M. E. Colvin, S. G. Leibowitz, B. Miller, B. Hansen, H. R. LaVigne, D. White, J.
P. Baker, M. R. Church, J. R. Brooks, M. A. Cairns, and J. E. Compton. 2006. Coho salmon dependence
on intermittent streams. Frontiers in Ecology and the Environment 4:513-518.
Wigington, P. J., S. M. Griffith, J. A. Field, J. E. Baham, W. R. Horwath, J. Owen, J. H. Davis, S. C. Rain, and J. J.
Steiner. 2003. Nitrate removal effectiveness of a riparian buffer along a small agricultural stream in
western Oregon. Journal of Environmental Quality 32:162-170.
Wigington, P. J., S. G. Leibowitz, R. L. Comeleo, and J. L. Ebersole. 2013. Oregon hydrologic landscapes: A
classification framework. Journal of the American Water Resources Association 49:163-182.
Wigington, P. J., T. J. T.J. Moser, and D. R. Lindeman. 2005. Stream network expansion: A riparian water
quality factor. Hydrological Processes 19:1715-1721.
Wigmosta, M. S., and W. A. Perkins. 2001. Simulating the effects of forest roads on watershed hydrology.
Pages 127-143 in Land use and watersheds: Human influence on hydrology and geomorphology in
urban and forest areas. M. S. Wigmosta and S. J. Burges, editors. American Geophysical Union,
Washington, DC.
Wilcock, P., J. Pitlick, and Y. Cui. 2009. Sediment transport primer estimating bed-material transport in
gravel-bed rivers. General Technical Report RMRS-GTR-226, U.S. Department of Agriculture, U.S.
Forest Service, Rocky Mountain Research Station, Fort Collins, CO.
Wilcox, B. P., D. D. Dean, J. S. Jacob, and A. Sipocz. 2011. Evidence of surface connectivity for Texas Gulf
Coast depressional wetlands. Wetlands 31:451-458.
Wilkes, G., J. Brassard, T. A. Edge, V. Gannon, C. C. Jokinen, T. H. Jones, N. Neumann, K. D. M. Pintar, N.
Ruecker, P. J. Schmidt, M. Sunohara, E. Topp, and D. R. Lapen. 2013. Bacteria, viruses, and parasites
in an intermittent stream protected from and exposed to pasturing cattle: Prevalence, densities, and
quantitative microbial risk assessment. Water Research 47:6244-6257.
Williams, D. D. 1996. Environmental constraints in temporary fresh waters and their consequences for
the insect fauna. Journal of the North American Benthological Society 15:634-650.
Williams, G. P., and M. G. Wolman. 1984. Downstream effects of dams on alluvial rivers. USGS
Professional Paper 1286, U.S. Department of the Interior, U.S. Geological Survey, Washington, DC.
Wilson, C. 0., and Q. Weng. 2011. Simulating the impacts of future land use and climate changes on
surface water quality in the Des Plaines River watershed, Chicago Metropolitan Statistical Area,
Illinois. Science of The Total Environment 409:4387-4405.
Wilson, J. L., and H. Guan. 2004. Mountain-block hydrology and mountain-front recharge. Pages 113-137
in Groundwater recharge in a desert environment: The southwestern United States. F. M. Phillips, J.
Hogan, and B. R. Scanlon, editors. American Geophysical Union, Washington, DC.
Wilzbach, M. A., and K. W. Cummins. 1989. An assessment of short-term depletion of stream
macroinvertebrate benthos by drift. Hydrobiologia 185:29-39.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Wilzbach, M. A., K. W. Cummins, and J. D. Hall. 1986. Influence of habitat manipulations on interactions
between cutthroat trout and invertebrate drift. Ecology 67:898-911.
Winston, M. R., C. M. Taylor, and J. Pigg. 1991. Upstream extirpation of four minnow species due to
damming of a prairie stream. Transactions of the American Fisheries Society 120:98-105.
Winter, T. C. 1999. Relation of streams, lakes, and wetlands to groundwater flow systems. Hydrogeology
Journal 7:28-45.
Winter, T. C. 2001. The concept of hydrologic landscapes. Journal of the American Water Resources
Association 37:335-349.
Winter, T. C. 2007. The role of groundwater in generating streamflow in headwater areas and in
maintaining base flow. Journal of the American Water Resources Association 43:15-2 5.
Winter, T. C., J. W. Harvey, 0. L. Franke, and W. M. Alley. 1998. Ground water and surface water: A single
resource. USGS Circular 1139, U.S. Department of the Interior, U.S. Geological Survey, Denver, CO.
Winter, T. C., and J. W. LaBaugh. 2003. Hydrologic considerations in defining isolated wetlands. Wetlands
23:532-540.
Winter, T. C., and D. 0. Rosenberry. 1998. Hydrology of prairie pothole wetlands during drought and
deluge: A 17-year study of the cottonwood lake wetland complex in North Dakota in the perspective
of longer term measured and proxy hydrological records. Climatic Change 40:189-209.
Winter, T. C., D. 0. Rosenberry, and J. W. LaBaugh. 2003. Where does the ground water in small
watersheds come from? Ground Water 41:989-1000.
Wipfli, M. S., and D. P. Gregovich. 2002. Export of invertebrates and detritus from fishless headwater
streams in southeastern Alaska: Implications for downstream salmonid production. Freshwater
Biology 47:957-969.
Wipfli, M. S., and J. Musslewhite. 2004. Density of red alder (Alnus rubrd) in headwaters influences
invertebrate and detritus subsidies to downstream fish habitats in Alaska. Hydrobiologia 520:153-
163.
Wipfli, M. S., J. S. Richardson, and R. J. Naiman. 2007. Ecological linkages between headwaters and
downstream ecosystems: Transport of organic matter, invertebrates, and wood down headwater
channels. Journal of the American Water Resources Association 43:72-85.
Wiskow, E., and R. R. van der Ploeg. 2003. Calculation of drain spacings for optimal rainstorm flood
control. Journal of Hydrology 272:163-174.
With, K. A., R. H. Gardner, and M. G. Turner. 1997. Landscape connectivity and population distributions
in heterogeneous environments. Oikos 78:151-169.
Withers, P. J. A., and H. P. Jarvie. 2008. Delivery and cycling of phosphorus in rivers: A review. Science of
The Total Environment 400:379-395.
Woessner, W. W. 2000. Stream and fluvial plain ground water interactions: Rescaling hydrogeologic
thought. Ground Water 38:423-429.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Wofford, J. E. B., R. E. Gresswell, and M. A. Banks. 2005. Influence of barriers to movement on within-
watershed genetic variation of coastal cutthroat trout. Ecological Applications 15:628-637.
Wohl, E. 2005. Disconnected rivers: Human impacts to rivers in the United States. Reviews in
Engineering Geology 16:19-34.
Wohl, E., and N. D. Beckman. 2014. Leaky rivers: Implications of the loss of longitudinal fluvial
disconnectivity in headwater streams. Geomorphology 205:27-35.
Wollheim, W. M., B. J. Peterson, S. M. Thomas, C. H. Hopkinson, and C. J. Vorosmarty. 2008. Dynamics of N
removal over annual time periods in a suburban river network. Journal of Geophysical Research:
Biogeosciences 113:G03038.
Wolman, M. G., and J. P. Miller. 1960. Magnitude and frequency offerees in geomorphic processes.
Journal of Geology 68:54-74.
Wolock, D. M., T. C. Winter, and G. McMahon. 2004. Delineation and evaluation of hydrologic-landscape
regions in the United States using geographic information system tools and multivariate statistical
analysis. Environmental Management 34:S71-S88.
Wood, P. J., and P. D. Armitage. 1997. Biological effects of fine sediment in the lotic environment
Environmental Management 21:203-217.
Woodford, D. J., and A. R. Mclntosh. 2010. Evidence of source-sink metapopulations in a vulnerable
native galaxiid fish driven by introduced trout. Ecological Applications 20:967-977.
Woodin, M. 1994. Use of saltwater and freshwater habitats by wintering redheads in southern Texas.
Pages 279-287 in Aquatic birds in the trophic web of lakes. J. Kerekes, editor. Springer Netherlands.
Woodward, G. U. Y., and A. G. Hildrew. 2002. Food web structure in riverine landscapes. Freshwater
Biology 47:777-798.
Wright, S. 1943. Isolation by distance. Genetics 28:114-138.
Yamamoto, S., K. Morita, I. Koizumi, and K. Maekawa. 2004. Genetic differentiation of white-spotted
charr [Salvelinus leucomaenis) populations after habitat fragmentation: Spatial-temporal changes in
gene frequencies. Conservation Genetics 5:529-538.
Yang, J., and X. Chu. 2013. Effects of DEM resolution on surface depression properties and hydrologic
connectivity. Journal of Hydrologic Engineering 18:1157-1169.
Yang, W., X. Wang, Y. Liu, S. Gabor, L. Boychuk, and P. Badiou. 2010. Simulated environmental effects of
wetland restoration scenarios in a typical Canadian prairie watershed. Wetlands Ecology and
Management 18:269-279.
Yetter, S. E. 2013. Freshwater macroinvertebrates of the Mid-Atlantic Region. Pages 339-379 in Mid-
Atlantic freshwater wetlands: Advances in wetlands science, management, policy, and practice. R. P.
Brooks and D. H. Wardrop, editors. Springer, New York.
Stream and Wetland Connectivity: January 2015
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Zaimes, G. N., R. C. Schultz, and T. M. Isenhart. 2004. Stream bank erosion adjacent to riparian forest
buffers, row-crop fields, and continuously-grazed pastures along Bear Creek in central Iowa. Journal
of Soil and Water Conservation 59:19-27.
Zedler, P. H. 2003. Vernal pools and the concept of "isolated wetlands". Wetlands 23:597-607.
Zimmer, K. D., M. A. Hanson, and M. G. Butler. 2001. Effects of fathead minnow colonization and removal
on a prairie wetland ecosystem. Ecosystems 4:346-357.
Zwieniecki, M. A., and M. Newton. 1999. Influence of streamside cover and stream features on
temperature trends in forested streams of western Oregon. Western Journal of Applied Forestry
14:106-113.
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APPENDIX A. GLOSSARY
Absorption—A reversible process that occurs when molecules in one state or phase penetrate those of
another phase.
Adsorption—Adhesion of molecules to a surface, either physically or chemically. Physical adsorption
occurs when the surface tension of a solid causes molecules to be held at its surface; this process can be
reversible, depending on environmental conditions. Chemical adsorption occurs when chemicals bond
at the surface of a solid, and is not readily reversible.
Allochthonous—Describing organic material that originates from outside of streams, rivers, wetlands,
or lakes (e.g., terrestrial plant litter, soil).
Alluvial Aquifer—An aquifer with geologic materials deposited by a stream or river (alluvium) that
retains a hydraulic connection with the depositing stream.
Alluvial Deposits—See Alluvium.
Alluvial Ground Water—Ground water occurring in an alluvial aquifer.
Alluvium—Deposits of clay, silt, sand, gravel, or other particulate materials that have been deposited by
a stream or other body of running water in a streambed, on a flood plain, on a delta, or at the base of a
mountain. See Colluvium.
Anastomosing Channel—A multithreaded stream or river channel in which the channels
(distributaries) branch and rejoin farther downstream; distributary channels are separated by stable
islands (usually vegetated) that are large relative to the size of the channels.
Anoxic Conditions—Without detectable dissolved oxygen; anaerobic. See Hypoxia.
Aquatic Ecosystem—Any aquatic environment, including all of the environment's living and nonliving
constituents and the interactions among them.
Aquifer—A geologic formation (e.g., soil, rock, alluvium) with permeable materials partially or fully
saturated with ground water that yields ground water to a well, spring, or stream.
Artificial Drainage—Use of constructed channels or subsurface structures to drain an area by
increasing the rate of flow of water from the area.
Assimilatory Processes—The incorporation or transformation of simple compounds into more
complex compounds.
Autochthonous—Describing organic matter that originates from production within streams, rivers,
wetlands, or lakes (e.g., periphyton, macrophytes, phytoplankton).
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Bank Storage—Storage of water that flows from a stream to an alluvial aquifer during a flood or period
of high streamflow. The volume of water is stored and released after the high-water event over days to
months. The volume of water stored and the timing of release depends on the hydraulic properties of
the alluvial aquifer.
Baseflow—Sustained flow of a stream (or river) in the absence of stormflow (direct runoff). Natural
baseflow is sustained by ground-water discharge in the stream network. Baseflow also can be sustained
by human sources (e.g., irrigation recharges to ground water).
Basin—See Drainage Basin.
Bedrock—Solid rock underlying loose deposits such as soil or alluvium.
Bog—A peat-accumulating wetland that is generally nutrient poor.
Braided Channel—A multithreaded channel in which the channels (distributaries) branch and rejoin
farther downstream and the channels are separated by mobile, transient bars (poorly vegetated) that
are small relative to the size of the channels.
Carolina Bays—Elliptical, ponded, depressional wetlands that range along the Atlantic Coastal Plain
from northern Florida to New Jersey. See Delmarva Bays.
Catchment—The area drained by a stream, river, or other water body; typically defined by the
topographic divides between one water body and another. Synonymous with Watershed and Drainage
Basin.
Channel—A natural or constructed passageway or depression of perceptible linear extent that conveys
water and associated material downgradient.
Channelization—A type of artificial drainage in which complex channels are straightened to increase
the rate of water flow from an area.
Channelized Flow—Flow that occurs in a natural or artificial channel.
Colluvium—A layer of unconsolidated soils, sediment and rock fragments deposited by surface runoff
and gravitational processes; colluvium generally occurs as a blanket of poorly sorted sediment and rock
fragments on the lower parts of hillslopes underlain by bedrock. See Alluvium.
Condition—General health or quality of an ecosystem, typically assessed using one or more indicators.
Confined Aquifer—An aquifer bounded above and below by confining units of distinctly lower
permeability than that of the aquifer itself.
Confluence—The point at which two stream channels intersect to form a single channel.
Connectivity—The degree to which components of a river system are joined, or connected, by various
transport mechanisms; connectivity is determined by the characteristics of both the physical landscape
and the biota of the specific system.
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Connectivity Descriptors (for streams and wetlands)—The frequency, duration, magnitude, timing,
and rate of change effluxes to and biological exchanges with downstream waters.
Contributing Area—Location within a watershed/river network that serves as a source of stream flow
or material flux.
Contaminants—Any material that might be harmful to humans or other organisms when released to the
environment.
Deep Ground Water—Ground-water flow systems having the deepest and longest flowpaths; also
referred to as regional ground-water flow systems, they can occur beneath local and intermediate
ground-water flow systems. See Local Ground Water, Regional Ground Water.
Delmarva Bays—Carolina bays that are geographically specific to the Delmarva Peninsula. These
wetlands frequently have the same elliptical shape and orientation as Carolina bays. See Carolina Bays.
Dendritic Stream Network—A stream network pattern of branching tributaries (see Figure 2-19B).
Depressional Wetland—A wetland occupying a topographic low point that allows the accumulation of
surface water. Depressional wetlands can have any combination of inlets and outlets or lack them
completely. Examples include kettles, prairie potholes, and Carolina bays. This category also includes
slope wetlands (wetlands associated with surface discharge of ground water or saturated overflow with
no channel formation).
Diadromous—Migratory between fresh and salt waters.
Direct Runoff—Runoff that occurs in direct response to precipitation. See Stormflow.
Discharge—The volume of water (surface water or ground water) that passes a given location over a
given period of time; the rate of runoff. Often expressed as ft3 s~l or m3 s~l.
Discontinuous Flow—Refers to stream and river reaches that have flow in one part of the reach but not
another part of the reach. See Reach.
Dispersal—Movement from natal breeding sites to new breeding sites.
Drainage Area—The spatial extent of a drainage basin. Typically expressed in mi2 or km2.
Drainage Basin—The area drained by a stream, river, or other water body; typically defined by the
topographic divides between one water body and another. Synonymous with Catchment one? Watershed.
Drainage Density—The total length of stream channels per unit drainage area (e.g., per mi2 or km2).
Drainage Network—See River Network.
Egg Bank—Viable dormant eggs that accumulate in soil or in sediments under water. See Seed bank.
Endorheic Basins—A closed drainage basin with no outflows to other water bodies.
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Endorheic Stream—A stream or river reach that experiences a net loss of water to a ground-water
system. See Losing Stream or Wetland.
Ephemeral Stream—A stream or river that flows briefly in direct response to precipitation; these
channels are always above the water table.
Eutrophication—Natural or artificial enrichment of a water body by nutrients, typically phosphates and
nitrates. If enrichment leads to impairment (e.g., toxic algal blooms), eutrophication is a form of
pollution.
Evapotranspiration—The combined loss of water to the atmosphere due to evaporation and
transpiration losses. Transpiration is the loss of water vapor to air by plants.
Fen—A peat-accumulating wetland characterized by mineral-rich water inputs.
Flood—The occurrence of stream or river flow of such magnitude that it overtops the natural or artificial
banks in a reach of the stream or river; where a floodplain exists, a flood is any flow that spreads over or
inundates the floodplain. Floods also can result from rising stages in lakes and other water bodies.
Flood (100-year)—Flood level (stage or discharge) with a 1% probability of being equaled or exceeded
in a given year.
Flood Flows—Discharge or flow of sufficient (or greater) magnitude to cause a flood.
Flood Stage—The stage at which streams or rivers overtop their natural or artificial banks.
Floodwater—Water associated with a flood event.
Floodplain—A level area bordering a stream or river channel that was built by sediment deposition
from the stream or river under present climatic conditions and is inundated during moderate to high
flow events. Floodplains formed under historic or prehistoric climatic conditions can be abandoned by
rivers and form terraces.
Floodplain Wetland—Portions of floodplains that meet the Cowardin etal. (1979) three-attribute
definition of a wetland (i.e., having wetland hydrology, hydrophytic vegetation, or hydric soils). See
Wetland.
Flow—Water movement above ground or below ground.
Flow Duration Class—A classification that assigns streamflow duration to ephemeral, intermittent, or
perennial classes.
Flow Regime—Descriptor of flow types in a temporal or magnitude sense (i.e., slow-flow regime, low-
flow regime)
Flowpath—See Hydrologic Flowpath.
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Fluvial—Refers to or pertains to streams; e.g., stream processes (fluvial processes), fluvial landforms,
such as fluvial islands and bars, and biota living in and near stream channels.
Flux—Flow of materials between system components per unit time.
Gaining Stream or Wetland—A wetland or a stream or river reach that experiences a net gain of water
from ground water (see Figure 2-5). In this situation, the water table elevation near the stream or
wetland is higher than the stream or wetland water surface. Conditions conducive to losing or gaining
streams and wetlands can change over short distances within river networks and river basins. See
Losing Stream or Wetland.
Geographically Isolated Wetland—A wetland that is completely surrounded by uplands; for example,
hydrophytic plant communities surrounded by terrestrial plant communities or undrained hydric soils
surrounded by nonhydric soils. This term often is mistakenly understood to mean hydrologically
isolated. Geographically isolated wetlands vary in their degree of hydrologic and biotic connectivity.
Ground Water—Any water that occurs and flows in the saturated zone. See Saturated Zone.
Ground-water Discharge —The flow of ground water to surface waters; discharge areas occur where
the water tables intersect land surfaces. See Seep, Spring
Ground-water Discharge Wetland—A wetland that receives ground-water discharge.
Ground-water Flow—Flow of water in the subsurface saturated zone.
Ground-water Flow-through Wetland—A wetland that has both ground-water inputs and outputs.
Ground water enters the wetland through the upgradient direction and exits the wetland downgradient.
Ground-water Recharge—The process by which ground water is replenished; a recharge area occurs
where precipitation or surface water infiltrates and is transmitted downward to the saturated zone
(aquifer). See Infiltration, Percolation, Transmission.
Ground-water Recharge Wetland—A wetland that recharges ground water.
Ground-water Reservoir—A saturated body of ground water having loosely definable spatial limits.
Ground-water System—Reference to the ground water and geologic materials comprising the saturated
zone; the ground-water system, as a whole, is a three-dimensional flow field.
Ground water-Surface water Interactions—Movement of water between surface-water bodies and
ground-water systems. Flows can occur in either direction.
Ground-water Withdrawal—Pumping of water from aquifers for human uses.
Habitat—Environment (place and conditions) in which organisms reside.
Headwater—Areas from which water originates within a river or stream network. This term typically
refers to stream channels but can also describe wetlands or open waters, such as ponds.
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Headwater Stream—Headwater streams are first- to third-order streams. Headwater streams can be
ephemeral, intermittent, or perennial. See Stream Order, Flow Duration Class.
Hillslope—A sloping segment of land surface.
Hydraulic Conductivity—A measure of the permeability of a porous medium. For a given hydraulic
gradient, water moves more rapidly through media with high hydraulic conductivity than low hydraulic
conductivity.
Hydraulic Gradient—Slope of the water table. See Water Table.
Hydraulic Head—The height above a standard datum of the surface of a column of water that can be
supported by the static pressure at a given point; for a well, the hydraulic head is the height of the water
level in the well compared to a datum elevation.
Hydraulics—The physics of water in its liquid state.
Hydric—An area, environment, or habitat that is generally very wet with plenty of moisture. See Mesic,
Xeric.
Hydrograph—A graph of stream or river discharge over time. Stage or water table elevation also can be
plotted.
Hydrologic Event—An increase in streamflow resulting from precipitation or snowmelt.
Hydrologic Flowpath—The pathway that water follows as it moves over the watershed surface or
through the subsurface environment.
Hydrology—The study of the properties, distribution, and effects of water as a liquid, solid, and gas on
Earth's surface, in the soils and underlying rocks, and in the atmosphere.
Hydrologic Landscape—A landscape with a combination of geology, soils, topography, and climate that
has characteristic influences on surface water and ground water.
Hydrologic Permanence—The frequency and duration of streamflow in channels or the frequency and
duration of standing water in wetlands.
Hyporheic Flow—Water from a stream or river channel that enters subsurface materials of the
streambed and bank and then returns to the stream or river.
Hyporheic Exchange—Water and solutes exchanged between a surface channel and the shallow
subsurface. See Hyporheic Flow.
Hyporheic Zone—The area adjacent to and beneath a stream or river in which hyporheic flow occurs.
The dimensions of the hyporheic zone are controlled by the distribution and characteristics of alluvium
and hydraulic gradients between streams and local ground water.
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Hypoxia—The condition in which dissolved oxygen is below the level necessary to sustain most animal
life. See Anoxic Conditions.
Infiltration—The downward entry of water from the land surface into the subsurface.
Infiltration Capacity—The maximum rate at which infiltration can occur at a given location.
Interfluve—The area of higher terrain between adjacent stream valleys.
Intermediate Ground Water—Ground-water flow systems representative of the wide range of
flowpath lengths and depths that occur between local and regional ground-water systems.
Intermittent—This term also can be applied to other surface-water bodies and ground-water flow or
level. See Intermittent Stream.
Intermittent Stream—A stream or portion of a stream that flows continuously only at certain times of
year; for example, when it receives water from a spring, ground-water source, or a surface source such
as melting snow. At low flow, dry segments alternating with flowing segments can be present.
Inundation—To cover dry land with floodwaters.
Isolation—Condition defined by reduced or nonexistent transport mechanisms between system
components.
Isotopic Tracer—See Stable Isotope Tracer.
Lag Function—Any function within a stream or wetland that provides temporary storage and
subsequent release of materials without affecting cumulative flux (exports = imports); delivery is
delayed and can be prolonged.
Lateral Source Stream—A first-order stream that flows into a higher order stream.
Lentic—Of, relating to, or living in still water. See Lotic.
Levee (Artificial)—An engineered structure built next to a stream or river from various materials to
prevent flooding of surrounding areas. The levee raises the elevation of the channel height to convey
greater discharge of water without flooding.
Levee (Natural)—A broad, low ridge or embankment of coarse silt and sand that is deposited by a
stream on its floodplain and along either bank of its channel. Natural levees are formed by reduced
velocity of flood flows as they spill onto floodplain surfaces and can no longer transport the coarse
fraction of the suspended sediment load.
Local Ground Water—Ground water with a local flow system. Water that recharges at a high point in
the water table that discharges to a nearby lowland. Local ground-water flow is the most dynamic and
shallowest of ground-water flow systems. Therefore, it has the greatest interchange with surface water.
Local flow systems can be underlain by intermediate and regional flow systems. Water in these deeper
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flow systems have longer flowpaths and longer contact time with subsurface materials. Deeper flow
systems also eventually discharge to surface waters and influence their condition.
Losing Stream or Wetland—A stream, wetland, or river reach that experiences a net loss of water to a
ground-water system (see Figure 2-5). In this situation, the water table elevation near the stream or
wetland is lower than the stream or wetland water surface. Conditions conducive to losing or gaining
streams and wetlands can change over short distances within river networks and river basins. See
Gaining Stream or Wetland.
Lotic—Of, relating to, or living in moving water. See Lentic.
Mainstem—Term used to distinguish the larger (in terms of discharge) of two intersecting channels in a
river network.
Materials—Any physical, chemical, or biological entity, including but not limited to water, heat energy,
sediment, wood, organic matter, nutrients, chemical contaminants, and organisms.
Meltwater—Liquid water that results from the melting of snow, snowpacks, ice, or glaciers.
Mesic—An area, environment, or habitat with a moderate amount of moisture. See Hydric, Xeric.
Migration—Long-distance movements undertaken by organisms on a seasonal basis.
Non-floodplain Wetland—An area outside of the floodplain that meets the Cowardin et al. (1979)
three-attribute definition of a wetland (i.e., having wetland hydrology, hydrophytic vegetation, or hydric
soils). For the purposes of this report, riparian wetlands that occur outside of the floodplain are not
included as non-floodplain wetlands, since these wetlands are subject to bidirectional, lateral hydrologic
flows. See Floodplain, Wetland.
Nutrients (In Aquatic Systems)—Elemental forms of nitrogen, phosphorus, and trace elements,
including sulfur, potassium, calcium, and magnesium, that are essential for the growth of organisms but
can be contaminants when present in high concentrations.
Nutrient Spiraling—Longitudinal cycles ("spirals") of nutrient uptake and release along the stream or
river continuum. The spirals are created as aquatic organisms consume, transform, and regenerate
nutrients, altering the rates of nutrient transport to downstream waters.
Open-channel Flow—Water flowing within natural or artificial channels.
Open Waters—Nontidal lentic water bodies such as lakes and oxbow lakes that are frequently small or
shallow.
Overbank Flow—Streamflow that overtops a stream or river channel.
Overland Flow—The portion of streamflow derived from net precipitation that fails to infiltrate the land
surface at any point and runs over the surface to the nearest stream channel.
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Oxbow Lakes—Water bodies that originate from the cutoff meanders of rivers; such lakes are common
in floodplains of large rivers.
Peatland—A wetland that accumulates partially decayed organic matter. Fens and bogs are common
examples.
Perched Ground Water—Unconfined ground water separated from an underlying body of ground
water by an unsaturated zone; perched ground water is supported by a perching layer (bed) for which
the permeability is so low that water percolating downward to the underlying unsaturated zone is
restricted.
Perching Water Tables—See Perched Ground Water.
Percolation—The downward movement of water through soil or rock formations.
Perennial— See Perennial Stream. This term can be applied to other surface-water bodies and to
ground-water flow or level.
Perennial Stream—A stream or portion of a stream that flows year-round and is maintained by local,
intermediate, or regional ground-water discharge or flow from higher in the river network.
Permanent Waters—Water bodies that contain water year-round; perennial waters.
Permeability—Property of a porous medium that enables it to transmit fluids under a hydraulic
gradient. For a given hydraulic gradient, water will move more rapidly through high permeability
materials than low permeability materials.
Phreatophyte—Plants that use water from the saturated zone.
Potential Evapotranspiration—The amount of water that would be lost to the atmosphere over a given
area through evaporation and transpiration, assuming no limits on the water supply. See
Evapotranspiration.
Potentiometric Surface—The surface representing the level to which ground water will rise in a well
penetrating a confined aquifer.
Prairie Potholes—Complex of glacially formed wetlands, usually lacking natural outlets, found in the
central United States and Canada.
Precipitation—Water that condenses in the atmosphere and falls to a land surface. Common types
include rain, snow, hail, and sleet.
Precipitation Intensity—The rate at which precipitation occurs; generally refers to rainfall intensity.
Primary Production—The fixation of inorganic carbon into organic carbon (e.g., plant and algae
biomass) through the process of photosynthesis. Primary production is the first level of the food web,
and provides most of the autochthonous carbon produced in ecosystems. The rate of fixation is referred
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to as gross primary productivity (GPP) or net primary productivity (NPP), where NPP is equal to GPP
minus respiration. See Respiration, Secondary Production.
Propagule—Any part of an organism that can give rise to a new individual organism. Seeds, eggs, and
spores are propagules.
Reach—A length of stream channel with relatively uniform discharge, depth, area, and slope.
Recession [of Flow)—Decrease in flow following a hydrologic event.
Recharge Area—An area in which water infiltrates the surface and reaches the zone of saturation.
Refuge Function—The protective function of a stream or wetland that allows an organism (or material)
to avoid mortality (or loss) in a nearby sink area, thereby preventing the net decrease in material flux
that otherwise would have occurred (exports = imports). This term typically refers to organisms but can
be used for nonliving materials. See Sink Function.
Regional Ground Water—Ground water with a deep, regional-scale flow system; also referred to as
deep ground water. These flow systems can occur beneath local and intermediate ground-water flow
systems. See Local Ground Water, Deep Ground Water.
Respiration—The chemical process by which organisms break down organic matter and produce
energy for growth, movement, and other biological processes. Aerobic respiration uses oxygen and
produces carbon dioxide.
Return Flow—Water that infiltrates into a land surface and moves to the saturated zone and then
returns to the land surface (or displaces water that returns to the soil surface).
Riparian Areas—Transition areas or zones between terrestrial and aquatic ecosystems that are
distinguished by gradients in biophysical conditions, ecological processes, and organisms. They are
areas through which surface hydrology and subsurface hydrology connect water bodies with their
uplands. They include those portions of terrestrial ecosystems that significantly influence exchanges of
energy and matter with aquatic ecosystems. Riparian areas are adjacent to perennial, intermittent, and
ephemeral streams, lakes, and estuarine-marine shorelines. See Upland.
Riparian Wetland—Portions of riparian areas that meet the Cowardin etal. (1979) three-attribute
definition of a wetland (i.e., having wetland hydrology, hydrophytic vegetation, hydric soils). See
Wetland.
River—A relatively large volume of flowing water within a visible channel, including subsurface water
moving in the same direction as the surface water, and lateral flows exchanged with associated
floodplain and riparian areas. See Stream.
River Network—A hierarchical, interconnected population of channels or swales that drain water to a
river. Flow through these channels can be perennial, intermittent, or ephemeral.
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River Network Expansion/Contraction—The extent of flowing water in a river network increases
during wet seasons and large precipitation events and decreases during dry periods. See Variable Source
Area.
River System—A river and its entire drainage basin, including its river network, associated riparian
areas, floodplains, alluvial aquifers, regional aquifers, connected water bodies, geographically isolated
water, and terrestrial ecosystems.
Runoff—The part of precipitation, snowmelt, or other flow contributions (e.g., irrigation water) that
appears in surface streams at the outlet of a drainage basin; it can originate from both above land
surface (e.g., overland flow) and below land surface sources (e.g., ground water). Units of runoff are
depth of water (similar to precipitation units, e.g., mm). This measurement is the depth of water if it
were spread across the entire drainage basin. Can also be expressed as a volume of water (i.e., m3, feet3,
acre-ft).
Saturated Zone—The zone below the land surface where the voids in soil and geologic material are
completely filled with water. Water in the saturated zone is referred to as ground water. The upper
surface of the saturated zone is referred to as the water table. See Ground Water, Unsaturated Zone,
Water Table.
Saturation Overland Flow—Water that falls onto a saturated land surface and moves overland to the
nearest stream or river.
Seasonality—Refers to the seasonal distribution of water surplus of a river system. See Water Surplus.
Secondary Production—The generation of biomass of consumer organisms that feed on organic
material from primary producers (algae, microbes, aquatic and terrestrial plants), and biomass of
predators that feed on consumer organisms. See Primary Production.
Seed Bank—Viable dormant seeds that accumulate in soil or in sediments under water. See Egg bank.
Seep—A small area where water slowly flows from the subsurface to the surface. A seep can also refer to
a wetland formed by a seep; such a wetland is referred to as a ground-water slope wetland.
Seepage—Water that flows from a seep.
Shallow Ground Water—Ground water with shallow hydrologic flowpaths. See Local Ground Water.
Sink Function—Any function within a stream or wetland that causes a net decrease in material flux
(imports exceed exports).
Snowpack—Accumulation of snow during the winter season; an important source of water for streams
and rivers in the western United States.
Snowmelt—The complete or partial melting and release of liquid water from seasonal snowpacks.
Solute—A substance that is dissolved in water.
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Source Area—The originating location of water or other materials that move through a river system.
Source Function—Any function within a stream or wetland that causes a net increase in material flux
(exports exceed imports).
Spillage—Overflow of water from a depressional wetland to a swale or channel.
Spring—A surface-water body formed when the side of a hill, a valley bottom, or other excavation
intersects a flowing body of ground water at or below the local water table.
Stable Isotope Tracer—Certain elements such as oxygen, hydrogen, carbon, and nitrogen have multiple
isotopes that occur in nature that do not undergo radioactive decay. These isotopes can be used to track
the source and movement of water and other substances.
Stage—The elevation of the top of a water surface.
Stream—A relatively small volume of flowing water within a visible channel, including subsurface water
moving in the same direction as the surface water, and lateral flows exchanged with associated
floodplain and riparian areas. See River.
Stream Burial—The process of incorporating streams—particularly headwaters—into storm sewer
systems, usually by routing through underground pipes.
Stream Power—A measure of the erosive capacity of flowing water in stream channels or the rate of
energy dissipation against the stream bed or banks per unit of channel length that has the mathematical
form: ooa = pgQS where ooa is the stream power, p is the density of water (1000 kg/m3), g is acceleration
due to gravity (9.8 m/s2), Q is discharge (m3/s), and Sis the channel slope.
Stream Network—See River Network. A stream network is the same as river network, but typically
refers to a smaller spatial scale.
Stream Reach—See Reach.
Storm—A precipitation event that produces an increase in streamflow.
Stormflow—The part of flow through a channel that occurs in direct response to precipitation; it
includes surface and subsurface sources of flow. See Direct Runoff.
Stream Order (Strahler)—A method for stream classification based on relative position within a river
network, when streams lacking upstream tributaries (i.e., headwater streams) are first-order streams
and the junction of two streams of the same order results in an increase in stream order (i.e., two first-
order streams join to form a second-order stream, two second-order streams join to form a third-order
stream, and so on). When streams of different order join, the order of the larger stream is retained.
Stream-order classifications can differ, depending on the map scale used to determine order.
Streamflow—Flow of water through a stream or river channel. See Discharge.
Subsurface Water—All water that occurs below the land surface.
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Surface Runoff—See Overland Flow.
Surface Water—Water that occurs on Earth's surface (e.g., springs, streams, rivers, lakes, wetlands,
estuaries, oceans).
Surface-water Bodies—Types of water bodies that comprise surface water. See Surface Water.
Swale—A nonchannelized, shallow trough-like depression that carries water mainly during rainstorms
or snowmelt. A swale might or might not be considered a wetland depending on whether it meets the
Cowardin et al. (1979) three-attribute wetland criteria. See Wetland.
Symmetry Ratio—The size ratio of a minor tributary (T2) to a major tributary (71) at a confluence.
Discharge (Q2/Q1), drainage area [A2/A1], or channel width [W2/W1] can be used to characterize the
ratio of tributary size.
Terminal Source Stream—A first-order stream that intersects another first-order stream.
Terrace—An historic or prehistoric floodplain that has been abandoned by its river and is not currently
in the active floodplain. See Floodplain.
Terrene Wetlands—"Wetlands surrounded or nearly so by uplands and lacking a channelized outlet
stream; a stream may enter or exit this type of wetland but it does not flow through it as a channel;
includes a variety of wetlands and natural and human-made ponds" (Tiner, 2011).
Tracer—A substance that can be used to track the source and movement of water and other substances.
Transformation Function—Any function within a stream or wetland that converts a material into a
different form; the amount of the base material is unchanged (base exports equal base imports), but the
mass of the different forms can vary.
Transmission Loss—The loss of runoff water by infiltration into stream and river channel beds as water
moves downstream; this process is common in arid and semiarid environments.
Transport Mechanism—Any physical mechanism, such as moving water, wind, or movement of
organisms, which can transport materials or energy. As used in this report, the term specifically refers to
physical mechanisms that move material or energy between streams or wetlands and downstream
waters.
Tributary—A stream or river that flows into a higher order stream or river.
Turnover Length—The ratio of the downstream flux of organic carbon to ecosystem respiration per
length of stream. It approximates the average distance that organic carbon is expected to travel before it
is consumed and mineralized by aquatic organisms.
Unconfined Aquifer—An aquifer that has a water table; the aquifer is not bounded by lower
permeability layers. See Confined Aquifer.
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Unsaturated Zone— Also referred to as the vadose zone. The zone between land surface and the water
table within which the moisture content is less than saturation and pressure is less than atmospheric.
Soil pore spaces also typically contain air or other gases. See Saturated Zone.
Uplands—(1) Higher elevation lands surrounding streams and their floodplains. (2) Within the wetland
literature, specifically refers to any area that is not a water body and does not meet the Cowardin et al.
(1979)-attribute wetland definition. See Wetland.
Uptake Length (for dissolved nitrogen in streams)—The distance traveled in the water column
before algal and microbial assimilation occurs.
Valley—A depression of the earth's surface that drains water between two upland areas.
Variable Source Area—Neither stormflow nor baseflow is uniformly produced from the entire surface
or subsurface area of a basin. Instead, the flow of water in a stream at any given moment is influenced by
dynamic, expanding or shrinking source areas, normally representing only a few percent of the total
basin areas. The source area is highly variable during stormflow. During large rainfall or snowmelt
events, the flowing portions of the river network, and associated source areas, expand. As the event
ends, the network and source areas contract.
Vernal Pool—Shallow seasonal wetlands that generally accumulate water during colder, wetter months
and gradually dry down during warmer, dryer months.
Water Balance—The accounting of the volume of water that enters, leaves, and is stored in a hydrologic
unit, area, or arbitrarily defined control volume, typically a drainage basin or aquifer, during a specified
period of time.
Water Body—Any sizable accumulation of water on the land surface, including streams, rivers, lakes,
and wetlands.
Water Surplus—Water that is available for streamflow or recharge of ground water; precipitation
minus evapotranspiration.
Water Table—The top of the zone of saturation of an unconfined aquifer.
Watershed—The area drained by a stream, river, or other water body; typically defined by the
topographic divides between one water body and another. Synonymous with Catchment and Drainage
Basin.
Wet Channel—Channel with flowing or standing water.
Wetland—An area that generally exhibits at least one of the following three attributes (Cowardin et al.,
1979): (1) is inundated or saturated at a frequency sufficient to support, at least periodically, plants
adapted to a wet environment; (2) contains undrained hydric soil; or (3) contains nonsoil saturated by
shallow water for part of the growing season.
Wetland Storage—The capacity of a wetland to detain or retain water from various sources.
Xeric—An area, environment, or habitat that is generally dry with very little moisture. See Hydric, Mesic.
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References
Cowardin, L. M., V. Carter, F. C. Golet, and E. T. LaRoe. 1979. Classification of wetlands and deepwater
habitats of the United States. U.S. Department of the Interior, U.S. Fish and Wildlife Service, Office of
Biological Services, Washington, DC.
Tiner, R. W. 2011. Dichotomous keys and mapping codes for wetland landscape position, landform,
water flow path, and waterbody type descriptors: Version 2.0. U.S. Fish and Wildlife Service,
National Wetlands Inventory Program, Northeast Region, Hadley, MA.
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APPENDIX B. CASE STUDIES
B.I Case Study: Carolina and Delmarva Bays
B.1.1 Abstract
Carolina and Delmarva bays are ponded depressional wetlands that occur along the Atlantic Coastal
Plain from northern Florida to New Jersey. Most bays receive water through precipitation, lose water
through evapotranspiration, and lack natural surface outlets. Both mineral-based and peat-based bays
have shown connections to shallow ground water. Bays typically are near each other or near permanent
waters, providing the potential for surface-water connections in large rain events via overland flow. Fish
are reported in bays that are known to dry out, indirectly demonstrating surficial connections.
Amphibians and reptiles use bays extensively for breeding and for rearing young. These animals can
disperse many meters on the landscape and can colonize, or serve as a food source to, downstream
waters. Similarly, bays foster abundant insects that can become part of the downstream food web.
Humans have ditched and channelized a high percentage of bays, creating new surface connections to
other waters and allowing transfer of nutrients, sediment, and methylmercury.
B.1.2 Introduction
B.1.2.1 Definition and Geographic Extent
Carolina bays are elliptical, ponded, depressional wetlands that occur along the Atlantic Coastal Plain
from northern Florida to New Jersey (Prouty, 1952; Williams, 1996; Hunsinger and Lannoo, 2005). They
have been called "geographically isolated" wetlands (i.e., wetlands surrounded by uplands; Tiner, 2003),
and range from permanently inundated to frequently dry (Sharitz, 2003). Carolina bays range in size
from greater than 3,600 ha to less than 1 ha and are most abundant in North Carolina and South
Carolina (Sharitz and Gibbons, 1982; Sharitz, 2003). Carolina bays that are geographically specific to the
Delmarva Peninsula are often referred to as Delmarva bays. Delmarva bays frequently have the same
elliptical shape and orientation as other Carolina bays (Stolt and Rabenhorst, 1987a), yet some lack the
shape or rim (Sharitz, 2003).
The number of Carolina bays was estimated at 500,000 in the 1950s (Prouty, 1952), but only
10,000-20,000 remained by the early 1990s (Richardson and Gibbons, 1993). Carolina and Delmarva
bays have been ditched and drained for agricultural purposes (Figure B-l; Sharitz, 2003). A study of
2,651 Carolina bays in South Carolina found that 97% of bays larger than 0.8 ha had been disturbed by
agriculture or logging (Bennett and Nelson, 1991). The northern Delmarva Peninsula has an estimated
1,500-2,500 Delmarva bays remaining (Stolt and Rabenhorst, 1987a). The number of Carolina and
Delmarva bays is likely an underestimation, because many are too small to be readily mapped. The
National Wetlands Inventory maps have mapping units of 0.4-1.2 ha, but the Department of Energy's
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Figure B-l. Aerial photograph of Carolina bays within a region of the upper Coastal Plain of South
Carolina. (A) Infrared image showing the pattern of intact and disturbed Carolina bays within a region of
the upper Coastal Plain of South Carolina (scale: 1 cm = 1.5 km), and (B) the same image with bays (or
former bays that have been disturbed by agriculture) outlined. Reprinted with permission from Sharitz
(2003).
Savannah River Site on the upper Coastal Plain of South Carolina has 371 known Carolina bays with 46%
having an area of 1.2 ha or less (Sharitz, 2003).
B.l.2.2 Geology
The origin of Carolina and Delmarva bays is unknown, but has been attributed to meteorite impacts,
substrate dissolution, and historic modification of shallow ponds through the action of waves generated
by winds (Johnson, 1942; Savage, 1982; Ross, 1987; Stoltand Rabenhorst, 1987a; Grant etal., 1998).
The soils of Carolina and Delmarva bays range from mineral to organic depending on the position in the
landscape, hydrologic conditions, vegetation, and disturbance (Stolt and Rabenhorst, 1987b; Sharitz,
2003). Most bays have alternating layers of sand or silt with impervious clay (Bliley and Pettry, 1979).
The organic horizons in bays can range from 1 to 200 cm, with bays near the coast more likely to have
the thicker peat deposits (Newman and Schalles, 1990). Despite variation in soil content, water often
quickly infiltrates these soils before reaching an impervious clay layer (Sharitz, 2003).
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B.1.2.3 Hydrology
Carolina and Delmarva bays gain water primarily from precipitation and lose water by
evapotranspiration (Sharitz, 2003). Thus, these systems respond to seasonal rainfall, snowmelt, and
temperature. The water levels of Carolina and Delmarva bays therefore fluctuate. The water level in a
bay can change from 1-2 m above the soil surface to more than 1 m below the surface (Knight et al.,
1989; Schalles and Shure, 1989; Lide et al., 1995; Sharitz, 2003). Bays often are wetter in winter and
early spring, when evapotranspiration rates are low, and tend to dry down in summer when
evapotranspiration rates are high. Recent work by Lang et al. (2012) using highly accurate LiDAR-
derived stream maps has shown that the proportion of wetlands intersected by stream channels (and
thus not geographically isolated) is higher than previously thought.
In an analysis of the Tuckahoe Creek watershed in the Delmarva Peninsula, the High Resolution NHD
and NHD Plus were found to underestimate the number of wetlands intersected by natural stream
channels by 13% and 27%, respectively (Lang et al., 2012). Other hydrologic inputs to bays include
artesian wells (Wells and Boyce, 1953), shallow ground water (Phillips and Shedlock, 1993; Lide et al.,
1995; Caldwell etal., 2007b), inlet channels (Sharitz, 2003), and some surface runoff during periods of
high rainfall. Some bays, particularly those along the coast, can be flooded by high tides and thus are
connected to coastal waters (Bliley and Pettry, 1979; Sharitz, 2003).
Despite the prevalence of clay substrates below many of these bays, some studies have found that bays
exchange shallow ground water with the surroundings (Phillips et al., 1993; Lide et al., 1995; Sun et al.,
2006; Caldwell etal., 2007a; Pyzoha etal., 2008). Some Carolina bays have natural outlet channels
(Sharitz, 2003), and many have human-created outlet channels (i.e., ditches) typically resulting in
connections to other bays or small streams (Sharitz, 2003).
B.1.2.4 Water Chemistry
Water chemistry of Carolina and Delmarva bays is affected by their position on the landscape,
weathering of underlying mineral substrate, accrual and decomposition of organic matter, and the
degree to which surface runoff, precipitation, and ground water influence their hydrology (Sharitz,
2003). In general, precipitation-fed wetlands are typically acidic and low in nutrients (Whigham and
Jordan, 2003).
Newman and Schalles (1990) reported variable water chemistry in a study of 49 Carolina bays in North
Carolina and South Carolina that spanned two transects from inland to the coast. All 49 bays were acidic
(median pH = 4.6) and were classified as soft waters (median calcium = 1.69 mg Ca2+ L-1). DOC
represented 38% of the water anions (median DOC = 17.2 mg L-1). Bays with thick peat layers tended to
be low in nutrients, whereas bays with thin peat layers had water quality characteristics similar to local
ground water (Newman and Schalles, 1990). Phillips and Shedlock (1993) also associated bay water
chemistry with shallow ground water; their study found similarities in water chemistry between upland
ground water and the margins of three Delmarva bays. The few studies of nutrient cycling within bays
indicate some have the proper wetting and drying cycles to promote denitrification.
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Several studies have shown that Carolina bays have the proper hydrology, organic matter content, and
pH for the methylation of mercury (Snodgrass et al., 2000b; Brant et al., 2002). Mercury pollution enters
water bodies from atmospheric deposition, typically in the ionic form of Hg2+. Bacteria can convert Hg2+
to methylmercury, the bioavailable form of mercury that can accumulate in fish, birds, and other
organisms. Periodic drying and flooding of Carolina bays, especially shallow ones, promotes mercury
methylation and release (Snodgrass et al., 2000b). Mercury levels did not reach acute doses but posed a
chronic risk to fish (Snodgrass etal., 2000b) and birds that feed on these fish (Brant etal., 2002).
B.I.2.5 Biological Communities
The wetting and drying cycles of Carolina and Delmarva bays promote a diverse biota, including the
presence of numerous rare and endemic species (Sutter and Krai, 1994; Edwards and Weakley, 2001;
Sharitz, 2003). Eleven types of vegetation communities have been described in regional surveys of
Carolina bays, including species-rich herbaceous communities and cypress ponds (Bennett and Nelson,
1991; Weakley and Schafale, 1991). A seed bank study at the Savannah River Site in South Carolina
reported higher diversity than any other reported freshwater wetland habitat (Kirkman and Sharitz,
1994). Researchers estimate that more than one-third of rare plant species in the Southeast occur in
nonalluvial wetlands, including Carolina bays (Sutter and Krai, 1994; Sharitz, 2003).
Carolina and Delmarva bays are highly valuable for providing habitat and food web support for
invertebrates and vertebrates (Sharitz, 2003). For example, a Savannah River Site study of zooplankton
found 44 species of cladocerans and 7 species of copepods (Mahoney et al., 1990). Another invertebrate
study showed that a 1.5-ha Carolina bay contained 115 taxa of aquatic and semiaquatic insects from 29
families and 7 orders; more than 11,600 and 8,400 insects emerged from the bay in 1992 and 1993,
respectively (Leeper and Taylor, 1998).
Approximately 10-21% of sampled Carolina and Delmarva bays had fish populations (Gibbons and
Semlitsch, 1991; Snodgrass etal., 2000a; Sharitz, 2003). The absence of predatory fish in many bays
enables abundant amphibian populations to thrive, especially those that have aquatic larval stages
(Sharitz and Gibbons, 1982; Sharitz, 2003). For example, one study sampled two 1-ha bays over the
course of a year and captured more than 72,000 amphibians, including 9 salamander and 16 frog species
(Sharitz and Gibbons, 1982). The Savannah River Site supports 34 species of amphibians, 16 of which
depend entirely on seasonal wetlands for breeding (Gibbons and Semlitsch, 1991). Several of these
amphibians are endangered or threatened, including the flatwoods salamander (Ambystoma
cingulatuni) and the gopher frog (Rana capita) (Sharitz, 2003).
Sharitz and Gibbons (1982) reported 6 turtle species, 9 lizard species, 19 snake species, and 13 small
mammal species in bays. American alligators [Alligator mississippiensis) are indigenous to southern
Carolina bays (Sharitz and Gibbons, 1982). Endangered wood storks [Mycteria americand) nest in
Carolina bays, and birds such as egrets, coots, wood ducks, and other migratory waterfowl also use
Carolina and Delmarva bays (Sharitz and Gibbons, 1982).
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B.1.3 Evidence of Connectivity
B. 1.3.1 Physical Connections
Research is ongoing on the hydrologic connectivity of Carolina and Delmarva bays to surrounding areas
via ground-water flows and intermittent surface flows. A few studies have found ground-water
connections or indirect evidence of surface-water connections.
A study by Lide et al. (1995) found a ground-water connection to a Carolina bay. The study examined a
7-ha Carolina bay on the Savannah River Site typical of other bays in western South Carolina with
loamy-sand substrate and an underlying clay layer (Lide et al., 1995). The 2-year study examined data
from 38 piezometers, borehole logs, pond-stage records, and weather data. They concluded that the
Carolina bay was not a perched wetland, but a surface expression of the water table. Although
fluctuation of pond stage was largely controlled by precipitation and evapotranspiration, nearly
continuous shallow ground-water recharge was present and shallow ground-water discharge occurred
periodically.
Phillips and Shedlock (1993) studied three Delmarva bays and also concluded that the bays were
connected to local ground water. They studied water table levels and chemistry in transects that ran
from uplands through the Delmarva bays. Local ground water strongly influenced the height of the
water table in the Delmarva bays. The ground water also was attributed to maintaining a low pH,
contributing dissolved aluminum and lowering bicarbonate in the Delmarva bay (Phillips et al., 1993).
Another Carolina bay study in western South Carolina also found evidence for ground-water
connectivity (Pyzoha et al., 2008). The more than 13-year study examined piezometer and bay water
levels monthly in an 8-ha bay with sandy-loam substrate and an underlying clay layer. Researchers
concluded that surface-water and ground-water connections were important to bay hydrology and the
bay was not an isolated system. Sun et al. (2006) incorporated climate, vegetation, and soil information
to model the hydrology of this bay, which confirmed that the bay was receiving ground-water discharge
and recharging ground water to lower topographic areas.
Caldwell et al. (2007b) also used a model to understand the hydrology of three Carolina bays in North
Carolina and inferred ground-water connections. All three bays were larger than 100 ha, and their
hydrology had not been altered by artificial drainage. Soil types were mineral on the perimeter to mostly
organic in the center. The team modeled bay hydrology using climate, vegetation, soils, and hydrology
data. They estimated that 10% of water inputs to the bays were surface runoff. Ground-water inflow was
the source of 3-26% of water volume into the perimeter of the bays, and ground-water outflow volume
(2-21%) was frequentin the center of the bays (Caldwell etal., 2007b).
In addition to ground water, several studies infer Carolina and Delmarva bays are connected to other
water bodies through surface-water connections. For example, a study of Carolina bays in Virginia
revealed that several of the largest bays were at sea level and bordered the Chesapeake Bay (Bliley and
Pettry, 1979). Tidal marshes have encroached and entered these Carolina bays, reflecting a direct link
between the Carolina bays and the estuarine environment.
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Researchers have used geographic information system methods to determine the nearest river or
tributary to Carolina bays (Sharitz, 2003). A geographic information system analysis at the Savannah
River Site of 371 Carolina bays showed that 8% were within 50 m of a stream or tributary and 12%
were within 100 m (mapping units with a minimum resolution of 0.22 ha; Sharitz, 2003). The same
methods showed that 12% of the 2,170 Delmarva bays in Maryland were within 50 m and 19% were
within 100 m of streams (mapping units with a minimum resolution of 0.40 ha; Sharitz, 2003). During
large storms, the bays located closest to the river network can exhibit hydrologic connections via
overland flow or shallow ground-water flow.
Perhaps the strongest evidence that Carolina bays are connected hydrologically to streams or estuaries
is that many of these bays are ditched, creating a conveyance for surface water. These ditches commonly
connect the surface water of bays to other bays that are lower on the landscape, and ultimately, to
streams (Sharitz, 2003).
B.I.3.2 Chemical Connections
Few peer-reviewed papers examine chemical connections between Carolina and Delmarva bays and
other waters. One, by Phillips et al. (1993), examined ground water in the Delmarva Peninsula and found
that the amount of nitrate in ground water decreased with the presence of forested depressional bays.
The authors speculated that the nitrate reduction was due to denitrification in the wetlands. These
systems do have the appropriate wetting and drying hydrology to promote denitrification, which could
reduce the amount of nitrate in both ground water and surface waters (Groffman et al., 1992).
Carolina and Delmarva bays are frequently connected chemically to downstream waters through
ditches. If the bays are sediment and nutrient sinks due to their surficial isolation, ditch connections
would make them sources for these materials. For example, Bennett and Nelson (1991) reported that
71% of 2,600 bays were disturbed by agriculture. Whereas the bays might have been a nutrient sink for
excess fertilizer that was in surface runoff, these nutrients now could pass through the bays and into the
ditches, reaching downstream locations. Additionally, the conditions in Carolina bays have been shown
to promote mercury methylation (Snodgrass etal., 2000b). If these bays connect to downstream waters
via ditches, some bioavailable mercury would be expected to move to other waters.
B.I.3.3 Biological Connections
Carolina and Delmarva bays are "hotspots" for regional biological diversity and animal use (Sharitz,
2003), which indicates a high potential for movement between bays and other water bodies. The current
published evidence for biological connections between bays and other waters is, however, limited or
indirect.
The presence offish in Carolina and Delmarva bays indirectly demonstrates that these bays are
connected to other waters. For example, fish were found in 21% of 63 Carolina bays on the Savannah
River Site, many of which dry out during parts of the year; fish likely colonized these bays through
intermittent or permanent surface hydrologic connections (Snodgrass et al., 1996). One Carolina bay in
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North Carolina, Mattamuskeet Bay, has been colonized by both freshwater and estuarine fishes through
four canals connecting the bay to Pamlico Sound (Rulifson and Wall, 2006).
Insect emergence from bays can affect nearby waters. Leeper and Taylor (1998) studied insects in a
1.5-ha Carolina bay and recorded 115 taxa representing 29 families. Of the 39 genera of the family
Chironomidae represented, 16 are known to live in both pond and stream environments (Hudson et al.,
1990; Leeper and Taylor, 1998). Although Leeper and Taylor (1998) did not directly document
movement, these species can hatch in Carolina bays and then become important food sources for fish in
nearby streams after adult emergence and aerial dispersal. The total number of chironomids emerging
from the aforementioned Carolina bay was moderate compared to other wetlands, but cumulative
emergence from thousands of bays across the landscape would create a significant food source for
organisms, including fishes, in other nearby waters.
Carolina and Delmarva bays are immensely productive amphibian breeding habitats, and are critical for
persistence of pond-breeding amphibian populations that can move to other water bodies (Sharitz and
Gibbons, 1982). Gibbons etal. (2006) documented more than 360,000 juvenile amphibians from 24
species, emigrating from one Carolina bay during a single breeding season. More than 95% of the
biomass (about 1,330 kg) came from juveniles of the southern leopard frog [Rana sphenocephald), which
is known to use both stream and wetland habitats (Table 4-2). Given the finding that 12-19% of
Carolina and Delmarva bays were within 100 m of a tributary (Sharitz, 2003), amphibians emigrating
from these bays could transfer extremely high levels of energy and organic matter into rivers and
streams. About 90% of Carolina bays located in the Savannah River Site have a tributary or river within
1,600 m (Sharitz, 2003).
B.1.4 Carolina and Delmarva Bays: Synthesis and Implications
The key findings of this case study are as follows:
• Both peat-based and mineral-based bays have been shown to have shallow ground-water inputs
and outputs.
• Some Delmarva bays have surface-water connections to the Chesapeake Bay, and the many bays
near each other and near permanent waters can be connected during high-precipitation events.
• Human channeling and ditching of the bays are widespread and create surface connections to
other waters.
• Fish are found in bays that periodically dry out, indirectly showing that a hydrologic connection
occurred at some time.
• Dispersive amphibians and reptiles use bays for breeding or rearing young.
• The abundant insects in bays could become part of the food web for downstream fish.
Although generally supporting the existence of or potential for connectivity between Carolina and
Delmarva bays and regional rivers or estuaries, the preponderance of evidence found in the literature
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we reviewed for this case study is indirect. Furthermore, evidence from this literature review that these
connections influence the physical, chemical, and biological conditions and functions of rivers or
estuaries is circumstantial. Therefore, the literature that we reviewed does not provide sufficient
information to evaluate fully the influence of Carolina and Delmarva bays on rivers and estuaries at this
time.
B.2 Case Study: Oxbow Lakes
B.2.1 Abstract
Oxbow lakes are water bodies that originate from the meanders of rivers that become cut off. They are
common in the floodplains of large rivers around the world. In the following case study, we provide
evidence from the peer-reviewed literature to support two conclusions: (1) oxbow lakes periodically
connect to the active river channel, and (2) the connection between oxbow lakes and the active river
channel provides for several ecological effects on the river ecosystem.
B.2.2 Introduction
B.2.2.1 Origin and Description
Oxbow lakes and ponds (hereafter referred to as oxbow lakes) originate from river meanders that are
cutoff from the active river channel. In floodplain rivers, natural erosion of the outer banks of curves in
the active river channel leads to increased meandering over time. As these meanders grow, the active
channel can come into contact with itself and cutoff the curved segment of the river; this cutoff channel
becomes an oxbow lake within the floodplain.
Oxbow lakes are dynamic ecosystems. Young oxbow lakes are located near the active river channel and
tend to have steep banks. As oxbow lakes are subjected to flooding over time and begin to fill with
sediment, they can become shallower and eventually develop terrestrial characteristics. Continued
movement and meandering of unconstrained, shallow river channels can leave some oxbow lakes at
considerable distances from the active river channel (Winemiller et al., 2000). Owing to the dynamic
physical processes that create and promote succession in oxbow lakes, among-lake variation in the
character and connectivity of individual oxbow lakes within a floodplain often is large.
Oxbow lakes are an integral element in alluvial floodplain valleys of meandering rivers around the world
(Winemiller et al., 2000; Glinska-Lewczuk, 2009). Studies of these ecosystems have been conducted in
river floodplains in Australia (Crook and Gillanders, 2006), Europe (Hein et al., 2003), North America
(Winemiller etal., 2000; Zeugetal., 2005), and South America (da Silva etal., 2010). Due to the common
origin, characteristics of, and interactions between oxbow lakes and rivers, evidence from around the
world is presented here.
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B.2.3 Evidence
Oxbow lakes commonly connect with the active river channel. The most evident connections are direct
physical linkages, in which water movement between the active river channel and oxbow lakes is
traceable. Although these physical connections are intrinsically important, they also facilitate the
movement and exchange of chemical and biological material between the river and lake ecosystems.
B.2.3.1 Physical Connections
Physical connections between the active river channel and oxbow lakes can be through water movement
as overland surface flow, subsurface flow from river infiltration, and subsurface flow from hillslope
aquifers (Amoros and Bornette, 2002). In some cases, natural or constructed stream channels are
present between the river and the oxbow lake. For the purpose of this report, oxbow lakes with this type
of permanent physical connection are a priori considered an integrated part of the river network.
Evidence presented here is largely for oxbow lakes that lack permanent physical connections to the
river network; therefore, we focus on overland flow events (i.e., temporary connections occurring
during high river stages and floods) and shallow ground-water flow as the dominant surface
connections between ecosystems.
Regional- and local-scale climate and hydrogeologic patterns are important for understanding the
dynamics of physical connectivity between oxbow lakes and active river channels. Regional differences
influence the predictability of hydrologic connectivity between rivers and oxbow lakes. In temperate
rivers (e.g., Brazos River, TX), surface flow connections between the river channel and oxbow lakes are
likely to occur at irregular intervals, in response to flow magnitude and lake geomorphology
(Humphries et al., 1999; Zeug and Winemiller, 2008). Tropical rivers, in contrast, are likely to have more
regular inundation patterns associated with seasonal flooding (Junk et al., 1989; da Silva et al., 2010).
The predictability of subsurface connections also can vary regionally. An isotope tracer analysis of lakes
in the Old Crow Flats, Yukon Territory, Canada, indicated that oxbow lakes receive much of their water
input from shallow ground-water flow during the relatively short thaw season (Turner etal., 2010). The
regularity of connectivity has important implications for the exchange of chemical and biological
material between oxbow lakes and the river (Junk et al., 1989; Humphries et al., 1999).
Local landscape characteristics and position of water bodies in the floodplain influence the relative
contribution of surface-water and subsurface-water movement between individual lakes and the active
river channel, as a study of oxbow lakes on the Loire and Allier Rivers, France, demonstrates. Water in
two oxbow lakes had different geochemical signatures, suggesting a difference between when river
water was introduced to the lakes (Negrel et al., 2003). The younger oxbow lake was more connected to
the surface network due to its closer proximity to the river channel and a small stream connection, while
an older oxbow lake, which was more distant from the river channel, was more dependent on
subsurface flow (Negrel etal., 2003).
In addition to these spatial differences, temporal differences can occur in the short-term dynamics of
hydrologic connectivity. Amoros and Bornette (2002) describe a system of pulsing connectivity, where
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the direction of water exchange between floodplain water bodies, including oxbow lakes, and a river is
related to river stage. At low water stage, floodplain water bodies might receive water from a hillslope
aquifer, and water from the oxbow lake likely drains through the alluvium toward the river. In contrast,
when a river has a high water stage, water is more likely to seep through the alluvium from the river to
the oxbow lake. Finally, inundation would result in surface-water connectivity, where river water moves
overland to the oxbow lake. This pattern of pulsing connectivity is influenced by the local topography
and the characteristics of the floodplain alluvium (Amoros and Bornette, 2002) and is an illustration of
the expansion and contraction concepts described in the framework (Section 2.2.3; Figure 1-2).
Physical connectivity between oxbow lakes and the river network has direct consequences on the
hydrologic dynamics of that river network. Oxbow lakes provide flood protection. Like other floodplain
water bodies, they retain water. This retention lowers water velocity and can reduce the height of
floodwater over nearby terrestrial landscapes (Winemiller et al., 2000). In addition to storing
floodwaters, oxbows trap sediment as the velocity of floodwaters declines during the process of
retention, allowing sediment to settle out of suspension.
Human alterations of natural flow patterns in rivers can influence connectivity between oxbow lakes
and the active river channel. On one hand, connectivity can be enhanced. Channels between oxbow lakes
and the river channel often are constructed for their benefits to biological productivity (Glinska-
Lewczuk, 2009). On the other hand, isolation might be enhanced. An analysis of sediment cores in two
small oxbow lakes in the Vistula River valley, Poland, showed changes in sedimentation rate and grain
size following flood dike construction along the river (Galbarczyk-Gasiorowska etal., 2009). These
changes in sedimentation can alter the balance of subsurface connections. The absence of channel
migration since the 1980s has restricted flooding to areas close to the main channel of the Ebro River,
Spain. The effects of this diminished river-floodplain interaction (e.g., erosive floods) left two of three
oxbow lakes examined relatively isolated from the river channel, with a thick layer of fine sediment and
thus little connection to subsurface flows (Cabezas et al., 2009).
B.2.3.2 Chemical Connections
The dynamics of hydrologic connectivity are important for understanding the chemical character of
oxbow lakes. Flooding of the river facilitates exchange of chemicals between the river water and the
water in oxbow lakes. In some cases, these surface-water exchanges reset the chemical environment in
oxbow lakes (e.g., periodic floods introducing well-aerated water to oxbow lakes in Poland; Obolewski et
al., 2009). The chemical effects of flooding are not limited to changes in the water column. For example,
the isolation of oxbow lakes from the active river channel corresponded with changes in sediment
chemistry, and ultimately, an acceleration of eutrophication (Galbarczyk-Gasiorowska etal., 2009).
Subsurface connections also influence oxbow lake chemistry in important ways. For example, an
assessment of oxbow lakes on the River Lyna, Poland indicated that nutrient concentrations in oxbow
lakes likely were influenced by a combination of river water from surface connections, ground-water
seepage from the alluvial aquifer, infiltration from hillslope runoff, and inlake nutrient processing
(Glinska-Lewczuk, 2009). In some cases, these other connection types can play a more important role in
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oxbow lake chemistry than periodic surface connections created during flood events. An examination of
sediment chemistry in floodplain water bodies on the River Havel, Germany showed little effect of
flooding on sediment chemistry (particulate organic matter, carbon, nitrogen, phosphorus, and iron) in
oxbow lakes (Knosche, 2006). As is the case with physical connectivity, the relative importance of
surface and subsurface connectivity depends on local characteristics of the floodplain ecosystem.
Alterations of natural flood dynamics affect the exchange of chemical materials between the river and
oxbow lakes. Total organic carbon accretion and total nitrogen accretion in river floodplains are
important ecosystem functions of floodplain water bodies, like oxbow lakes, that might improve water
quality in rivers (Mitsch, 1992). An analysis of sediment, carbon, and nitrogen accretion in oxbow lakes
on the River Ebro, Spain showed lower recent accumulation (1963-2007) compared to the past
(1927-1963; Cabezas etal., 2009). In this example, the reduced accumulation of carbon and nitrogen
concentrations in oxbow lake sediment was related to reduced size and frequency of flood events in this
floodplain ecosystem (Cabezas etal., 2009).
Importantly, oxbow lakes reduce pollution loading to the river network. Oxbow lakes can intercept
nutrients from upland runoff, leaving them in the oxbow lake rather than in the river (Glinska-Lewczuk,
2009). A similar process of physical interception is observed in riparian wetlands, where wetland
ecosystems have been considered habitats that might control nonpoint-source pollution of nutrients
(Mitsch, 1992), sediment (Brix, 1994), or pesticides (Gregoire et al., 2009) to rivers. In addition to being
areas of deposition, high mineralization rates in oxbow lakes suggest that these lakes can process and
remove some nutrients in terrestrial runoff before the runoff reaches the river channel (Winemiller et
al.,2000).
B.2.3.3 Biological Connections
Hydrologic connectivity influences the biological character of oxbow lakes and facilitates exchange of
biological material between oxbow lakes and the active river channel. Evidence also suggests a
temporally dynamic relationship between biological assemblages of river and oxbow lake ecosystems.
Oxbow lakes represent important areas of relatively high biological productivity in the floodplain
landscape. Oxbow lakes can be a source of plankton to the active river channel (Hein et al., 2003). In
contrast to terrestrial sources of carbon that often dominate the water column of rivers, plankton is
more labile and easier to assimilate into aquatic food webs (Thorp and Belong, 2002; Bunn et al., 2003).
The connectivity relationship has added complexity for plankton, because oxbow lakes need to be
periodically isolated from the river to establish populations of these organisms. Intermediate residence
times (i.e., the amount of time a water molecule spends in a lake) of between 10 and 27 days in oxbow
lakes along the River Danube resulted in the highest carbon flow between phytoplankton and
zooplankton (Keckeis et al., 2003). Likewise, the time since inundation is an important factor influencing
the composition of zooplankton communities. Recently inundated floodplain water bodies are
dominated by rapid-colonizing rotifers, and then become dominated by cladocerans as the time since
inundation increases (Baranyi et al., 2002). In this study, total zooplankton biomass, crustacean
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biomass, and the number of crustacean species were positively related to time since inundation. These
results indicate a relationship between the time since inundation and plankton assemblages, and
suggest that this relationship exists because colonization and reproduction within an oxbow lake
requires time without disturbance.
Although short periods of isolation are necessary for the development of within-oxbow productivity,
periodic connections are important for plankton exchange between oxbow lakes and the active river
channel. Exchange can occur from the river to the oxbow lake (e.g., juvenile riverine fish might feed in
floodplain water bodies; Baranyi et al., 2002) or from the oxbow lake to the river (e.g., phytoplankton;
Hein et al., 2003). These periodic connections between floodplain water bodies and the corresponding
export of labile phytoplankton from floodplain water bodies to rivers contribute to the food sources of
biological assemblages in nearby rivers (Thorp and Belong, 2002; Bunn etal., 2003; Keckeis etal., 2003).
Connectivity between oxbow lakes in the floodplain and the active river channel is important for
maintaining mollusk populations in oxbow lakes. A comparison of three oxbow lakes with different
levels of connectivity (lotic, semilotic, and isolated) showed the highest level of mollusk diversity in the
semilotic lake (eight vs. four taxa in each of the other lakes) on the Lyna River, Poland (Obolewski et al.,
2009). In this example, the occurrence of taxa was associated with physiochemical characteristics
(oxygen, temperature, and phosphorus) of oxbow lakes. These findings support the idea that the degree
of oxbow lake-river connectivity influences the abundance and composition of mollusk communities in
floodplain water bodies, and these communities support the diversity of mollusk taxa throughout the
river system (Reckendorfer etal., 2006).
Physical connectivity between oxbow lakes and the active river channel influences the composition of
benthic macroinvertebrate communities in oxbow lakes. For example, hydrologic connection explained
28% of the variability in benthic invertebrate communities among sites in the active river channel,
constructed oxbow lakes, and natural oxbow lakes of the Middle Ebro River, Spain (Gallardo et al.,
2008). Macroinvertebrate richness and abundance increased with hydrologic connectivity (i.e., floods
and flow pulses) between oxbow lakes and the river channel, and a diversity metric (Shannon index)
peaked at intermediate levels of connectivity (Gallardo etal., 2008).
Oxbow lakes have food resources and habitat that often support abundant fish populations (Winemiller
et al., 2000; Zeug et al., 2005; Zeug and Winemiller, 2008; Zeug et al., 2009). A comparison offish
biomass in oxbow lakes and a river channel showed that fish biomass in oxbow lakes was three times
the biomass caught in rivers. Average catch per unit effort in oxbow lakes was 364.3 g per 10-m seine
haul and 5,318 gnr1 ha-1 of gillnet sampling, versus 138.1 gper 10-m seine haul and 495 gnr1 ha-1 of
gillnet sampling in the river (Winemiller et al., 2000). Additional studies by this research group have
found similar patterns for juvenile fish (Zeug and Winemiller, 2008).
Periodic surface-water connections between the river and oxbow lakes facilitate the movement offish
from the river to oxbow lakes, where riverine fish can exploit these relatively productive floodplain
water bodies before moving back to the river. Dietary data provide evidence that oxbow lakes are
important spawning and nursery habitats for gizzard shad in the Brazos River, TX (Zeug et al., 2009).
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Isotope analysis showed that gizzard shad in oxbow lakes had different isotopic signatures based on
habitat type: oxbow, river, and an oxbow-river mix (Zeug et al., 2009). Although oxbow lakes clearly
provided habitat for both juvenile and adult shad, the authors did not observe oxbow-specific isotopic
signatures in shad in the river channel (Zeug et al., 2009). In addition, an analysis of otolith chemical
signatures by Crook and Gillanders (2006) indicates that floodplain lakes were an important source of
carp recruitment to the Murray-Darling River, where floodplain lakes were estimated to be the source of
98% of the young-of-year carp for areas 140 km downstream of the floodplain lakes. In a third example,
floodplain water bodies, with their diverse and productive habitats, were considered nurseries for
drifting larvae of migratory fish (Meschiatti et al., 2000). Half the migratory fish species from the Mogi-
Guacu River, Brazil also were observed as juveniles in oxbow lakes along the river (24 of the 46
migratory riverine species were observed in 2 oxbow lakes), and most of the migratory fish observed in
oxbow lakes were juveniles, rather than larvae or reproductively mature age classes (Meschiatti et al.,
2000). This age structure suggests that the oxbow lakes were not the site of reproduction, but were
important habitats for juvenile fish.
Individual fish species have specific habitat and reproductive requirements and use floodplain habitats
in different ways, giving the dynamic hydrologic connectivity of oxbow lakes and the river network
added significance. For example, owing to variable flow in the Rio Grande, NM, recruitment success
varies between years of high (Junk et al., 1989) and low flow (Humphries et al., 1999), which contributes
to overall fish diversity in the Rio Grande (Pease et al., 2006). Likewise, in a 5-year study offish in
floodplain lakes, Shoup and Wahl (2009) discuss how individual oxbow lakes had different conditions
and thus varied in suitability for different fish species. In their study, interannual variability was present
in oxbow lake hydrology (lake-river connectivity ranged from 0 to more than 21 weeks per year) and
water chemistry, and in associated differences in fish assemblages (Shoup and Wahl, 2009). Because of
the complex relationships observed in their study, Shoup and Wahl (2009) concluded that the entire
floodplain should be considered a single functioning unit that supports the overall biological integrity of
a river.
B.2.4 Oxbow Lakes: Synthesis and Implications
The key findings of this case study are as follows:
• Evidence indicates the presence of physical, chemical, and biological connections between
oxbow lakes and the river channel. The specific local and regional characteristics of both the
oxbow lakes and the river influence these connections.
• Some of the best-documented observed functions of oxbow lakes are as sources or sinks for
water, sinks for nutrients from upland runoff that might otherwise flow into rivers, and sources
of food and refuges for riverine organisms.
• Human alteration of these connections can be detrimental to the dynamics that balance
connectivity and exchange between oxbow lakes and the active river channel. Practices that
alter the natural flow regime of the river (e.g., river regulation) or inhibit periodic flooding of
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oxbow lakes (e.g., levees) affect movement of water and sediment, the use of oxbow lakes by
riverine fish, and the regional biological diversity of floodplain water bodies.
• Interannual variability in oxbow lake hydrology, water chemistry, and fish assemblages
demonstrate complex relationships between rivers and floodplain open waters and river
systems, in which the water bodies in floodplains function as single unit supporting the overall
biological integrity of the river.
Although the incidence of observed connectivity between oxbow lakes and river networks varies
according to spatial, temporal, physical, and biological factors, most of the evidence examined indicates
that oxbow lakes are important determinants of the physical, chemical, and biological condition and
function of rivers.
B.3 Case Study: Prairie Potholes
B.3.1 Abstract
Prairie potholes are a complex of glacially formed wetlands, usually occurring in depressions that lack
permanent natural outlets, that are found in the central United States and Canada. The vast area they
occupy is variable in many aspects, including climatically, topographically, geologically, and in terms of
land use and alteration, which imparts variation on the potholes themselves. Potholes demonstrate a
wide range of hydrologic permanence, from holding permanent standing water to wetting only in years
with high precipitation, which in turn influences the diversity and structure of their biological
communities. Owing in large part to their spatial and temporal variability, individual prairie potholes
span the entire continuum of connectivity to and isolation from the river network and other bodies of
water. Potholes generally accumulate and retain water effectively due to the low permeability of their
underlying soil, which can modulate flow characteristics of nearby streams and rivers. Potholes also can
accumulate chemicals in overland flow, thereby reducing chemical loading to other bodies of water.
When potholes are artificially connected to streams and lakes through drainage, isolation is eliminated
and they become sources of water and chemicals. Potholes also support a community of highly mobile
organisms, from plants to invertebrates to birds, that travel among potholes and that can biologically
connect the entire complex to the river network.
B.3.2 Introduction
Prairie potholes are a complex of wetlands and water bodies that cover more than 700,000 km2 of the
north-central United States and southern Canada, in an area referred to as the prairie pothole region
(PPR; Kantrud et al., 1989). Formed by the retreat of Pleistocene glaciers, potholes are shallow
depressions underlain by low-permeability, clay-rich glacial tills that allow for the collection and
temporary retention of water. Prairie potholes range widely from more than 200 ha to less than 0.5 ha
in surface area with an average of 1 ha or less (Cowardin et al., 1981; Kahara et al., 2009). Their density
across the landscape varies from region to region, from roughly 5 potholes knr2 in the eastern part of
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the region to up to 90 knr2 in the western part as a result of several factors, including patterns of glacial
movement, topography, and climate (van der Valk and Pederson, 2003; Kahara et al., 2009).
By the 1980s, more than 50% of potholes in the region were filled, drained, or ditched, with much higher
percentages lost in agriculturally intensive regions like Iowa (Figure 2-21; Dahl, 1990). Conservation of
remaining potholes and restoration of others have been prompted by various means, including the
"Swampbuster" provision of the 1985 Food Security Act and the Wetland Reserve Program
(administered by the U.S. Department of Agriculture National Resource Conservation Service since
1990).
B.3.2.1 Hydrologic Dynamics
Prairie potholes are hydrologically dynamic and heterogeneous, varying both spatially and temporally
(Euliss et al., 2004). Water inflows consist largely of precipitation in the form of spring snowmelt runoff
or summer rain falling directly into the depressions (Carroll et al., 2005). Some potholes also receive
ground-water discharge (Winter and Rosenberry, 1998). Evapotranspiration accounts for most of the
water outflow in most potholes (Carroll etal., 2005; van der Kamp and Hayashi, 2009). In some
situations, water can leave the basin as overland flow (known as "fill-and-spill") and shallow or regional
ground-water recharge. Potholes with ground-water flow-through or with directional reversal of
ground-water flow (discharge under some conditions and recharge under others) also have been
identified (Rosenberry and Winter, 1997).
Prairie potholes experience seasonal cycles in water level. Potholes fill in the spring, typically reaching
maximum water volume as melting snow, unable to infiltrate frozen upland soils, runs overland into
topographically low places on the landscape. Water levels decline through the summer, although they
can be maintained or increase due to summer rains (Winter and Rosenberry, 1995). Hydrologic
permanence of these systems varies among prairie potholes in response to precipitation, pothole depth,
underlying soil permeability, and position in relation to the water table. Temporary potholes have
intermittent standing water only in periods of high precipitation. Seasonal potholes collect water in
spring, but typically dry by mid-summer each year. Semipermanent potholes usually maintain standing
water throughout the year and occasionally dry in years with low precipitation. Permanent potholes
have standing water year-round and maintain standing water from year to year. Importantly, loss of
temporary and seasonal potholes has occurred at higher rates than loss of permanent pothole wetlands,
because shallower, less permanent basins are easier to drain (Miller et al., 2009).
Spatial variation in precipitation affects interannual variation in water level and hydrologic permanence.
The east-west gradient across much of the PPR delivers more than 800 mm of average precipitation to
northwestern Iowa each year and less than 500 mm of average precipitation to most of North Dakota.
These dynamics also depend on 20- to 200-year, large-scale climate cycles, including periodic flood and
drought conditions (Ashworth, 1999; Leibowitz and Vining, 2003). Annual average climate and longer
climate cycles profoundly affect individual pothole dynamics and the interactions both among potholes
and between potholes and broader landscape features (Winter and Rosenberry, 1998; Johnson et al.,
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2004). Hydrologic dynamics can have major effects on the diversity and abundance of organisms (Euliss
and Mushet, 2004).
In addition, topography at multiple scales, soil characteristics, and underlying geology influence pothole
dynamics and interactions. Three major physiographic regions comprise the PPR from east to west: the
Red River Valley, Drift Prairie, and Missouri Coteau. The Red River Valley was formerly a vast lake filled
with glacial melt, and today consists of the relatively topographically flat, clay-rich till surrounding the
Red River of the North. The Drift Prairie is higher in elevation than the Red River Valley, and consists of
rolling, hummocky terrain formed by glacial deposits. The Missouri Coteau has the highest elevation of
the region and relatively steep relief due to thick glacial debris deposits (Kantrud et al., 1989). More
restricted local landform zones, various till plains in the Des Moines Lobe in Iowa and the Prairie Coteau
in eastern South Dakota for example, also influence hydrologic characteristics of potholes (Miller etal.,
2009).
B.3.2.2 Chemical Functions
The chemical composition of prairie potholes is determined largely by the degree of connectivity with
ground water and the position of the wetland with respect to local and regional ground-water systems.
Seasonal wetlands located high in the landscape tend to be less saline than the wetlands situated low in
the landscape. This simplistic view is made more complex, however, by watershed characteristics,
concentration of solutes by evapotranspiration, variability in ground-water and surface-water residence
times, changing wetland volumes, and climatic variability. For example, LaBaugh et al. (1996)
documented substantial interannual changes in dominant ionic species in response to climatic
variability. These changes persisted beyond the climatic inputs, indicating that antecedent moisture
conditions also influence wetland response to a changing climate.
Nutrient (including carbon, nitrogen, and phosphorus) cycling in prairie potholes likely depends on
fluctuating water levels, wet-dry cycles, and resulting effects of vegetation cycling. Potholes tend to be
nitrogen-limited environments, with the notable exception of potholes located on agricultural land that
tend to receive runoff high in nitrate (Crumpton and Goldsborough, 1998). Denitrification that takes
place in the anaerobic zone of these and other wetlands can make them effective nitrogen sinks (van der
Valk, 2006).
B.3.2.3 Ecological Characteristics
The high spatial and temporal abiotic heterogeneity, both within an individual pothole and between
potholes across the region, creates a variety of ecological niches and contributes to high biodiversity in
these habitats. In response to hydrologic cycles, a semipermanent pothole can have up to four distinct,
concentric zones of vegetation, ranging from floating aquatic plants to upland plants. Depending on the
timing within annual or between interannual wet-dry cycles, a given pothole can have all zones or just
one zone. A pothole also could be in the process of developing zones (regenerative phase) or losing
zones (degenerative phase). Invasive species like reed canarygrass [Phalaris arundinacea) and cattail
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(Typha angustifolia and T, xglaucd) have established in streams and wetlands across the region,
disrupting natural pothole vegetation communities.
Perhaps the best-known and most well-studied attribute of prairie potholes is their role as productive
feeding and nesting habitat for waterfowl. Of the 34 species of duck that breed in North America, 12 are
common in the region, which contributes up to 80% of the continent's waterfowl game (Batt et al.,
1989). In addition, a diverse assemblage of microorganisms, invertebrates, amphibians, reptiles, and
sometimes fish, obligately or facultatively use potholes to feed or reproduce. For example, 44 different
invertebrate taxa, including nematodes, mollusks, and arthropods, were collected in Iowa potholes
(Hentges and Stewart, 2010).
B.3.3 Evidence
B.3.3.1 Physical Connections
Because prairie potholes are small wetlands that form in depressions often lacking permanent outlets,
they have been described as hydrologically isolated from each other and from other waters. In some
instances, this generalization has proved true but in others, it is false.
One of the most noted hydrologic functions of potholes is water storage. Because most of the water
outflow in potholes is via evapotranspiration, potholes can become water sinks, preventing flow to other
waters in their river or terminal lake basins. Several studies have quantified the large water storage
capacity of prairie pothole complexes. A conservative estimate puts the amount of precipitation that can
be retained in prairie potholes on land enrolled in the federal Conservation Reserve Program and
Wetland Reserve Program atmore than 555 million m3 (Gleason et al., 2008). In various subbasins
across the PPR, including those that feed Devils Lake and the Red River of the North, both of which have
a long history of flooding, potholes have consistently been estimated to hold tens of millions of cubic
meters of water (Hubbard and Linder, 1986; Vining, 2002; Gleason et al., 2007).
Water storage by prairie potholes can affect streamflow. Simulations of the Starkweather Coulee
subbasin that drains to Devils Lake indicate that streamflow declines substantially with increased
wetland storage capacity. Increasing the volume of pothole storage across the subbasin by
approximately 60% caused simulated total annual streamflow to decrease by 50% during a series of dry
years and by 20% during wet years. The weaker effect of potholes on streamflow during wet years is
likely due to high soil moisture conditions and maintenance of high water levels within potholes across
years, which causes a greater proportion of runoff to reach streams relative to dry years (Vining, 2002).
Similar simulation studies of watersheds in the Red River basin (one in North Dakota and one in
Minnesota) produced qualitatively comparable results, suggesting that the ability of potholes to
modulate streamflow can be widespread across the PPR (Vining, 2004). This work also indicates that
reducing water storage capacity of wetlands by connecting formerly isolated potholes through ditching
or drainage to the Devils Lake and Red River basins could increase stormflow and contribute to
downstream flooding. In many agricultural areas already crisscrossed by extensive surface and
subsurface drainage systems (Figure 2-21), total streamflow and baseflow are increased by directly
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connecting potholes to stream networks (Blann etal., 2009). The ensuing impacts of changing
streamflow are numerous, including effects on stream geomorphology, habitat alteration, and ecological
effects (reviewed in Blann et al., 2009).
Studies in some regions show a lack of association between pothole water storage and aspects of
streamflow. For instance, modeling of an Iowa watershed indicated that total pothole outflow and total
maximum pothole volume do not affect streamflow characteristics (Du et al., 2005). At the Minnesota
watershed within the Red River basin discussed previously, simulated annual and daily streamflow
decreased with increased pothole water storage capacity but peak streamflow was not reduced during a
simulated flooding event, possibly due to an overwhelmed capacity of wetlands and upland soils to
retain additional water (Vining, 2004). In yet another Minnesota watershed, wetland water storage
provided no explanatory power in estimating peak streamflows for small streams (Lorenz et al., 2010).
The presence or absence of an effect of pothole water storage on streamflow depends on many factors,
including patterns of precipitation, topography, and degree of human alteration. For instance, in parts of
the PPR with low precipitation, low stream density, and little human alteration, the extreme hydrologic
isolation of potholes likely results in few effects on larger waters. Neither a comprehensive examination
of the downstream effects nor a systematic characterization of potholes for the factors that determine
those effects has been conducted.
Surface-water isolation is common for many prairie potholes under average precipitation conditions,
but intense precipitation events or high cumulative precipitation over one or more seasons can result in
temporary hydrologic connectivity via overland flow. These "fill-and-spill" events between potholes
have been witnessed and measured in the Missouri Coteau and in the Drift Prairie zones of the PPR in
North Dakota (Winter and Rosenberry, 1998; Leibowitz and Vining, 2003), and inferred using digital
aerial photography (Kahara et al., 2009). All else being equal, a wetter climate such as that experienced
in the southeastern part of the PPR should promote hydrologic connectivity (Johnson et al., 2005). Local
topography can enhance or diminish the likelihood and frequency of temporary surface-water
connections. Authors have reasoned that the relatively wet and topographically low Red River Valley
zone of the PPR should display greater surface-water connectivity of potholes than either the Drift
Prairie or Missouri Coteau zones. Furthermore, they suggest that stream density will influence the
chance that pothole spillage connects to the larger river network. Thus, potholes in the Missouri Coteau,
with its limited network of streams, should be more hydrologically isolated than potholes in the Red
River Valley or Drift Prairie (Leibowitz and Vining, 2003).
Individual potholes range from isolated to highly connected to other potholes via shallow local and
deeper regional ground-water flows. A high water table and soil pocketed with root pores or fractures
from wet-dry cycles promote water movement between wetlands via shallow ground-water aquifers. In
these cases, water moves most often from topographically high, recharge wetlands to low, discharge
wetlands (van der Kamp and Hayashi, 2009), although a single wetland can shift from recharge to
discharge in years where the water table is high (Carroll et al., 2005). Other wetlands shift multiple
times from recharge to discharge conditions during a single year, which can either facilitate or prevent
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ground-water connections to nearby wetlands (Rosenberry and Winter, 1997). Potholes can connect to
the river network via ground water if both are located within the zone of shallow local aquifer flows.
One study in North Dakota described prairie wetlands and lakes as water sources to the topographically
low James River via shallow ground-water flow (Swanson et al., 1988). Broader, regional movement of
ground water is restricted by very low permeability clay-rich tills that can keep deep ground-water
recharge to only millimeters per year on average over a drainage basin (van der Kamp and Hayashi,
1998).
Human alterations of the landscape have had an impact on the connectivity of prairie potholes. Presence
or absence of a crop on the upland near a wetland can alter the degree to which the wetland receives
overland flow from the upland and the removal of water via transpiration that otherwise would
recharge ground water (Hayashi etal., 1998). Up to 30% of cropland in the Upper Midwest is artificially
drained to increase agricultural productivity (Pavelis, 1987). Filling potholes and lowering the water
table through use of field tiling for agriculture has likely increased isolation of remaining potholes by
decreasing the density of depressions containing water. Extensive surface draining and ditching,
however, have directly and dramatically increased connectivity between pothole basins and surface
waters of the river network, converting these systems from precipitation sinks to water sources (Blann
et al., 2009). Ditches create new surface-water outlets from potholes, allowing collected water to flow
into streams and rivers; drains fitted at the bottom of potholes connected to shallow subsurface pipes
often discharge to open ditches or streams (Ginting et al., 2000).
B.3.3.2 Chemical Connections
The chemical connectivity of prairie potholes is largely mediated by their hydrologic connectivity.
Hydrologically isolated potholes tend also to be isolated chemically. Unaltered potholes with no outlet
can accumulate nutrients, sediment, and other chemical compounds as they collect runoff (Crumpton
and Goldsborough, 1998; Donald et al., 1999). Such accumulations have measurable effects on the water
quality of potholes and the resident organisms (Gleason etal., 2003). Presence of these materials in
potholes is influenced by inflow, itself a function of precipitation and surrounding land use. Potholes
surrounded by tilled fields with higher precipitation, for example, tend to accumulate nutrients,
sediment, and pesticides (Gleason et al., 2008). Additionally, potholes within agricultural areas that have
not been drained or ditched are hypothesized to be nitrogen sinks, transforming nitrate in the
agricultural runoff they receive to nitrous oxide or nitrogen gas. Denitrification can transform up to 80%
of nitrate that runs off into potholes (Crumpton and Goldsborough, 1998 and references therein).
On the other hand, potholes that periodically are connected hydrologically to other bodies of water via
overland flow can transfer chemicals, such as dissolved ions (Leibowitz and Vining, 2003). Potholes
modified by ditching or drainage also have increased hydrologic connectivity and, therefore, chemical
connectivity to other water bodies (Whigham and Jordan, 2003). Wetlands drained for agriculture can
contribute nitrogen, phosphorus, sediment, pesticides, and herbicides to the waters into which they
drain (reviewed in Blann et al., 2009). For example, two wetlands in southwestern Minnesota fitted with
surface drains that connected to subsurface tiles emptying into the Watonwan River (a tributary of the
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Minnesota River) were found to be sources of total solids and total phosphorus to the river during
periods of high runoff (Ginting et al., 2000).
Although the chemical sink and periodic chemical source functions of potholes have been documented in
the literature, the overall influence of these functions on larger waters and river networks have been
difficult to quantify. This inability is partly because altered and unaltered potholes are embedded in a
matrix of land use and land management types, and many different parts of this complex landscape
affect downstream water quality and ecological communities (Blann et al., 2009). The most fruitful
future approach might be to model drainage basin sediment, nutrient, and pesticide transport under
various climatic conditions, using pothole characteristics and functions as independent, explanatory
variables (Gleason et al., 2008).
B.3.3.3 Biological Connections
Dispersal capabilities of organisms residing in potholes and features of the landscapes they must
traverse help determine the strength of biological connectivity. Although some research has focused on
internal seed and egg bank dynamics (van der Valk and Davis, 1978; Gleason et al., 2004), increasing
evidence suggests that potholes are not biologically isolated. In fact, the observation that potholes lack
an endemic aquatic and semiaquatic flora or fauna suggests that, at least over evolutionary time,
potholes have been well connected biologically to communities in other ecosystems (van der Valk and
Pederson, 2003).
Organisms can move into and out of potholes via wind, water, or land, by either self-propelling or
hitchhiking on other mobile organisms. Many species of wetland plants and insects are dispersed on the
wind (Keiper et al., 2002; Soons, 2006), including cattail (Typha spp.) seeds, which can disperse over
huge areas (more than 80 ha; van Digglen, 2006) and have been found to colonize, quickly and passively,
previously drained, restored potholes (Galatowitsch and van der Valk, 1996). Plants and invertebrates
also can travel by becoming attached to or consumed and excreted by waterfowl (Amezaga et al., 2002).
Seeds of up to half a dozen common pothole plants can be consumed and excreted by ducks in a viable
state; because migrating waterfowl fly such long distances, the maximum dispersal distance of these
hitchhiking plants is estimated to be 1,400 km (Mueller and van der Valk, 2002). Additionally, fast and
efficient recolonization of species in restored potholes, including floating aquatics and emergent
perennials, is likely facilitated by waterfowl movement (Aronson and Galatowitsch, 2008). Waterfowl
often move between wetlands during the breeding season in search of food and cover, and some species
also use habitats within the river network as wetlands dry or freeze (Pattenden and Boag, 1989; Murkin
and Caldwell, 2000). Water also can provide a means for biologically connecting potholes. Fish and other
organisms or parts of organisms that can be suspended in water (e.g., floating insect larvae or seeds)
have been hypothesized to move between potholes during spillage events (Zimmer et al., 2001; van der
Valk and Pederson, 2003; Herwigetal., 2010). Dispersal of waterborne organisms also can occur
through manmade waterways (i.e., ditches) that connect potholes to stream networks (Hanson et al.,
2005; Hentges and Stewart, 2010; Herwigetal., 2010). Most of these studies cite only anecdotal
evidence for dispersal through ditches. Populations of aquatic plants in agricultural ditches in Europe,
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however, are genetically highly structured along these man-made waterways, suggesting that these
watercourses determine dispersal pathways (Gornall et al., 1998).
Finally, overland dispersal of amphibians and mammals can connect potholes. Eight of twelve amphibian
species were able to quickly recolonize restored potholes near source populations (Lehtinen and
Galatowitsch, 2001). Although muskrat territories in the PPR are usually restricted (less than 100 m
from the home stream or wetland), they can disperse longer distances to feed and breed in prairie
wetland habitat under certain conditions (Clark, 2000 and references therein). In North Dakota,
muskrats have been observed taking up residence in potholes for a series of years, provided suitable
water levels and vegetation existed, and then emigrating, presumably to more permanent and larger
lakes and streams (Winter and LaBaugh, 2003). Not all wetland animals disperse widely, however.
Populations of the pothole-dwelling salamander Ambystoma tigrinum (studied in small, non-pothole
wetlands, in this case) can be genetically differentiated from each other down to 1.5 km, indicating low
dispersal (Routman, 1993).
Landscape features, including distance, relief, and human alterations, can promote or restrict biological
connections between wetlands and larger bodies of water. Spatial distance is one important factor to
consider. For a given species, wetlands located closer together will exchange more organisms than
wetlands that are farther apart. Therefore, landscapes in which potholes are located in relative
proximity to each other and to the river network are likely to be connected more frequently and by
more species. For example, restored potholes in pothole-dense areas tend to be recolonized by plants
more efficiently (Mulhouse and Galatowitsch, 2003), and high pothole density promotes greater
movement of waterfowl (Krapu et al., 1997). Unfortunately, quantification of biological effects of
potholes on larger waters is severely limited. In most cases, studies involving biological isolation or
connectivity in the PPR have focused on the potholes themselves as sources and recipients of organisms.
B.3.4 Prairie Potholes: Synthesis and Implications
The key findings for this case study are as follows:
• The degree to which prairie potholes are connected or could connect to river networks depends
on many factors. These factors include distance to rivers or streams, topography, precipitation,
climate cycles (seasonal and on longer time scales), biotic community composition, and artificial
drainage. Within the PPR, distance to rivers and streams is strongly influenced by the three
major physiographic regions (Red River Valley, Drift Prairie, and Missouri Coteau), which vary
in the number of potholes and stream density (e.g., Figures 2-20A and 2-20B).
• On a watershed scale, unaltered potholes often function as hydrologic sinks, sequestering water
and reducing annual streamflow, but can become sources as they spill overland under high
precipitation or low relief, or both. When artificially drained or ditched, potholes can become
sources of water, nutrients, sediment, and pesticides. Their roles as sinks and sources affect
river geomorphology and biological communities.
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• Potholes also might have direct biological effects on river networks via connectivity of resident
populations, although these effects are less well known and studied.
Because of wide variation in the conditions that determine the incidence or magnitude of connections
between prairie potholes and river networks, pothole complexes in some watersheds are more likely to
have important effects on associated rivers and lakes than others are. Given evidence in the current
literature, however, when proper climatic or topographic conditions occur, or biotic communities are
present that promote potential or observed connections, measurable influence on the physical, chemical,
and biological condition and function of downstream waters is highly likely.
B.4 Case Study: Prairie Streams
B.4.1 Abstract
Prairie streams drain temperate grasslands in the central United States. Periods of flooding and drying
characterize their hydrology, with spring-fed, perennial pools and reaches embedded within more
intermittently flowing reaches; thus, water flow along prairie stream networks exhibits high temporal
and spatial variability. Existing evidence indicates that small prairie streams are connected to
downstream reaches, most notably via flood propagation and the extensive transport and movement of
fish species throughout these networks. Nutrient retention in small prairie streams also significantly
influences nutrient loading in downstream rivers.
B.4.2 Introduction
B.4.2.1 Geography and Climate
Prairies are temperate grasslands located in the Great Plains physiographic region of the central United
States and Canada (Figure B-2). Grasses and forbs (broad-leaf plants other than grasses) dominate the
region, particularly in upland areas. Shrubs and trees can be found in lowlands, and are commonly called
gallery forests. Native prairie ecosystems once covered approximately 1.62 million km2 in North
America but have been lost almost completely since European settlement, mainly replaced by row-crop
agriculture (Samson and Knopf, 1994). Because of drastic alterations to much of the historical eastern
plains (Iowa, Illinois, Missouri, Indiana, Minnesota), our discussion centers principally on river networks
in the high plains subregion of the Great Plains (Subregion 2 in Figure B-2), where drier climate and
thin, rocky soil have limited row-crop agriculture.
Prairies generally can be characterized by their relatively low topographic relief, although areas such as
the Flint Hills in eastern Kansas, the Arikaree Breaks in northwestern Kansas, and the Arbuckle
Mountains in south-central Oklahoma have relatively steep terrain compared to that of western Kansas
or the Oklahoma panhandle (Osterkamp and Costa, 1987; Matthews, 1988). The underlying geology
consists of extensive limestone deposits, but sandstone and shale deposits or unconsolidated sands,
silts, and clays characterize other areas (Brown and Matthews, 1995). Soils in the Great Plains are
predominantly loess, but some areas such as Nebraska's Sand Hills have high percentages of sand
(Wolock et al., 2004). Although prairie soils tend to be less permeable than more humic forest soils,
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Figure B-2. Map of the United States showing physiographic subregions and major rivers of the Great
Plains: (1) glaciated prairie; (2) high plains; (3) eastern plains; and (4) Ozark Plateau. Modified from
Covich et al. (1997).
fractures and macropores of the limestone geology in some prairie areas, such as the Flint Hills, allow
for relatively rapid percolation and recharge of local ground water (Macpherson and Sophocleous,
2004).
Most of the large rivers draining the high plains subregion (e.g., the Missouri, Yellowstone, Milk,
Cheyenne, White, Niobrara, Platte, Kansas-Republican, Arkansas, Cimarron, Canadian, Red, and Washita
Rivers) are major tributaries to the Mississippi River. The southern portions of the subregion contain
the headwaters of the Rio Grande River (Pecos River) or rivers that flow directly into the Gulf of Mexico
(the Guadalupe, San Antonio, Colorado of Texas, Brazos, and Nueces Rivers). Some rivers in the northern
portions of the glaciated prairie flow north, eventually into the Hudson Bay (notably the Red River of the
North).
The climate in this region ranges from semiarid in the western portions to moist subhumid in the
eastern portions. Mean annual precipitation ranges from 200 to 1,000 mm y1 from west to east across
the Great Plains (Lauenroth et al., 1999). Potential evaporation typically exceeds precipitation
(Transeau, 1905,1935). Mean annual temperatures increase from north (4-8 °C) to south (16-20 °C;
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Lauenroth et al., 1999). Winters tend to be dry, with less than 20% of the annual precipitation (Borchert,
1950; Lauenroth et al., 1999; Boughton et al., 2010). Most precipitation falls in late spring and early
summer (Borchert, 1950; Lauenroth et al., 1999), and much of the summer precipitation results from
localized convective thunderstorms. Because of the region's geographic location relative to the Gulf of
Mexico and the Rocky Mountains, however, substantial interannual variation exists, particularly in
terms of summer rainfall deficit (Borchert, 1950).
B.4.2.2 Hydrology and Geomorphology
The hydrology of most prairie river networks is highly variable (Matthews, 1988; Brown and Matthews,
1995; Dodds etal., 2004). These systems are frequently subjected to the extremes of drying and
flooding, and intermittent or flashy hydrology is prevalent in river networks throughout most of the
Great Plains (Matthews, 1988; Zale etal., 1989; Poff, 1996; Dodds etal., 2004). The topology of most
prairie river networks is dendritic due to the relatively flat landscape and uniform geology (Brown and
Matthews, 1995). Prairie river networks tend to have high drainage density (Section 2.4.2), and are
therefore efficient at transferring rainfall from uplands to downstream reaches (Gregory, 1976;
Osterkamp and Friedman, 2000). Flood magnitudes tend to be higher in the semiarid Great Plains than
in other regions, despite comparable rainfall intensities, due to low infiltration and vegetation
interception (Osterkamp and Friedman, 2000). Although floods tend to occur in late fall through late
spring, they can occur any time during the year (Brown and Matthews, 1995; Poff, 1996). Like most
river networks, those draining prairie landscapes often contain ephemeral, intermittent, and perennial
streams. Although many headwater prairie streams are ephemeral or intermittent (Matthews, 1988;
Brown and Matthews, 1995; Dodds etal., 2004), some have perennial spring-fed reaches located at the
network origins or distributed between intermittent reaches along headwater streams (Matthews et al.,
1985; Sawin et al., 1999; Dodds et al., 2004; Bergey et al., 2008).
The flow regimes of streams draining the Rocky Mountains, Black Hills, and northern prairies are largely
tied to snowmelt. Most systems originating in the mountains quickly transition in flow and morphology
as they cross the Great Plains, becoming intermittent and then slowly gaining flow from large streams
before joining the Mississippi River (Brown and Matthews, 1995). Some areas, however, have stable
streamflow with few intermittent streams because flow is derived from large, permeable ground-water
sources (e.g., Sand Hills in Nebraska; Winter, 2007).
The High Plains (Ogallala) aquifer system and other aquifers (e.g., Edwards-Trinity) are important
hydrologic features interconnected with Great Plains river networks. The High Plains aquifer system is
the largest (450,658 km2) and most intensively pumped U.S. aquifer, underlying much of the Great
Plains from southern South Dakota and southeastern Wyoming to central Texas (Sophocleous, 2005;
Ashworth, 2006; Sophocleous, 2010). The High Plains aquifer is composed of blanket sand and gravel
derived mainly from alluvial deposits and ancient marine sands. It is unconfined regionally, but locally
can be confined where beds of silt, clay, or marl are present. Regional movement of water through the
aquifer is from west to east, but locally the water moves toward major tributaries. Northern areas of the
Great Plain are underlain by glacial deposit aquifers that can be a mixture of till (unsorted material
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ranging from clay to boulders) and outwash (stratified sand and gravel) that was deposited by glacial
meltwater.
Most headwater streams originating in the prairie have riffle-pool morphology with alluvial gravel; only
headwater streams originating in the western mountains have high gradient, cobble-boulder channels
(Brown and Matthews, 1995). Southern prairie headwater streams tend to have finer substrate than
those in the northern and central Great Plains (Brown and Matthews, 1995). Larger streams tend to
have broad sand beds that are frequently braided (Section B.4.2.5). In contrast to headwater streams in
forested regions, the riparian areas of prairie headwater streams typically lack overhanging trees.
Grasses and shrubs are the dominant riparian vegetation, so channels lack woody debris and generally
receive direct sunlight. Because of intense flooding, prairie streams tend to form wide, deep channels
relative to their drainage areas, regardless of flow permanence (Hedman and Osterkamp, 1982; Brown
and Matthews, 1995). Because of similarity in topography, climate, geology, and soils, stream
geomorphology across the Great Plains is largely comparable (Miller and Onesti, 1988). High plains
channels, however, tend to be slightly steeper in gradient and more sinuous than wider and deeper
channels of the eastern plains (Miller and Onesti, 1988). During floods, the relatively incised channels
and lack of woody debris in prairie headwater streams make them less retentive of organic matter and
other materials than those of high-gradient forested channels; their pool-riffle morphology, high
sinuosity, and seasonal drying, however, can enhance retention (Brown and Matthews, 1995).
B.4.2.3 Physicochemistry
The factors discussed above are strong drivers of prairie stream physicochemistry (Matthews, 1988;
Brown and Matthews, 1995). Hot summers and cold winters in this region cause substantial direct and
indirect changes in water temperature, dissolved oxygen, and nutrient concentrations. Isolation of
surface water into pools during summer drying exacerbates these changes (Zale et al., 1989; Ostrand
and Marks, 2000; Ostrand and Wilde, 2004). For example, water surfaces can be covered with ice in
winter, whereas summer water temperatures can reach 35-40 °C with 9-10 °C diel (i.e., daily)
fluctuations (Matthews, 1988; Matthews and Zimmerman, 1990). Concomitant fluctuations in dissolved
oxygen occur, which when combined with stream respiration, contribute to dissolved oxygen values
approaching anoxic conditions.
Prairie rivers and streams naturally have higher concentrations of dissolved solids (e.g., calcium,
carbonate, bicarbonate, sodium, chloride, magnesium, sulfate) due to dissolution of the underlying
geologic layers (Huntzinger, 1995). Associated with these high levels of dissolved ions are elevated
alkalinity and pH. Mean total dissolved solids concentrations for many Great Plains rivers are among the
highest in the United States, exceeding 500 mg Ir1; many Great Plains rivers, however, also receive
anthropogenic total dissolved solid inputs from wastewater treatment effluents, agricultural runoff,
irrigation contributions to baseflow, and disposal of produced water associated with fossil fuel
production (Mathis and Dorris, 1968; Huntzinger, 1995; Farag et al., 2010). Some river networks, such
as the headwaters of the Red River in Texas and Oklahoma, are saline because they derive from brine
springs (Taylor etal., 1993).
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Streams and rivers of the central United States are often cited as having elevated nutrient (i.e., nitrogen
and phosphorus) loads. These loads are primarily attributable to nonpoint source runoff from fertilizer
application and livestock waste, especially during higher flows in winter and spring (Huntzinger, 1995;
Royer et al., 2006; Alexander et al., 2008). Data from streams draining native prairie indicate that
nitrogen and phosphorus concentrations and fluxes are lower or comparable to other intact ecosystems
(McArthur etal., 1985a; Dodds etal., 1996a; Kemp and Dodds, 2001).
B.4.2.4 Ecology
The low diversity of aquatic flora and fauna of prairie river networks, especially compared to
assemblages in the eastern and southeastern United States (Jewell, 1927; Fausch and Bestgen, 1997), is
likely due to the environmental instability of these river networks, their evolutionary history, and the
magnitude and extent of human alterations. Most organisms have adapted to erratic hydrologic regimes
and harsh physiochemical conditions in prairie streams by having rapid growth, high dispersal ability,
resistant life stages, fractional or extended reproduction (i.e., spawn multiple times during a
reproductive season), broad physiological tolerances, and life cycles timed to avoid predictably harsh
periods (Matthews, 1988; Dodds etal., 1996b; Fausch and Bestgen, 1997).
Algae are foundational components of prairie streams, acting to retain nutrients and provide an
important energy source to consumers (Gelwick and Matthews, 1997; Dodds et al., 2000; Evans-White et
al., 2001; Evans-White etal., 2003). Flooding and drying in prairie streams reset algal assemblages, spur
successional sequences, and maintain high levels of primary production (Power and Stewart, 1987;
Dodds etal., 1996b; Murdock et al., 2010). Algal assemblages are composed primarily of diatoms (e.g.,
Cymbella, Cocconeis, PinnuIaria,Achnanthes, Navicula, and Gomphonema], filamentous green algae (e.g.,
Cladophora, Spirogyra, Rhizoclonium, Stigeoclonium, Zygnema, and Oedogonium), and cyanobacteria (e.g.,
Oscillatoria, Nostoc).
Because of high light availability, algal primary production in prairie streams occasionally can be
substantially higher than in forested headwaters (Hill and Gardner, 1987a; Dodds et al., 1996b;
Mulholland et al., 2001; Bernot et al., 2010). Gallery forests farther downstream provide shade and
contribute organic matter. Shade from the gallery forests lowers light transmission to algae, resulting in
lower algal primary production in these reaches than in unshaded prairie headwater reaches. Thus, in
contrast to conventional longitudinal paradigms like the River Continuum Concept, the organic matter
driving prairie headwater streams derives mainly from within the channel (autochthonous production),
whereas leaf litter and other detritus from nearby gallery forests (allochthonous production) dominate
in intermediate-sized streams (Gurtz etal., 1982; Gurtz etal., 1988; Wiley etal., 1990). Despite having
greater primary production than forested headwaters, prairie streams—like forested ones—tend to be
net heterotrophic systems (Mulholland et al., 2001), but those that agricultural activities (e.g., elevated
nutrients, channelization) influence can at times be net auto trophic (Prophet and Ransom, 1974; Gelroth
and Marzolf, 1978; Wiley et al., 1990).
Invertebrates in prairie streams are represented by various aquatic insect groups (e.g., Diptera,
Coleoptera, Plecoptera, Ephemeroptera, Trichoptera), crustaceans (crayfish, isopods, amphipods),
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mollusks, and oligochaetes. Consumers of fine benthic organic matter, epilithic algae, and other
invertebrates tend to dominate invertebrate communities (Gray and Johnson, 1988; Harris et al., 1999;
Stagliano and Whiles, 2002). Diversity and abundance of invertebrates tend to increase with flow
permanence, but species composition generally highly overlaps, with intermittent stream assemblages
representing a nested subset of those from perennial streams (McCoy and Hales, 1974; Miller and
Golladay, 1996; Fritz and Dodds, 2002).
As with algae, flooding and drying are important drivers of invertebrate assemblages in prairie streams.
Distinct successional transitions are apparent following these disturbances (Chou et al., 1999; Fritz and
Dodds, 2002), and recovery to predisturbance levels can be rapid (Miller and Golladay, 1996; Miller and
Nudds, 1996; Fritz and Dodds, 2004). Woody debris is often rare in prairie streams, but where it is
present, invertebrates tend to be more abundant and more resistant to flooding, relative to those
associated with less stable sand and gravel substrates (Golladay and Hax, 1995; Hax and Golladay, 1998;
Johnson and Kennedy, 2003).
Fish are a well-studied component of river networks in the Great Plains, and are among the most
threatened (Rabeni, 1996; Fausch and Bestgen, 1997; Hubert and Gordon, 2007; Hoagstrom etal.,
2010). Approximately 200 fish species are found across prairie river networks, about 50 of which are
endemic to these streams. The most common taxa are minnows (Cyprinidae), suckers (Catastomidae),
darters (Percidae), sunfishes (Centrarchidae), and catfishes (Ictaluridae).
Longitudinal organization offish assemblages has been recognized widely in Great Plains river networks
(Harrell et al., 1967; Smith and Powell, 1971; Schlosser, 1987), and like macroinvertebrates these
assemblages often are nested such that intermittent headwater communities are subsets of those in
downstream perennial segments. Unlike algae and macroinvertebrates, fish inhabiting intermittent
headwater streams have no terrestrial or drying-resistant life stages. Fish, however, are highly mobile
and avoid desiccation by moving into downstream perennial reaches or perennial spring-fed pools in
upstream segments (Deacon, 1961; Fausch and Bramblett, 1991). Periodic floods are important for
creating perennial refuges and providing connectivity between habitats for the dispersal offish and
their eggs in prairie stream networks (Section B.4.3.3; Labbe and Fausch, 2000; Franssen et al., 2006).
B.4.2.5 Human Alterations
Human alterations to prairie river networks have affected physical, chemical, and biological connectivity
in these systems both directly and indirectly. Crop and livestock agriculture are predominant land uses
in the Great Plains (Galat et al., 2005; Matthews et al., 2005) and represent major nonpoint sources of
nutrients, sediment, and pesticides (Battaglin et al., 2003; U.S. EPA, 2006; Alexander et al., 2008).
Livestock concentrate in and near streams for shade, food, and water, leading to bank erosion, increased
soil bulk density, sedimentation, and elevated fecal bacteria concentrations (Armour et al., 1991; Strand
and Merritt, 1999).
To support these agricultural enterprises, water has been diverted from channels, withdrawn from
regional aquifers, and stored in reservoirs. Ground-water withdrawals in the Great Plains are the highest
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in the United States (Sophocleous, 2010), causing many once perennial river segments to regularly dry
up completely during summer months, particularly in the drier western portions of the Great Plains
(Cross and Moss, 1987; Ferrington, 1993; Falke etal., 2011). Nearly all river networks in prairie regions
have been altered by impoundments for irrigation storage and flood control, from small farm ponds in
headwaters to large reservoirs on river mainstems (Smith et al., 2002; Galat et al., 2005; Matthews et al.,
2005). Decline in flood magnitude, altered flow timing, and reduced flow variability and turbidity are
evident in many prairie rivers compared to historically documented conditions (Cross and Moss, 1987;
Hadley et al., 1987; Galat and Lipkin, 2000). Reductions in peak discharge derived from prairie streams
have contributed to the narrowing of the region's once broad and shallow river channels (Friedman et
al., 1998; Wohl et al., 2009). Dynamic mosaics of sand bars common in most prairie rivers have become
stabilized and coalesced islands. The establishment of trees along prairie river riparian zones was
limited by floods prior to settlement, but now dense zones of native and invasive trees and shrubs
further reduce flows through high evapotranspiration (Johnson, 1994; Dahm et al., 2002).
B.4.3 Evidence
B.4.3.1 Physical Connections
B.4.3.1.1 Water
As in other river systems, water is the primary medium by which materials are transported from
streams to rivers in prairie networks. Floods are common in Great Plains streams (Fausch and
Bramblett, 1991; Hill etal., 1992; Fritz and Dodds, 2005), and propagation of these floods from streams
to downstream rivers demonstrates hydrologic connectivity. Fritz and Dodds (2004, 2005)
characterized the hydrology of intermittent streams draining native tallgrass prairie in a study that
coincided with the highest flow on record (on May 13,1995, with a return interval of at least 50 years).
Kings Creek and one of its headwater streams (NO IB) are both headwater streams draining into the
Kansas River, downstream of the USGS gaging station at Fort Riley and upstream from the confluence of
the Big Blue and Kansas Rivers and the USGS gaging station at Wamego (Figure B-3). The peak-flow
rising and descending limbs were very rapid at Kings Creek and NO IB compared to those recorded for
the Kansas River at Wamego, where the peak arrived approximately 12 hours later (Figure B-4).
Hydrographs for the upstream Fort Riley gage on the Kansas River and the Big Blue River indicate that
the May 13,1995 peak at the downstream Wamego gage was associated with floods propagating from
Kings Creek and other small streams (Figure B-4). The subsequent peak at the Wamego gage that
occurred 5 days later was associated with a storm mainly affecting portions of the Kansas River basin
upstream of the Fort Riley gage, which elicited only a slight increase in discharge at Kings Creek and
N01B (Figure B-4).
A flood occurring June 14-20,1965 on the Platte River (Colorado and Nebraska) is among the largest
U.S. floods in recorded history, with a recurrence interval of 900 to 1,600 years (Matthai, 1969). This
flood originated from runoff of intense rainfall (360 mm in 4 hours) over headwater portions of the
drainage south of Denver, CO. Normal annual precipitation for this area is approximately 400 mm. Flows
in Plum Creek, one of the intermittent headwater streams to the Platte River that received the heaviest
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Figure B-3. Map showing the location of Kings Creek and N01B, intermittent tributaries to the Kansas
River.
• Big Blue @ Manhattan
Wamego
Kansas River@
Wamego
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ft Kansas River @ Ft. Riley
Kings Creek
• N01B
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Natural Area
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§ Kings Creek
Watershed, KS
rains, rose from <5 m3 s-1 to 4,360 m3 s-1 in only 40 minutes. Under the Federal Flood Control Act of
1944, detention impoundments were extensively constructed on headwater streams in the Great Plains
to retard flooding in downstream rivers (Schoof et al., 1978; Van Haveren, 1986). Headwater
impoundments reduced runoff to the Washita River in Oklahoma by 36%, but channel dredging of
streams offset these reductions by increasing flow from ground water and reducing transmission loss
(Schoof etal., 1978).
Machavaram et al. (2006) examined hydrologic connectivity between intermittent prairie streams, a
headwater pond, and a perennial stream reach approximately 10 km downstream using chemical and
isotopic tracers in a southeastern Kansas system. They found that, following precipitation, 20% of
downstream water originated from the upstream pond, fed by ephemeral and intermittent streams;
elevated oxygen stable isotope tracer associated with the pond water took 26-31 hours to reach the
downstream site (Machavaram etal., 2006). Streams connected to lakes and wetlands contributed
proportionally more flow to a southeastern Minnesota river in summer, when other water sources were
minimal, than in spring (Lenhart et al., 2010). Flow from these streams has a delayed or lagged release
because of storage in lakes and wetlands, and stream flow backed up because of high mainstem flows
(Lenhartetal., 2010).
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Figure B-4. Hydrographs (instantaneous and daily mean) showing propagation of the 13 May 1995
(Julian date 133) flood downstream from headwater sites (N01B and Kings Creek) to the Kansas River
at Wamego. Also shown are hydrographs from upstream gages on the Kansas River at Fort Riley and
the Big Blue River (see Figure B-3 for all site locations). Instantaneous data were not available at Kings
Creek immediately following the flood because of damage to the USGS gage and were not available
from Big Blue River. The peak instantaneous discharge for Kings Creek was estimated by USGS.
N01b (1.18km2)
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Julian date
Big Blue Riser at Manhattan (24, 968 km2)
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50 132 134 136 138 140
Julian date
8.4.3.1.2 Temperature (heat energy)
Water temperatures represent a substantial stress to biotic communities in Great Plains rivers (Section
B.4.3.3). Rivers to the north experience cold winters, and those to the south and west experience hot
summers. Streams, particularly those strongly connected to more stable ground water, can provide
thermal refuges for avoiding temporary hypothermic and hyperthermic stress (Section B.4.3.3.2). Wide,
shallow channels with little overhead canopy can result in high water temperatures under summer low
flows. Over a 1-km reach of the South Canadian River in Oklahoma, summer (August 18-19,1976)
maximum mainstem water temperatures were 36-37 °C, with cooler water (32-35 °C) in backwater
pools and a tributary stream (Matthews and Zimmerman, 1990). Mean water temperatures of seven
streams immediately upstream from confluences with the Missouri River (at the Kansas-Missouri
border) did not differ from water temperatures in the mainstem river, 200-300 m downstream of the
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confluences, except during March when streams were warmer than the river (Braaten and Guy, 1999).
Mean water temperature was determined to be homogeneous with no relationship between drainage
area and water temperature across two agriculturally dominated drainages in Illinois, where most flow
was derived from surface and shallow subsurface runoff (agricultural tiles) rather than deeper ground
water [Wiley et al., 1990).
B.4.3.1.3 Sediment
Great Plains rivers are naturally turbid (Jewell, 1927; Cross and Moss, 1987; Huntzinger, 1995), with
suspended sediment derived from the fine soils through which these river networks flow. Turbidity and
suspended sediment concentration increase in prairie networks with increasing discharge and drainage
area (Hill and Gardner, 1987b; Wiley et al., 1990; Lenhart et al., 2010), and can vary seasonally (Lenhart
et al., 2010). Seasonal turbidity levels at tributary outlets and nearby mainstem reaches, however, were
not related across seven Missouri River confluences in Kansas and Missouri (Braaten and Guy, 1999),
suggesting that these streams did not influence river turbidity at baseflow conditions. In contrast to
other studies in the prairie region, no relationship was found between suspended particle concentration
and stream size among 22 sites ranging in land use and network position (second- to eighth-order) in
the Kansas River basin (Whiles and Dodds, 2002). A significant positive relationship did exist when the
authors excluded suburban sites and sites influenced by impoundments. Concentrations of suspended
fine inorganic and organic matter were highest in the smallest stream draining suburban land use,
whereas a comparably small stream draining native tallgrass prairie had among the lowest
concentrations (Whiles and Dodds, 2002).
The downstream transport of metal-contaminated sediment was documented from mine tailings near a
South Dakota headwater stream down through the river network to a reservoir approximately 200 km
downstream, at the confluence of the Cheyenne and Missouri Rivers (Horowitz etal., 1988; Marron,
1989). The total amount of mine tailings transported from the headwater stream to downstream waters
and floodplains over a 100-year span was estimated to be approximately 100 million metric tons
(Marron, 1989). Contributions from streams to large rivers can therefore depend on the quantities
available for transport from headwater streams from surrounding land uses.
B.4.3.2 Chemical Connections
8.4.3.2.1 Nutrients and other chemicals
Studies show that chemical constituents are exported from small prairie streams (Dodds et al., 1996a)
and these chemical connections, or the downstream, flow-associated transport of nutrients, ions,
dissolved and particulate organic matter, and other substances along prairie stream drainage networks,
can significantly influence downstream water quality (Kemp and Dodds, 2002; Dodds etal., 2004; Dodds
and Oakes, 2006).
Small prairie streams also can be important in preventing downstream nutrient transport. Studies
conducted in Kings Creek, a stream draining a 1,060-ha tallgrass prairie catchment in Kansas, indicate
that small prairie streams are highly nitrogen retentive (Tate, 1990; Dodds et al., 1996a; Dodds et al.,
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2000). For example, Dodds etal. (1996a) found that nitrogen transport through four second- and third-
order streams in the Kings Creek watershed ranged from 0.01 to 6.0% of the total nitrogen supplied by
precipitation, the balance being retained by the stream system. Similar patterns of nutrient retention
have been demonstrated at larger spatial scales, as well. Alexander et al. (2000; 2008) modeled the
contribution of different-sized streams and rivers (including prairie streams) to nutrient loading in the
Gulf of Mexico. They found that large rivers deliver more of their nitrogen and phosphorus loads to the
Gulf of Mexico than small streams, largely due to increased instream nutrient uptake and removal by
small streams (Alexander et al., 2000). Despite their relative retentiveness, however, small streams do
make substantial contributions to downstream nutrient loading due to their large numbers, with small
to mid-sized streams in the western regions of the Mississippi River basin (which includes the Great
Plains) delivering approximately 25-50% of their nitrogen loads to the Gulf (Alexander et al., 2008).
Correlations between water quality and upstream land use also indicate that prairie stream headwaters
affect downstream reaches. Dodds and Oakes (2006, 2008) examined relationships between water
quality and watershed land use at different spatial scales, along one fifth-order prairie stream network
(Dodds and Oakes, 2006) and across 68 small prairie streams (Dodds and Oakes, 2008) in eastern
Kansas. In the single drainage study, they found that concentrations of total nitrogen and nitrate were
significantly related to riparian cover in the 2 km upstream of sampling sites, even when controlled for
catchment land cover at each site (Dodds and Oakes, 2006). In the cross-drainage study, riparian cover
along first-order streams was more closely correlated with total nitrogen, nitrate, ammonium, total
phosphorus, atrazine, dissolved oxygen, and fecal coliform concentrations than riparian cover 2 or 4 km
immediately upstream of sites across the 68 drainages (Dodds and Oakes, 2008). Nutrients are elevated
in most prairie streams and rivers and nutrient concentrations in these systems are related to nonpoint
land uses (Dodds and Oakes, 2004). These, along with widespread nature of headwater streams in river
networks, are highly indicative that streams have strong chemical connection, functioning as important
links between the surrounding lands to downstream waters.
Because prairie streams frequently experience intermittent flow, their influence on downstream waters
is often discharge-dependent and temporally variable. For example, nitrate concentrations tend to be
higher in intermittent prairie streams immediately after flows resume, versus when flow recedes (Tate,
1990). In addition, nitrogen uptake lengths (Dodds etal., 2000) and total phosphorus loads (Banner et
al., 2009) increase with discharge. The effect of precipitation-driven flows on downstream water quality
can depend on the relative contributions of surface water delivered from upstream channels and ground
water. Prairie streams typically are tied closely to ground-water sources (Section B.4.2.2), so the
influence of headwaters can be especially pronounced during periods of high precipitation. Kemp and
Dodds (2001) found that nitrate concentrations in fourth- and fifth-order lowland prairie reaches were
lowest during periods of high precipitation, when more low-nitrate water was delivered downstream
from second- and third-order reaches and high-nitrate ground-water influences were minimized.
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8.4.3.2.2 Dissolved and participate organic matter
Differences in DOC inputs along the prairie stream longitudinal gradient provide further indirect
evidence of chemical connections between prairie stream headwaters and downstream reaches.
McArthur et al. (1985b) isolated bacteria from stream sediments of grassland reaches and gallery forest
reaches of a prairie stream and exposed them to leachates derived from grasses and bur oak (a common
gallery forest species). Grassland bacteria only grew when provided with grass leachates as a carbon
source, whereas gallery forest bacteria grew when provided with either grass or bur oak leachates. This
finding suggests that either (1) grass-derived DOC-consuming bacteria are transported downstream and
then coexist with bacteria consuming forest-derived DOC, or (2) grass-derived DOC is transported
downstream, and local bacterial communities have adapted to use more refractory DOC exported from
upstream reaches (McArthur etal., 1985b).
Studies measuring POM exported from low-order prairie stream reaches show significant temporal and
spatial variability. For example, Golladay (1997) documented little POM export from a third-order
prairie stream in Kansas, whereas two prairie streams in Texas had much higher rates of POM transport
(Hill and Gardner, 1987b). In part, these differences might reflect variability between stoonflow and
baseflow sampling, as organic matter concentrations can be positively correlated with stream discharge
(Hill and Gardner, 1987b; Golladay, 1997). Whiles and Dodds (2002) examined seston (suspended fine
particles) dynamics along the Kansas River drainage network (second- to eighth-order), and found that
seston concentrations showed a significant positive relationship with stream size, increasing
approximately 17-fold along the longitudinal gradient. This increase in seston was correlated with an
increase in the taxa richness of filter-feeding invertebrates (Whiles and Dodds, 2002), illustrating that
detrital transport along the stream gradient can influence invertebrate assemblages, which is a basic
tenet of the River Continuum Concept (Vannote et al., 1980).
Stagliano and Whiles (2002) found that the standing stock of fine particulate organic matter FPOM in a
perennial reach of a tallgrass prairie stream was insufficient to support the annual secondary
production (i.e., the rate of heterotrophic biomass formation) of collector-gatherers (Cummins and Klug,
1979), the dominant group of macroinvertebrates feeding on deposited FPOM. The replenishment of
FPOM standing stocks, at least in part from upstream sources via algal senescence, the transport and
settlement of suspended POM, and the breakdown and transport of coarse POM, likely accounted for this
apparent imbalance: Turnover of FPOM standing stocks was estimated to occur every 20 days (Stagliano
and Whiles, 2002). Whiting et al. (2011) examined organic matter dynamics and trophic structure along
a tallgrass prairie stream network (first- to fifth-order). They found that collector-filterers
(macroinvertebrates that feed on suspended POM; Cummins and Klug, 1979) in upstream reaches
consumed <1% of suspended POM flux; gatherers that feed on fine and very fine POM dominated
secondary production in downstream reaches; and predators in downstream forested reaches
consumed 107% of locally derived macroinvertebrate production. Predators in the upstream and
middle reaches consumed 65% and 74% of available macroinvertebrate production, respectively. These
findings support the idea that downstream secondary production depends in part on the export of
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energy sources (i.e., POM and invertebrates) from upstream reaches (in order for downstream
predators to consume >100% of locally derived macroinvertebrate production).
As discussed earlier (Section B.4.2.4), prairie stream headwaters typically are open-canopied systems
that receive little organic matter from terrestrial inputs, relative to forested headwaters (Jewell, 1927).
Given the importance of autochthonous production in these systems, that algal-based contributions to
prairie stream seston can be significant (Swanson and Bachmann, 1976; Hill and Gardner, 1987b;
Lenhart et al., 2 010) is not surprising. In four Iowa streams, export of chlorophyll a (a measure of algal
biomass) was positively correlated with upstream channel bottom area, suggesting that downstream
suspended algae originated as benthic algae in upstream portions of the network (Swanson and
Bachmann, 1976). This downstream transport of algae also can provide colonists for downstream
reaches after flooding or drying of stream channels. For example, Dodds et al. (199 6b) examined the
recovery of periphyton biomass upon channel rewetting in an intermittent prairie stream. Within 2
weeks, chlorophyll had returned to maximum levels on rocks placed in the stream, even when the rocks
had been treated and scrubbed to remove desiccation-resistant propagules; this finding suggests that
algal colonists in this stream were transported downstream from permanent upstream pools (Dodds et
al., 1996b).
Coarse POM can connect prairie stream headwaters to downstream reaches. Johnson and Covich (1997)
examined detrital inputs along a second- to fifth-order prairie stream network in Oklahoma. They found
that leaves in the stream originated from farther upstream than expected, with the percentage of whole
leaves at a site best explained by riparian forest cover in reaches 500 and 1,000 m upstream. The
percentage of leaf fragments >1 mm was best explained by downstream distance along the stream
network (Johnson and Covich, 1997), suggesting increased processing and fragmentation of leaves as
they move down the longitudinal gradient.
B.4.3.3 Biological Connections
B.4.3.3.1 Invertebrates
Existing evidence for invertebrate-mediated biological connectivity along prairie stream networks
mainly comes from studies of invertebrate assemblage recovery following flooding and drying in small
prairie streams. Recovery from these disturbances tends to be relatively rapid, with substantial gains in
invertebrate taxa richness and density observed within days to weeks (Miller and Golladay, 1996; Hax
and Golladay, 1998; Fritz and Dodds, 2004), suggesting that these reaches are quickly repopulated by
invertebrate drift from upstream sources, aerially dispersing adults, or disturbance-resistant survivors.
Fritz and Dodds (2002, 2004, 2005) examined postflooding and postdrying recovery of invertebrates in
small intermittent and perennial prairie streams along an approximately 5-km stretch of Kings Creek in
Kansas. They found that initial recovery of invertebrate taxa richness in intermittent reaches, and taxa
richness of invertebrate drift and aerially colonizing insects, were negatively related to distance from
upstream perennial water (Fritz and Dodds, 2002, 2004). Distance from upstream refuges, however,
was not a significant predictor of invertebrate diversity measures across annual time scales (Fritz and
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Dodds, 2005); they speculated that movement of water along the entire stream network (i.e.,
maintenance of hydrologic connectivity) makes proximity to colonists less important over longer time
scales. These findings suggest that recovery from disturbance in these systems depends on biological
connectivity via both downstream drift of colonizers and downstream (and potentially upstream)
movement of aerially dispersing, egg-depositing adults (Miller and Golladay, 1996; Dodds etal., 2004).
B.4.3.3.2 Fishes
Research on fish assemblages in prairie streams provides perhaps the strongest and most well-studied
evidence of biological connections throughout these networks. Much of this evidence focuses on two
related aspects of the ecology of prairie stream fish: the dispersal and recruitment of pelagic-spawning
prairie stream fish and the recovery offish assemblages after disturbance, especially flooding and
drying.
Many prairie stream fish broadcast spawn nonadhesive, semibuoyant eggs, which develop (typically
hatching within 1 to 2 days) as they are transported downstream with water flow (Cross and Moss,
1987; Fausch and Bestgen, 1997; Platania and Altenbach, 1998; Durham and Wilde, 2006). The distance
these eggs travel downstream depends on discharge and several other factors (e.g., development time);
Platania and Altenbach (1998) estimated, however, that unimpeded eggs could travel as far as 144 km
before hatching, and another 216 km as developing protolarvae (i.e., the swim-up stage), illustrating that
downstream transport of these drifting organisms can be extensive. Without adequate water flow along
sufficient lengths of the stream network, eggs can drop out of suspension before hatching (Platania and
Altenbach, 1998; Durham and Wilde, 2006). Based on historical and contemporary fish surveys, eight
species of pelagic-spawning cyprinids require a minimum length of greater than approximately 100 km
(ranging from 103 to 297 km, depending on the species) of undisrupted stream channel (e.g., channels
with no impoundments and no drying associated with human withdrawal) to support persistent
populations (Perkin and Gido, 2011).
This pelagic-spawning reproductive strategy also necessitates upstream movement by adult fish, if
populations are to be maintained in small prairie streams (Fausch and Bestgen, 1997; Durham and
Wilde, 2008). Prairie stream fishes generally are highly vagile, with adults capable of long-distance
migrations. For example, individuals of one species of prairie fish (Hybognathus placitus) in the South
Canadian River, NM were observed to move approximately 250m upstream over a 15-minute period,
illustrating that prairie fishes can move substantial distances over relatively short periods (Fausch and
Bestgen, 1997).
The effect that impoundment of prairie streams and rivers has had on the region's native fish
assemblages highlights the importance of hydrologic connectivity in these systems. Many studies have
documented statistically significant associations between impoundment of prairie streams and loss of
native fishes (Winston et al., 1991; Luttrell et al., 1999; Schrank et al., 2001; Falke and Gido, 2006;
Matthews and Marsh-Matthews, 2007). For example, Schrank etal. (2001) found that, across 26 streams
in the Flint Hills region of Kansas, sites from which Topeka shiners [Notropis topekd) had been
extirpated had significantly more small impoundments on them and higher largemouth bass
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(Micropterus salmoides) catch-per-unit-effortthan sites at which the shiners were extant. Fewer studies
have specifically examined the mechanisms by which impoundments affect these changes, although
impoundments likely disrupt both the downstream transport of developing eggs and larvae (Platania
and Altenbach, 1998) and the upstream and downstream movement of adult fish.
Because many small prairie streams have intermittent flow, maintenance offish populations often
depends on dispersal out of intermittent reaches before drying occurs and recolonization of these
habitats once water flow resumes—both of which require hydrologic connectivity along the stream
network. Many fishes also require different habitats during different life stages, further necessitating
hydrologic connectivity across these areas (Labbe and Fausch, 2000; Falke etal., 2010).
For dispersal and recolonization to occur, fishes must be able to access refuge habitats under adverse
conditions, and then expand into newly habitable areas once adverse conditions abate. Small, spring-fed
prairie streams serve as key refuges for endemic prairie fishes (Hoagstrom et al., 2010), because they
are ground water fed and maintain permanent pools that can provide habitat during periods of channel
drying (Wohl et al., 2009). This ground-water influence also allows these spring-fed streams to provide
refuge from adverse temperatures. For example, a spring-fed stream in Missouri had more stable
temperatures than the mainstem river, with cooler summer and warmer winter temperatures; in winter,
fish from the mainstem river moved into this habitat, where their food availability, growth, and average
egg size were greater than those offish that stayed in the mainstem (Peterson and Rabeni, 1996).
During and after floods, juvenile and adult fishes can move upstream or downstream (or get displaced
downstream) into newly available habitat (Fritz et al., 2002; Franssen et al., 2006). Once channels are
rewetted, prairie stream fishes can move quickly into these previously unoccupied habitats (Harrell et
al., 1967; Fritz et al., 2002; Franssen et al., 2006). For example, Harrell et al. (1967) examined fish
response to channel drying in third- to sixth-order reaches of Otter Creek, an intermittent prairie stream
in north-central Oklahoma, and found that most fish species collected after 8 months of flow prior to
channel drying were already present 3 days after channel rewetting (Harrell et al., 1967). After a flood in
an intermittent prairie stream in Kansas, fish dispersed into the headwaters from a perennial reach
approximately 5 km downstream (Franssen etal., 2006).
B.4.4 Prairie Streams: Synthesis and Implications
Prairie streams typically represent a collection of spring-fed, perennial pools and reaches, embedded
within larger, intermittently flowing segments (Labbe and Fausch, 2000). Due to the region's geographic
location, substantial interannual variation in rainfall exists. Expansion (flooding) and contraction
(drying) of these systems, particularly in terms of summer rainfall deficit (Borchert, 1950), determine
the timing of hydrologic connectivity at any given time. Because of this temporal variability, connectivity
in prairie river networks must be considered over relatively long time scales (multiple years).
• Studies have demonstrated significant physical, chemical, and biological connections from
prairie headwater streams to larger rivers, despite extensive alteration of historical prairie
regions by agriculture, water impoundment, water withdrawals, and other human activities
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(Matthews and Robinson, 1998; Dodds et al., 2004), and the challenges these alterations create
for assessing connectivity.
• The most compelling evidence for connectivity along prairie river networks comes from
examples of streams as sources of water via flood propagation (Matthai, 1969; Fritz and Dodds,
2004, 2005), sources of contaminated sediment transport (Horowitz etal., 1988; Marron, 1989),
sites of nutrient lags and transformation (Dodds et al., 1996a; Alexander et al., 2008), the
downstream transport of prairie fish eggs and larvae (Platania and Altenbach, 1998; Perkin and
Gido, 2011), and refuges for prairie fishes (Fausch and Bestgen, 1997; Franssen etal., 2006).
• Impoundments for irrigation storage and flood control have altered flood magnitude, altered
flow timing, and reduced flow variability and turbidity across the prairie regions (Cross and
Moss, 1987; Hadley etal., 1987; Galatand Lipkin, 2000). The effect that impoundment of prairie
streams and rivers has had on the regions' native fish assemblages highlights the importance of
hydrologic connectivity in these systems. Maintenance offish populations often depends on
dispersal out of intermittent reaches before drying occurs and recolonization of these habitats
once water flow resumes—both of which require hydrologic connectivity along the stream
network—and many fishes also require different habitats during different life stages (Labbe and
Fausch, 2000; Falke etal., 2010).
B.5 Case Study: Southwestern Intermittent and Ephemeral
Streams
B.5.1 Abstract
Ephemeral and intermittent streams are abundant in the arid and semiarid landscapes of the West and
particularly the Southwest (Figure B-5.). These areas are characterized by low and highly variable
precipitation where potential evapotranspiration exceeds precipitation. Based on the National
Hydrography Dataset; 94%, 89%, 88%, and 79% of the streams in Arizona, Nevada, New Mexico, and
Utah are intermittent or ephemeral (NHD, 2008). The heavily studied Upper San Pedro Basin in
southeastern Arizona is discussed in detail because it is a well-understood example of the hydrologic
behavior and connectivity of rivers common to the southwestern United States where ephemeral and
intermittent tributaries comprise the majority of the basin's stream reaches. Flows and floods from
ephemeral and intermittent streams are also major drivers of the dynamic hydrology of the relatively
few perennial reaches in the Southwest. These streams also supply water to mainstem alluvial aquifers
and regional ground-water aquifers. Both alluvial and regional aquifers, in turn, supply baseflow to
perennial mainstem stream reaches over extended periods (sometimes months) when little or no
precipitation occurs. It is this baseflow and shallow ground water that supports the limited naturally
occurring, vibrant riparian communities in the region. In addition, ephemeral streams export sediment,
which contributes to shaping the fluvial geomorphology and alluvial aquifers of streams in the regions
(Shaw and Cooper, 2008), and nutrients, which contribute to river productivity. Several studies found
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Figure B-5. Upper: Geographic distribution of intermittent and ephemeral (red) and perennial (blue)
streams in the Continental United States and two example watersheds in Arizona and
Michigan/Ohio/lndiana from the National Hydrography Dataset (NHD) stream map
(http://nhd.usgs.gov/). Lower: maps of mean precipitation and the precipitation coefficient of variation
(equal to the standard deviation divided by the mean) of annual precipitation from 1895 to 2012. Note
that the NHD might not accurately reflect the total extent of ephemeral or intermittent streams, as it
does not include stream segments less than 1.6 km (1 mile) long, combines intermittent and
ephemeral streams, and is based on l:100,000-scale topographic maps.
JS Streams
IrnermitKnl
— rttetia*
Pi ol 'A'alef sh*d
Cienega Creek Watershed
Percent Perennial = 1.7%
Iftc 1.;,, 7 r PP i prc-ip Mi nn I inn)
SI. Joseph's Watershed
Percent Perennial = 58.5%
1«*2011 Preclprattsn Coefficient »t Vulrion
TOO .*
««
am p
£00
IK*
-1JD -110
-to -at -TO
that native fishes and invertebrates are well adapted to the variable flow regimes common in rivers of
the Southwest and are heavily influenced by ephemeral tributary streams (Turner and List, 2007).
B.5.2 Introduction
This case study addresses the hydrologic and ecological influence of ephemeral and intermittent
streams on perennial or intermittent rivers in the arid and semiarid southwestern United States with
Stream and Wetland Connectivity:
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B-38
January 2015
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particular emphasis on Arizona and New Mexico. The structure of this case study differs slightly from
the other case studies because of the uniquely thorough understanding of one particular southwestern
river system, the San Pedro River, which has been the subject of a long-term research program
(Goodrich et al., 2000; Stromberg and Tellman, 2009). Hence, evidence for the function and connectivity
of ephemeral and intermittent tributaries to the San Pedro River is described in detail, and its
application to other river systems in the South west is subsequently explored.
B.5.3 Southwestern Rivers
Understanding the unique characteristics of southwestern American rivers is necessary to evaluate the
connectivity and influence of ephemeral and intermittent streams on these rivers (Levick et al., 2008).
Southwestern rivers differ in many ways from rivers in the humid eastern United States or in the
Midwest and West. Southwestern rivers typically can be divided into two main types, particularly in the
Basin and Range geologic province. The first type comprises rivers in the mountainous upper basins that
receive more precipitation, often as snow, and the second type comprises those rivers located in the arid
or semiarid plateau regions and valley plains dominated by ephemeral streams (Blinn and Poff, 2005).
For example, more than 80% of the Gila River corridor in New Mexico and Arizona meanders through
desert scrublands. Precipitation is seasonal. In summer, precipitation is strongly influenced by
atmospheric moisture flowing from the Gulf of Mexico and the Gulf of California (Mexican monsoon),
where local heating triggers high-intensity air-mass thunderstorms. In fall, tropical depressions, often
remnants of hurricanes, can bring infrequent but long-duration rainfall events; such storms are
responsible for many of the larger floods in the region (Webb and Betancourt, 1992). Cyclonic storms
from the Pacific Ocean, resulting in large frontal systems, dominate winter precipitation in the form of
snow in higher elevations and typically as low-intensity rainfall in lower elevations (Blinn and Poff,
2005). Figure B-6 illustrates the 2003 calendar year hydrograph from the White River near the Fort
Apache USGS gaging station (upper) in east-central Arizona, and the San Pedro River near Tombstone, in
southeastern Arizona (lower). Although the two gaging stations differ in elevation by less than 200 m,
the watershed contributing to the White River is substantially larger and is higher in elevation than the
San Pedro watershed, resulting in long-duration spring runoff from snowmelt. Monsoon-generated,
short-duration runoff dominates the San Pedro watershed butmonsoonal influence also is apparent in
the White River hydrograph. Runoff generated from late monsoon precipitation in September caused a
major increase in discharge in the White River and a minor increase in the San Pedro. Most perennial
and intermittent rivers in the Southwest are ground water dependent, flowing primarily in a baseflow
regime and supported by discharge from a connected regional or alluvial aquifer or both. As discussed in
more detail below, part of the baseflow is often sustained or augmented by slow drainage of a shallow
alluvial aquifer from past flooding. In arid and semiarid regions, the riparian areas that perennial and
intermittent streams support occupy a small percentage of the overall landscape but they host a
disproportionately greater percentage of the biodiversity than the areas surrounding them (Goodrich et
al., 2000; Stromberg et al., 2005). Reservoir construction, irrigation withdrawals, and the cumulative
impacts of ground-water pumping have converted many historical, perennially flowing reaches into
intermittently flowing reaches (Blinn and Poff, 2005).
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
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Figure B-6. 2003 calendar year hydrographs from (a) the White River near Fort Apache, AZ and (b) the
San Pedro River near Tombstone, AZ.
1000
800
600
Q
400
CD
Q
200
500
^400
fr
| 300
6
c
I 200
r=
100
(a) WHITE RIVER NEAR FORT APACHE
Jan Feb
Nov Dec
(b) SAN PEDRO RIVER NEAR TOMBSTONE
Tropical
Storm
A
Baseflow
Monsoon
Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec
Abrupt changes in streamflow regimes (i.e., a change from perennial to intermittent or ephemeral and
back again) can also result from underlying geology. Streams with abrupt changes are often referred to
as interrupted streams (Meinzer, 1923; Hall and Steidl, 2007). A constriction and rise in bedrock geology
can force regional ground water to the surface resulting in perennial flow while streamflow
encountering highly fractured bedrock or a highly porous karst system can virtually disappear over very
short distances. Another relatively abrupt transition in arid and semiarid stream hydrology and
morphology occurs where steep mountain slopes transition into lower valley slopes. At this transition,
watersheds with high sediment transport out of the mountainous portion often form alluvial fans. The
stream channel system above the transition is typically dendritic and below the transition, the channel
Stream and Wetland Connectivity:
A Review and Synthesis
B-40
January 2015
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system often becomes a diffusive set of shallow braided channels. Runoff over alluvial fans typically
becomes less concentrated or confined to a single large channel but more diffuse and shallower turning
into broad sections of sheet flow (Parker etal., 1998). The diffuse runoff is more likely to infiltrate into
the alluvial fan. Very large flows may be required for runoff to cross the alluvial fan and connect to
downstream waters.
Dominant hydrologic flowpaths vary with location within southwestern river basins. After climate and
weather, recharge and infiltration mechanisms are the next most important factors determining the
occurrence of ephemeral, intermittent, and perennial streams. Recharge over longer time scales
(months to centuries) is essential to replenishing regional ground water and near-stream alluvial
aquifers, which in turn are essential to maintaining baseflow in perennial streams. Primary recharge
mechanisms include mountain block recharge, mountain front recharge, diffuse hillslope or interchannel
recharge, and ephemeral channel recharge. Key advances brought forth in a recent synthesis of research
on ground-water recharge in the southwestern and western United States include (1) desert vegetation
effectively eliminates diffuse recharge in most areas of the basin floor; (2) ephemeral channel recharge
can be very important in wet years and greatly dominates recharge in basin-floor environments; and
(3) environmental tracers are now available to "fingerprint the sources and amounts of ground-water
recharge at the basin scale" (Phillips et al., 2004).
Mountains with deeper soils or those consisting of fractured rock will have higher infiltration capacities,
less frequent occurrences of overland flow, and serve as recharge areas for regional ground water
(Wilson and Guan, 2004; Blasch and Bryson, 2007; Wahi et al., 2008). Mountains with shallow soils and
more consolidated rock will shed stoonflow and shallow ground water off the mountain block onto the
valley, which often consists of deep alluvium, particularly in the basin and range geologic province. This
area is where mountain-front recharge occurs. High-elevation perennial streams often become
intermittent or ephemeral at this transition, with their downstream disappearance of surface flow
dependent on the flow rates coming off the mountain block and the permeability of the valley alluvium
into which they enter. During periods of high flow, they can reconnect with other perennial stream
reaches maintained by ground-water flow (Blinn and Poff, 2005; Blasch and Bryson, 2007; Yuan and
Miyamoto, 2008).
Runoff generation in arid and semiarid valley floors and lowlands where basin alluvium is relatively
porous and deep is dominated by the infiltration-excess mechanism in which precipitation rates exceed
infiltration rates. In the arid and semiarid Southwest, high-intensity convective thunderstorms typically
trigger this situation. Generally, such storms are relatively short in duration, resulting in ephemeral
flows with short runoff duration (Goodrich et al., 1997). As water flows down dry ephemeral channels, it
infiltrates the channel bottom and sides (i.e., channel transmission losses occur) where channel
substrate is porous. If restricting soil or geologic layers underlying the channel do not substantially
inhibit downward motion, channel transmission losses will recharge either the regional or alluvial
ground water (Tang etal., 2001; Constantz etal., 2002; Harrington etal., 2002; Goodrich etal., 2004;
Goes and Pool, 2005; Vivoni et al., 2006; Blasch and Bryson, 2007). In this influent stream environment
typical of many southwestern streams, the volume of transmission water losses in ephemeral channels
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
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increases as watershed size increases, resulting in a losing stream environment as opposed to a gaining
stream environment encountered in wetter hydroclimatic regimes (Goodrich et al., 1997). As noted
above and discussed in Phillips et al. (2004), these ephemeral tributary channels are the dominant
source of recharge in valley floors, and at the basin scale they can provide substantial recharge during
wet years. Typically, as stream drainage area increases, the alluviums under and next to streams begin
to serve as important shallow aquifers that receive and store streamflow infiltration during hydrologic
events and sustain baseflow and riparian communities between storms (Stromberg et al., 2005; Baillie
et al., 2007; Dickinson et al., 2010).
The magnitude of aquifer recharge is highly temporally variable in the Southwest Winter precipitation,
which has a predominant effect on mountain-block and mountain-front recharge in the Arizona-New
Mexico portion of the Southwest, is correlated with the El Nino/Southern Oscillation (Woolhiser etal.,
1993) at interannual time scales. Over decadal climate cycles, winter precipitation also is related to the
Pacific Decadal Oscillation (Pool, 2005). The magnitude of ephemeral channel recharge varies widely
from year to year, depending on the strength of the monsoon season (Goodrich et al., 2004) and the
occurrence of relatively infrequent and prolonged precipitation events resulting from tropical
depressions. Floods and large runoff events caused by any of these mechanisms can have a long-lasting
influence (6 to 10 months) on baseflow of southwestern rivers by recharging near-stream alluvial
aquifers and thereby sustaining streamflow as they drain (Brooks and Lemon, 2007).
B.5.4 San Pedro River
B.5.4.1 Basin Characteristics
Because of a rich research and long-term monitoring history, the San Pedro Basin and River in
southeastern Arizona represents an excellent case study of the hydrologic behavior and connectivity of
southwestern rivers (Goodrich et al., 2000; Stromberg and Tellman, 2009; Brookshire et al., 2010). The
San Pedro River originates in Mexico, flowing undammed north to its confluence with the Gila River. The
San Pedro is the only significant un-impounded river in Arizona and the last remaining stream in
southern Arizona with long perennial reaches (Figure B-7; Kennedy and Gungle, 2010). Most tributaries
to the river are ephemeral at their confluence with the mainstem. The river basin, located in the Basin
and Range Province, has a valley that is generally 30-50 km wide, comprising sedimentary fill deposits,
and slopes upward from the river to mountains with elevation ranging from 2,000 to 2,900 m. The San
Pedro Basin consists of 93% nonperennial reaches (including ephemeral and intermittent), 6.3%
artificial path (canals, diversions, pipeline, connectors), and 0.7% perennial reaches in the U.S. portion of
the basin as derived from the USGS NHD1 (Figure B-8). The percentage of streams types is not static but
varies from year-to-year. The Nature Conservancy and its partners annually map the wet and dry
reaches along the San Pedro mainstem and several large tributary streams since 2007
on USGS National Hydrography Dataset (NHD) stream map (htt]3^/nhd.usgs.goy^). Note that the NHD
might not reflect the total extent of ephemeral or intermittent streams accurately, as it does not include stream
segments less than 1.6 km (1 mile) long, combines intermittent and ephemeral streams, and is based on
l:100,000-scale topographic maps.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
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Figure B-7. San Pedro River basin map showing major physiographic features and current and
historical perennial reaches. From Levick et al. (2008), courtesy of The Nature Conservancy, Arizona.
Available online at
http://azconservation.org/map_gallery/current_and_formerly_perennial_san_pedro_river_surface_water.
Current and Formerly
Perennial Streams in the
San Pedro River Watershed
-/\-~— Currently perennial
Historically perennial
/\f Major roads
The San Pedro Rher in the U.S. ha*
Ins! more than hair of its historic;)!
perennial surface water.
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(Turner and Richter, 2011). The wet-dry mapping is conducted roughly in the middle June, historically
the time of lowest streamflow, prior to the onset of the monsoon. For 2014, about 25% (54.1 km) of the
214 km surveyed were found to be wet (214 km is ~1.1% of the stream length plotted in Figure B-8).
Stream and Wetland Connectivity:
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B-43
January 2015
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Figure B-8. Perennial (blue) and nonperennial (red) streams in the San Pedro Basin from the U.S.-
Mexico border to its confluence with the Gila River based on USGS National Hydrography Dataset
(NHD) stream map (http://nhd.usgs.gov/).
NHD Streams
Nonperennial
Perennial
20 10
20 Kilometers
The wet-dry survey data is accessible at:
rhttp://azconservation.org/downloads/category/san pedro river).
Annual precipitation within the basin ranges from 300 to 750 mm with highest amounts occurring in the
mountains. Vegetation includes desert scrub, grasslands, oak woodland savannah, mesquite woodland,
riparian forest, coniferous forest, and agriculture (Kepner etal., 2000; Kepner etal., 2004). Brush and
grasses typical of southwestern semiarid landscapes (Goodrich et al., 1997) dominate the valley floor
vegetation.
Stream and Wetland Connectivity:
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B-44
January 2015
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At the Walnut Gulch Experimental Watershed (WGEW—a subwatershed of the San Pedro watershed
near Tombstone, Arizona), operated by U.S. Department of Agriculture, Agricultural Research Service
(USDA-ARS), approximately two-thirds of the annual precipitation on the watershed occurs as high-
intensity, convective thunderstorms of limited aerial extent (Goodrich etal., 1997). Winter rains (and
occasional snows) are generally low-intensity events associated with slow-moving cold fronts and are
typically of greater aerial extent than summer rains. Runoff on the lower elevations of the WGEW is
generated almost exclusively from convective storms during the summer monsoon season via
infiltration excess that produces overland flow. The hydrogeology of the San Pedro River basin is typical
of many alluvial basins in the Southwest (Dickinson et al., 2010). Ground water flows through the basin-
fill aquifer (regional aquifer) from recharge areas near the mountains and beneath ephemeral
tributaries to perennial reaches of the San Pedro River (Wahi et al., 2008; Dickinson et al., 2010). A
narrow band of highly permeable stream alluvium is incised into the basin-fill along the major stream
channels (Figure B-9). The stream and floodplain alluvium is an important alluvial aquifer that receives
discharge from the basin-fill aquifer and streamflow via streambank infiltration occurring during high
stream stages.
This bank and alluvial aquifer storage supports riparian vegetation during periods lacking runoff
(Dickinson et al., 2010). The San Pedro River network with associated shallow alluvial aquifers
(mainstem and portions of some tributaries) supports extensive riparian vegetation communities
(Stromberg et al., 2005) thatprovide habitat for more than 350 species of birds, 80 species of mammals,
and 40 species of reptiles and amphibians (Kennedy and Gungle, 2010). Alluvial aquifers also are zones
of extensive hyporheic exchange (Stanford and Ward, 1988; Fernald et al., 2001).
Figure B-9. Generalized east-west section and stratigraphic units in the middle San Pedro watershed.
From Dickinson et al. (2010).
msr
EXPLANATION
Stratigraphic unit;
I I Pre- and post-entrenchment stream alluvium --
I I Upper and lower basin fill sand and gravel
JITj Upper and lower basin fill silt and clay
!>'./ 1 Pre-Basin and Range sedimentary rock$
L—I Sedimentary rock
— P12; ometric surface
- Line separating upper and lower basin fill
Sedimentary rock 1 ^^
Crystalline rock J
B.5.4.2 Ephemeral Stream Connections and Their Influence on the San Pedro River
Overland runoff generation and associated ephemeral streamflow is common in San Pedro tributary
streams. Goodrich etal. (1997) examined hundreds of hydrologic events in different-sized catchments at
the USDA-ARS WGEW and found that the relationship between watershed area and runoff volume was
Stream and Wetland Connectivity:
A Review and Synthesis
B-45
January 2015
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increasingly nonlinear as drainage area increased. The authors found a critical threshold watershed area
of approximately 36-60 ha, at which runoff responses became much less linear and channel
transmission losses increased more rapidly with increasing watershed area. This relationship is very
different from commonly observed relationships in humid streams of the East, where runoff generally is
proportional to watershed area (Section 3.3.1). Two reasons for this variability in runoff produced per
unit watershed area are: (1) the spatial variability and limited spatial extent of runoff producing
precipitation, and (2) the loss of runoff by infiltration into the bed of ephemeral channels (transmission
losses). Figure B-10 illustrates this process. During a major rainstorm on 17 August 2 006, most of the
precipitation from multiple air-mass thunderstorm cells occurred over relatively localized areas in the
upper and lower portions of the USDA WGEW. As overland flow occurred and became concentrated in
the ephemeral tributary network, streamflow dramatically diminished as the runoff hydrograph
traveled downstream through the channel network. However, a substantial amount of runoff from this
storm traversed the ephemeral Walnut Gulch tributary and reached the mainstem of the San Pedro
River, augmenting the flow as measured at the USGS Tombstone stream gage. Runoff in Walnut Gulch
(149 km2 drainage area) and many arid and semiarid streams is characterized by short duration, highly
episodic flows. The longitudinal extent of the effects of these flows on downstream waters is a function
of the flow magnitude, its duration, the depth, conductivity and antecedent moisture conditions of the
ephemeral channel substrate that the runoff flows across, and the depth to ground water. For example,
in 2006 there were 23 runoff flows measured at Walnut Gulch flume 1 (the outlet of the WGEW). The
average volume, peak runoff rate, and duration of these runoff events was 31,460 m3, 7.23 m3/s, and
239 minutes, respectively. Four (4) of the 23 runoff events recorded at Flume 1 were estimated to have
measureable impacts on flows measured at the downstream USGS Tombstone stream gage (4510 km2)
on the San Pedro River (including the event shown in Figure B-10).
Evidence is strong that transmission losses in ephemeral tributary streams recharge alluvial and
regional aquifers (Goodrich et al., 1997; Callegary et al., 2007). Using three fundamental approaches to
estimate ephemeral channel recharge (1—closing the water balance for the channel reach, 2—
measuring changes in ground-water volume directly [well levels] or indirectly [microgravity], and 3—
using geochemical tracers), Goodrich et al. (2004) estimated that during the relatively wet 1999 and
2000 monsoon seasons, regional aquifer ground-water recharge from ephemeral streams ranged from
approximately 15 to 40% of total average annual recharge as estimated from a calibrated regional
ground-water model (Pool and Dickinson, 2007). During the dry monsoon seasons of 2001 and 2002,
limited ephemeral runoff and stream channel infiltration occurred, but no discernible deep aquifer
recharge was detected.
The influence of stoonflows from ephemeral tributary streams extends to the San Pedro River
mainstem. As stoonflow is exported from the tributaries to the mainstem and water moves
downstream, transmission losses and bank recharge occur within the mainstem river itself and supply
water to the alluvial aquifer of the mainstem (Kennedy and Gungle, 2010). Using geochemical tracers
(chloride, sulfate, and stable isotopes of hydrogen and oxygen in water), Baillie et al. (2007) found two
main sources of water in the alluvial aquifer for the upper San Pedro River: (1) regional ground water
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
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Figure B-10. Storm rainfall and downstream hydrographs with decreasing runoff volume and peak
rate due to channel transmission losses as measured by in the USDA-ARS Walnut Gulch Experimental
Watershed (WGEW) and the impact of this storm runoff on the San Pedro River in SE Arizona. Inset
photos show a typical air-mass thunderstorm and the front of surface flow progressing down an
ephemeral channel. Photo of ephemeral stream from Levick et al. (2008).
Walnut Gulch Experiment! Watershed
Rainfall Volume (mm)
Walnut Gulch Flume
Subwalerstied Boundary
USGS Gauging Station
recharged along the Huachuca Mountains (mountain block, mountain front) to the west, and (2) local
recharge from monsoon floodwaters. Alluvial ground-water composition varied between gaining and
losing reaches. Locally recharged floodwater comprised 60 to 85% of the alluvial ground water in losing
reaches but only 10 to 40% in gaining reaches. Baseflow also contained a significant component of
monsoon floodwater throughout the year, from 80% in upstream reaches to 55% after passing through
several gaining reaches.
Ephemeral tributary stormflows are also sources of sediment and alluvium for the main San Pedro
River. Only the largest, less frequent events can flush sediment completely through ephemeral
tributaries (Lane et al., 1997). For example, a reach-scale study in the WGEW estimated sand transport
distances of only 401 and 734 m in nine floods over two consecutive years (Powell et al., 2007). In
another study, Lekach et al. (1992) found thatmore than 90% of the bedload yield originated from the
mid-watershed channels during larger runoff events from an arid watershed in Israel. Ephemeral
tributary stormflows and their associated sediment loads influence the character of river floodplains
and alluvial aquifers (Nanson and Croke, 1992; Shaw and Cooper, 2008).
Extensive riparian plant communities along the mainstem San Pedro River depend on the availability of
water in the alluvial aquifer along the river, including water derived from ephemeral stream stormflows
(Stromberg et al., 2005; Baillie et al., 2007). These riparian areas, in turn, strongly influence river
Stream and Wetland Connectivity:
A Review and Synthesis
B-47
January 2015
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attributes through stream shading, channel stabilization, nutrient cycling, inputs of invertebrates and
other organisms, and inputs of detritus, wood, and other materials (Gregory et al., 1991; National
Research Council, 2002; Naiman et al., 2005).
Ephemeral tributary stoonflow inputs heavily influence the nutrient and biogeochemical status of the
San Pedro River. Brooks and Lemon (2007) performed synoptic sampling on a 95-km reach of the San
Pedro River to identify the effects of regional hydrology and land use on dissolved carbon and nitrogen
concentrations. They found that, during the summer monsoon season, baseflow increased 5- to 10-fold,
and dissolved organic matter and inorganic nitrogen increased 2- to 10-fold. The fluorescence index of
water samples indicated a large input of terrestrial solutes with the onset of monsoon runoff inflows,
and values of both chloride and oxygen isotope tracers indicated that stream water and alluvial ground
water were well mixed along the entire 95-km reach. Meixner et al. (2007) used chloride tracer samples
and mixing analyses to examine sources of San Pedro River water during six summer floods in 2001
(wetyear) and 2002 (dryyear). Results of mixing models indicated that both a ground water-soil water
end-member and a precipitation end-member (indicative of overland flow) contributed to the floods.
The highest percentage of ground water-soil water in the flood flow (46%) occurred during an early
2001 flood and the lowest during large monsoonal floods of 2002. They noted that ground water
probably made lower contributions than soil water to streamflow, because high river stage during flood
events created hydraulic gradients from the river to alluvial ground water in the riparian area (water
moved from the river to alluvial ground water via bank storage, Figure 2-13B). During the first floods of
each year, nitrate and dissolved organic carbon increased dramatically in the river, whereas dissolved
organic nitrogen did not exhibit increases in 2001 but did in 2002. During floods, nitrate-nitrogen (NOs-
N) concentrations in river water were 0.2-0.5 mg NOs-N Ir1 higher in 2002 than during 2001. This
result was consistent with higher observed nitrate-nitrogen concentrations in soil water of the riparian
zone (alluvial aquifer) in 2002 than in 2001.
In summary, ephemeral tributary streams have strong physical and chemical connections to the San
Pedro River. The river ecosystem, including its abiotic and biotic components, depends on the influences
exerted by the ephemeral tributary streams on the river environment.
B.5.5 Other Southwestern Rivers
B.5.5.1 Physical Connections
Hydrologic behavior and river-system connectivity similar to the San Pedro River have been observed in
other southwestern rivers, increasing confidence that the observations made within the San Pedro are
applicable to other southwestern river systems.
Plummer et al. (2004) found that the Rio Grande in New Mexico has two primary sources of regional
ground water: (1) recharge from mountains and (2) seepage from the Rio Grande and Rio Puerco, and
from Abo and Tijera Arroyos (arroyos are ephemeral streams). Vivoni et al. (2006) observed ground-
water recharge processes in the Rio Puerco, a tributary river to the Rio Grande, and in the Rio Grande
itself. They note that a summer monsoonal rainstorm produced a flood event on the Rio Puerco that, in
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
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turn, generated a pulse of floodwaters along a losing reach of the Rio Grande (Figure 3-2). Forty-nine
percent (49%) of flood volume was lost to the shallow alluvial aquifer of the Rio Grande. Loss of river
water to the alluvial aquifer was observed to decrease with distance down the river reach.
Another important drainage basin type in the western and southwestern United States is endorheic or
closed drainage basins draining to lakes and playas having no outlet to the ocean. The largest of these
western basins is the Great Basin, which is approximately 490,000 km2 (~5% of the area of the United
States) and covers most of Nevada and parts of Oregon, Utah, and California (Atwood, 1994). Closed
basins can contain ephemeral, intermittent, and perennial stream reaches. Although endorheic streams
do not drain into oceans, many support downstream waters and habitat and numerous large perennial
lakes such as Lake Tahoe in California and Nevada and the Great Salt Lake in Utah.
The Pecos River basin in eastern New Mexico and western Texas comprises part of southern Rocky
Mountains in the north and grasslands, irrigated farmlands, deserts, and deep canyons in the southern
lower reaches of the river (Yuan and Miyamoto, 2008). Precipitation occurs as snow in the mountains
and summer monsoonal rainfall in the lower river valley. Based on hydrogen and oxygen isotope
composition of river water, Yuan and Miyamoto (2008) separated the river basin into three subbasins:
(1) the upper basin, (2) the middle basin, and (3) the lower basin. Snowmelt dominates the mountainous
upper basin. The river in the topographically gentle middle basin had mixed sources of water. Thirty-
three percent (33 %) of river water was lost through evaporation occurring in the streams channels and
irrigated fields of the middle basin. Similar to the San Pedro River, up to 85% of streamflowin the lower
basin was estimated to derive from local freshwater sources, mainly monsoonal rainfall. This finding is
consistent with significant contributions of flow from ephemeral tributary streams.
Shaw and Cooper (2008) studied the 14 ephemeral stream reaches in the Little Colorado River Basin in
northeastern Arizona. As derived from the USGS National Hydrography Dataset, this basin contains a
lower percentage of ephemeral and intermittent stream reaches (70%) as compared to the ~93% of
such reaches in the Upper San Pedro. Shaw and Cooper (2008) related watershed characteristics of the
Little Colorado to downstream reaches and the riparian plant communities of those reaches. They found
that, as the watershed area draining to the studied reaches increased, the overall basin channel slope
deceased, which resulted in less erosive capacity due to channel transmission losses and a decrease in
the variability of alluvial ground water in these channels. This resulted in "decreased disturbance
potential and increased moisture availability in the downstream direction," and these reaches had a
greater abundance of obligate riparian vegetation. Shaw and Cooper (2008) went on to develop a stream
classification system that related the functional linkages between contributing upstream watersheds,
stream reaches, and riparian plant ecology. Type I stream reaches have relatively small drainage areas
(less than 10 km2), which have the greatest disturbance potential with in-channel and near-channel
plants resembling those of surrounding upland species. Between 10 and 100 km2, Type II streams
exhibit "more moderate shear stresses and more persistent alluvial groundwater" with riparian
vegetation that is a mixture of upland and riparian species. Having larger areas (greater than 100 km2),
Type III reaches are "controlled mainly by upstream hydro-climatic conditions" with wetland tree and
shrub communities. Shaw and Cooper (2008) concluded that the connection of streamflow and ground-
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water regimes to riparian vegetation in the larger Type III watersheds, draining greater than 100 km2, to
upstream reaches far removed from larger regional floodplain rivers "... were driven by climatic
patterns from distant portions of the upper watershed and were relatively insensitive to local rainfall."
This finding reinforces the fact that stream-reach characteristics are influenced and connected, often
episodically, to distant portions of the contributing watershed.
B.5.5.2 Human Alterations
Anthropogenic uses and activities on arid and semiarid landscapes can have significant effects—both
good and bad—on downstream waters and overall health of watersheds. Human alteration to arid and
semiarid watersheds occurs in many forms and includes livestock grazing, land clearing, mining, timber
harvesting, ground-water withdrawal, streamflow diversion for water supply and irrigation,
channelization, urbanization, agriculture, roads and road construction, off-road vehicle use, camping,
hiking, and vegetation conversion (Levick et al., 2008). Climate change likely will have increasing
influence on streams and their connectivity in the Southwest. Most climate models predict important
changes for the southwestern United States, including increased warming and drying, intensification of
droughts, and increased variability of precipitation (Seager et al., 2007). These changes will result in less
runoff, reduced snowpack, and changes in streamflow patterns. Reduced snowpack will result in shorter
periods of longitudinal stream connectivity in intermittent streams, as snowmelt will occur more rapidly
in a warmer climate.
Streamflow augmentation can occur in human-dominated watersheds in the form of treated municipal
and industrial wastewater effluent discharges. Streams that would dry without these discharges are
effluent-dependent streams, whereas those that receive most, but not all, of their flow from effluent are
effluent-dominated streams (Brooks etal., 2006). Streams draining human-dominated areas also can
acquire baseflow from ground water recharged by over-irrigation and leaky infrastructure (Lerner,
1986; Roach etal., 2008; Townsend-Small etal., 2013).
Riparian areas near mainly perennial streams, but also in many cases intermittent streams, historically
have been attractive for human development, leading to their alteration on a scale similar to that of
wetlands degradation nationally (National Research Council, 2002). This situation is especially true in
arid and semiarid regions because riparian areas typically are indicative of water availability either as
surface water or as shallow ground water. Riparian areas in arid and semiarid regions are also greener
and cooler than most upland areas. Riparian areas are more sensitive to development impacts than
wetter areas, however, because of their limited geographical extent, drier hydrologic characteristics, and
fragile nature (e.g., erodible soils). Historically, riparian habitats represented about 1% of the landscape
in the West, and within the past 100 years, an estimated 95% of this habitat has been lost due to a wide
variety of land-use practices such as river channelization, unmanaged livestock grazing, agricultural
clearing, water impoundments, and urbanization. The following subsections present some of the types
of human-caused impacts on ephemeral and intermittent streams and their associated riparian areas.
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8.5.5.2.1 Land development
Land development includes urban, suburban, and exurban development but is referred to here
collectively as urban development. Before the 2008 recession, the Southwest was one of the fastest
growing regions of the United States, having an increase in population of approximately 1,500% over
the previous 90 years. In contrast, the population of the country as a whole grew by just 225% over that
time. Arizona and Nevada have grown the most, with population increases of 2,880% and 2,840%,
respectively. Typical urban development significantly changes the hydrologic characteristics of a
watershed by covering uplands with impervious surfaces, and removal, channelization, or armoring of
headwater streams (Box 3-1; Kennedy et al., 2013). Alteration of the natural stream network disrupts
natural flow patterns and sediment transport and storage, resulting in downstream flooding and
changes to the clarity and quality of the downstream flows and receiving waters. These effects can
damage downstream water supplies and habitat. The aerial photograph presented in Figure B-ll shows
a network of ephemeral streams that flows through a small community southeast of Tucson, AZ, to
Cienega Creek, a protected perennial stream.
Figure B-ll. Aerial photograph showing ephemeral tributaries to Cienega Creek, a perennial stream,
flowing through the small community of Vail, southeast of Tucson, AZ. Photograph: Lainie
Levick/Aerial flight courtesy of Lighthawk, www.lighthawk.org.
The impact of urbanization increases as the percentage of impermeable surface increases. Various
studies have shown that semiarid stream systems become irreparably impaired once the impervious
surfaces within the watershed exceed about 10% and experience dramatic morphological changes once
those surfaces exceed about 20% (Schueler, 1994; Miltner et al., 2004).
As the amount of impervious surface increases, runoff increases and infiltration decreases (Kennedy et
al., 2013), starting a chain of events that includes flooding, erosion, stream-channel alteration, increases
in human-caused pollutants, and ecological damage. Floods become more severe and more frequent, and
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peak flows and runoff volumes will be many times greater than in natural basins. The greater volume
and intensity of flooding causes increased erosion and sediment transport downstream. To
accommodate the increased flow and sediment load, streams in urbanized areas tend to become deeper
and straighter over time. The resulting bank erosion can destroy established streamside habitat and tree
cover, leading to higher temperatures, sedimentation, and disruption of wildlife corridors.
Storm sewers and lined drainages increase the rate of water delivery to the downstream channel
network. Erosion and sedimentation increases during construction and road building for new urban
areas. Improperly constructed and maintained roads, especially unpaved roads, can alter hillslope
drainage, and change baseflow and precipitation-runoff relationships, causing erosion and
sedimentation in streams (USDA, 2002). The primary geomorphic consequence of these hydrologic
changes is the erosional entrenchment of nearby channels and associated transportation of the
excavated sediment downstream, causing a significant increase in sediment load. Sediment is of
particular concern in arid and semiarid regions because many other pollutants tend to adhere to eroded
soil particles. Additional pollutants from urban runoff can include pathogens, nutrients, toxic
contaminants, sediment, and debris. Consequently, urban areas require stormwater management plans
both during and after construction to control runoff and offsite pollution.
Streams are channelized in urbanizing areas to protect private property and control streambank
erosion. Channelization typically straightens and steepens the stream, however, resulting in increased
flow velocity and sediment movement. These changes transfer flooding and bank erosion downstream
of the protected area. In the channelized reaches, the greatly reduced out-of-bank flow disrupts water,
sediment, organic matter, and nutrient enrichment of the flood plain (National Research Council, 2002).
In addition, removal of vegetation as part of the channelization process degrades wildlife habitat.
Habitat fragmentation is a common consequence of urbanization (Hilty etal., 2006). New developments
can alter large areas of land, removing natural drainage systems and wildlife habitat, and replacing them
with houses and roads. Altering, bisecting, or channelizing streams effectively can eliminate the main
biological functions of the stream channel by disrupting vegetation communities and hydrologic
function. Habitat fragmentation reduces wildlife diversity and abundance and might cause sensitive
species to disappear (England and Laudenslayer, 1995).
B.5.5.2.2 Land use
In addition to urbanization, agriculture (livestock and crops) and mining, including sand and gravel
operations, are major land uses in the desert Southwest. Livestock grazing is one of the more common
uses of rural land in the Southwest. Late 1800s estimates of cattle numbers in Arizona and New Mexico
exceeded 1.5 million and 2 million, respectively. During this period, the region experienced both
significant droughts and floods. During drought, the resulting desiccation of the uplands drove cattle to
the riparian areas, which were heavily damaged as a result. When the rains returned to the denuded
landscape, erosive processes were greatly enhanced. The overgrazing that occurred during this time is
one of the factors attributed to a relatively widespread period of channel downcutting, forming deep
arroyos and lowering ground-water levels (Schumm and Hadley, 1957; Hastings, 1959; Graf, 1988).
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In modern grazing-land management, livestock are provided with watering sources away from streams
when possible, but frequently they must depend on the streams for water. Livestock management
efforts attempt to avoid overuse of an area, but because water is scarce in arid environments, cattle and
wildlife tend to linger near water sources. Where not properly managed, cattle can remain too long in a
riparian area and trample streambanks, eat the riparian vegetation to the ground, contaminate the
water with wastes, and compact the soil (Levick et al., 2008). Several literature sources have stressed
that the cumulative impacts of unmanaged livestock in southwestern riparian ecosystems for the past
several hundred years probably have been the single most important factor in riparian ecosystem
degradation (Wagner, 1978; Ohmart, 1995).
Mining is another activity that historically has played a large role in the economy and land use in the
Southwest. Some of the largest copper and gold mines in the world are found in this region, and some
cover many thousands of hectares. Mining can cause major impacts on riparian areas along tributaries
and downstream waters by altering the local hydrology. Mining not only dewaters the area, it removes
vegetation and soil and changes the topography, severely affecting the watershed. Instream and
floodplain gravel mining can alter channel dimensions, increase sediment yield, and increase fine
sediment loading and deposition that can reduce infiltration into ephemeral channels (Bull and Scott,
1974).
Cultivated agriculture has had a long history in the southwestern deserts, and areas such as the Central
Valley in California provide much of the country's food supply. Most crops, however, must be irrigated
due to the low annual rainfall. Impacts to local hydrology from agricultural activities include (Levick et
al.,2008):
• Increased salinity caused by clearing of native vegetation that raises the ground-water
reservoir;
• Reduced flows from ground-water pumping or stream diversions for irrigation;
• Increased nutrients and turbidity from the use of fertilizers that run off into the streams across
the land surface or through the soil, causing excessive algal growth; and
• Fish, aquatic invertebrate, and bird kills from pesticides that run off into the streams or leach
into the ground water.
Due to the abundant solar resources in the arid and semiarid Southwest, numerous, large-scale solar
energy projects are envisioned or already under development. O'Connor et al. (2014) note that
development of solar energy zones will significantly affect ephemeral channel systems; the authors have
developed a scoring system to conduct ephemeral stream assessments using publicly available
geospatial data and high-resolution aerial imagery.
8.5.5.2.3 Water resources impacts
The South west has experienced rapid growth over the past several decades. This growth can be
sustained only with reliable water supplies. Lack of surface-water flows has placed increased reliance on
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ground water for human and agricultural uses. The percentage of population dependent on ground
water for domestic water needs in New Mexico, Arizona, and California is 90%, 60%, and 45%
respectively (Anderson and Woosley, 2005). When ground-water pumping is sufficiently large or
prolonged, it can result in lower water-table levels in regional and alluvial aquifers. If these aquifers are
a primary source of water for sustaining surface-water flow in perennial or intermittent streams and if
the drop in aquifer water levels is large enough, the pumping can effectively dewater these stream
reaches, severing longitudinal and vertical connectivity (Winter et al., 1998; Scanlon et al., 2012). The
perennial and intermittent streams effectively become ephemeral streams, and the habitat supported by
reliable surface flow or shallow ground water is lost (Stromberg et al., 1996).
The impact of aquifer water-level declines is illustrated in Figure B-12 with repeat photography of the
Santa Cruz River south of Tucson from 1942 and 1989 showing changes in riparian vegetation. Tucson's
population in 1940 was roughly 36,000 and increased to approximately 405,000 by 1990. Until the
Central Arizona Project brought Colorado River water to Tucson in the early 1990s, Tucson's domestic
water supply was solely provided by ground water. As ground-water pumping increased to supply the
growing population, the aquifer water level dropped by more than 25 meters and the riparian habitat
was completely altered, as all phreatophytic vegetation died out. The growing population of Tucson also
resulted in proportional increases in discharge of treated effluent. Portions of the Santa Cruz River
downstream of the reach photographed in Figure B-12 near treatment plant outfalls are now effluent-
dependent perennial stream reaches. Depending on the level of treatment, effluent can have various
effects on the stream ecosystem (Brooks et al., 2006). Without careful water management and reuse
(Bischel et al., 2013), the benefits of baseflow augmentation can be overshadowed by potential risks,
such as increased contaminant and pathogen exposures (Jackson and Pringle, 2010).
Dams and retention or detention basins frequently are used to store water or as flood-control devices in
the Southwest. They disrupt natural surface flow and sediment transport, interfere with natural
geomorphic processes, alter water temperatures, and fragment the natural stream systems both
upstream and downstream of the structure (Williams and Wolman, 1984). Upstream locations can
experience flooding, whereas downstream locations can be dewatered and become starved of sediment
(Sections 2.4.4 and 3.3.2).
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Figure B-12. Change in riparian vegetation along the Santa Cruz River, Tucson, AZ, as the result of
water-level declines in the regional aquifer. Photographs of the Santa Cruz River looking south from
Tucson, AZ, provided by Robert H. Webb, U.S. Geological Survey Anderson and Woosley (2005).
1942
i i
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B.5.5.3 Biological Connections
Much of the material in this section, as in Section B.5.5.2, is derived from the material presented in
Levick et al. (2008). As noted in Chapter 3, ephemeral and intermittent streams perform many of the
same functions in a watershed as perennial streams. In particular, in arid and semiarid regions, riparian
areas, including those near ephemeral and intermittent streams, support the vast majority of wildlife
species, are the predominant sites of woody vegetation including trees, and surround what are often the
only available surface-water sources, even if they are available only for limited periods. Desert washes
are easily recognizable by their dense corridors of vegetation that strongly contrast with the more
sparsely vegetated uplands (Figure B-13). In contrast to the nearby uplands, these stream corridors and
their associated vegetation communities provide structural elements of food, cover, nesting, and
breeding habitat, and movement/migration corridors for organisms. These corridor vegetation
communities moderate soil and air temperatures, stabilize channel banks, provide seed banking and
trap silt and fine sediment that favor the establishment of diverse floral and faunal species, and dissipate
stream energy (Levick et al., 2008). The resulting microclimates in and around ephemeral and
intermittent stream vegetation corridors are used extensively by fauna. In arid climates, such conditions
often benefit less mobile species that cannot avoid the harsh desert environment by moving to
microclimates that are more favorable. These stream corridors provide primary habitat, predator
protection, breeding and nesting sites, shade, travel corridors, migration stopover sites, and food
sources.
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Figure B-13. Aerial photograph showing dense corridor of vegetation lining ephemeral washes in
southeastern Arizona. Image accessed from Google Earth from May 2005 imagery date.
UP-
m
Both passive and active biological connections exist in the intermittent and ephemeral streams network.
Passive connections involve the transport of organisms and organic matter driven by water flow; these
connections thus depend on hydrologic connectivity. Active connections do not depend on flowing
water; instead, dispersal of organisms and organic matter occurs throughout the stream network
through walking, flying, or hitchhiking on mobile organisms. All these organism-mediated connections
form the basis of biological connectivity between headwater streams and downstream waters.
Movement can be both longitudinal along the stream network and lateral, and can occur over the life
cycles of numerous organisms (Schlosser, 1991; Fausch etal., 2002).
Meyer et al. (2007) noted the importance of headwater streams, including ephemeral and intermittent
streams, as vital parts of the biological integrity of U.S. waterways. Ephemeral and intermittent stream
channels are bordered by a zone of continuous or near-continuous vegetation, and thus they provide
important wildlife movement corridors as they afford both cover and food. Summer monsoons in the
Southwest coincide with periods when herptofauna such as snakes and amphibians are most active; the
episodic flows provide a generally continuous aquatic corridor for their dispersal. The translocation and
dispersal of species enables genetic interchange between subpopulations that are often isolated for most
of the year. In addition, recolonization of sites can occur when subpopulations are lost due to drought or
disturbance. Degradation of these habitats and loss of their connections to larger streams can have
negative consequences for the diversity of downstream and riparian ecosystems and for the biological
integrity of the entire river network. Nearly 81% of all streams are ephemeral or intermittent in the six
Southwestern states (USGS, 2006). From a strictly numerical viewpoint, therefore, degradation of these
ephemeral streams diminishes ecosystem functions in most southwestern watersheds.
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8.5.5.3.1 Physical features important to biological connectivity and integrity
Ephemeral and intermittent riverbanks in the arid Southwest provide shelter for numerous species of
wildlife, including reptiles, amphibians, birds, mammals, and invertebrates. These shelters typically are
independent of whether the streams contain water year-round. Shelters are created through the action
of water, wind, and gravity. Ephemeral dry-wash embankments notoriously are full of small caves and
crevices critical in the life of desert animals such as the desert tortoise [Gopherus agassizii) (Van
Devender, 2002). The alluvium in ephemeral and intermittent streams is often looser than the soils or
colluvium of surrounding uplands. These conditions provide enhanced habitat by specialized sand-
burrowing species of wildlife. High-value shelters also are created when woody debris is swept in from
the watershed and collects in the floodplain and stream channel. In cases of deep ephemeral stream
incision, cooler canyon-type environments might be created in which moisture loss is retarded.
8.5.5.3.2 Vegetation habitat features important to biological connectivity and integrity
Large ephemeral washes with shallow ground-water zones often are colonized with a variety of
phreatophytic trees, such as Fremont cottonwood (Populusfremontii), Arizona sycamore (Platanus
wrightii), and Arizona ash (Fraxinus velutind). These washes also include distinctive shrubs, such as
willow (Salix spp.), seepwillow (Baccharis spp.), burrobrush (Ambrosia monogyrd), and saltcedar
(Tamarix ramosissimd), and dense grass stands of sacaton (Sporobolus spp.). Those washes that lack a
shallow ground-water system or water augmentation by effluent discharge nonetheless give rise to a
distinctive vegetative habitat from the surrounding uplands. These environments often are referred to
as xeroriparian habitat.
The floral species in these habitats is moderated by the frequency and magnitude of runoff events.
Common tree species in xeroriparian habitat include subtropical legumes such as mesquite (Prosopis
spp.), catclaw acacia [Acacia greggii), ironwood (Olneya tesotd), and blue palo verde (Cercidium
floriduni). Mesquite has been identified as the key provider of food for numerous migrating birds (Van
Riper and Cole, 2004). Netleaf hackberry (Celtis reticulatatd) and Arizona sycamore (Platanus wrightii)
have been identified as providing exceptional cover for nesting birds on intermittent streams (Powell
and Steidl, 2002).
8.5.5.3.3 Hydrologic habitat features important to biological connectivity and integrity
Stanley et al. (1997) provide an excellent overview of the expansion and contraction of flowing waters
within southwestern streams in response to variable precipitation events. This phenomenon commonly
results in reaches of streams or rivers that have flow or residual pools with water surrounded by
reaches without water. This phenomenon is common in dryland rivers across the globe (Arthington et
al., 2005; Bunn et al., 2006). The isolated pools often serve as refuges for fish to survive in intermittent
streams during dry periods (Labbe and Fausch, 2000).
Episodic stream flow might be the most visually prominent hydrologic aspect of a stream but is seldom
the only hydrologic habitat feature of biological significance. An arid stream wash with a shallow
ground-water system also might have moist banks fed by capillary flow that provide sites for turtle or
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insect reproduction. Distinct invertebrate fauna can inhabit the hyporheic (subsurface) zone of flow
beneath a dry streambed. Episodic flooding, human excavations, and channel scour can produce in-
channel or off-channel pools where amphibians breed. Within-channel or floodplain springs can provide
distinct chemical compositions or thermal refuges from the main ephemeral or intermittent stream.
The natural episodic and intermittent flow regime in the arid South west is a competitive factor of native
species over exotics adapted to lake and pond conditions (Minckley and Meffre, 1987; Poff et al., 1997).
Louwand Seely (1982) and Williams (2005) concluded most desert species have developed adaptations
to the water-limited conditions of these regions that enable them to survive under harsh environmental
conditions. Fauna using ephemeral or intermittent waters include fish, mammals, amphibians, reptiles,
birds, and invertebrates. The variability of climate and flow regime, which influences species abundance
and diversity, however, makes evaluation difficult unless surveys are conducted over years in different
community types (Boulton and Lake, 1992).
8.5.5.3.4 Fish and aquatic insects
As discussed in the previous section, the interplay between stoonflow from ephemeral tributary
streams, water from alluvial aquifers, and water from regional ground water control the distribution
and timing of flowing water in southwestern rivers. Native fish species of southwestern streams and
rivers are adapted to these dynamic environments (John, 1964; Meffe, 1984). Rinne and Miller (2006)
compared fish assemblage data in river networks for two southwestern rivers, the Gila River (New
Mexico and Arizona) and the Verde River (Arizona) over 7 to 12 years. They included river hydrology
and geomorphology data in their analysis and found that variable streamflows and higher flow volumes
favor native fish species over nonnatives. They also noted that the presence of unconstrained alluvial
valley river reaches with shallow pools favored native fish. Furthermore, when humans alter the
hydrologic dynamics of ephemeral and intermittent tributaries such that flows connecting them to the
river network are more frequent or more consistent, nonnative fish can invade (Turner and List, 2007).
Recent nonnative invasion and a corresponding decline in native fish species diversity were observed in
the lower reaches of Aravaipa Creek, a tributary of the San Pedro River, which historically was only
rarely connected to the mainstem (Eby et al., 2003).
Lytle et al. (2008) found a similar adaptation strategy in populations of an aquatic insect (Abedus
herberti) occupying sites along a natural gradient of disturbance predictability. In their study,
predictability was defined as the ability of a signal or cue (rainfall) to cause a disturbance. In this case,
the disturbance was a flash flood. Using signal detection theory, they found that for 13 of 15 insect
populations, the observed insect response times "were an optimal compromise between the competing
risks of abandoning versus remaining in the stream, mediated by the rainfall-flood correlation of the
local environment." They concluded that these aquatic insect populations could evolve in their
responses to changes in the flow disturbance regime, providing evidence that these aquatic populations
can adapt to "among-stream differences in flow regime."
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B.5.6 Southwestern Intermittent and Ephemeral Streams: Synthesis and
Implications
Rivers of the arid and semiarid Southwest are products of a highly variable and dynamic environment.
Ephemeral and intermittent streams and their tributaries in the American South west provide a wide
range of functions that are critical to the health and stability of arid and semiarid watersheds and
ecosystems. Most importantly, they provide hydrologic connectivity within a basin, linking ephemeral,
intermittent, and perennial stream segments. This linkage and the corridor of connectivity facilitates the
movement of water, sediment, nutrients, debris, fish, wildlife, and plant propagules throughout the
watershed. The relatively more vegetated streams corridors connected to downstream perennial
reaches provide wildlife habitat and more humid environment than do the surrounding uplands. During
ephemeral and intermittent streamflow, energy dissipates as part of natural fluvial adjustment, and
sediment, organic matter, and debris are transported. The variability of the hydrologic regime in these
streams is the key determinant of spatial and temporal distribution of plant community structure and
the types of plants and wildlife present. Some of the major ways in which ephemeral streams are
connected with and influence rivers are as follows:
• Flows from ephemeral streams are a major driver of the dynamic hydrology of southwestern
rivers. Ephemeral tributary streamflows are especially important drivers of downstream floods
during monsoon seasons.
• Fishes and invertebrates native to mainstem rivers are adapted to the variable flow regimes that
ephemeral tributary streams strongly influence. Ephemeral flows prevent or mitigate invasion
by introduced species.
• Ephemeral tributary streams supply water to mainstem river alluvial aquifers; these alluvial
aquifers help sustain river baseflows.
• Ephemeral streams export sediment to rivers during major hydrologic events; the sediment
contributes to materials that comprise alluvial aquifers and shape the fluvial geomorphology of
rivers.
• Ephemeral tributaries export nutrients to mainstream rivers during hydrologic flow events;
nutrients occur in many forms and contribute to river productivity.
• Ephemeral and intermittent streams and their associated vegetation communities provide
structural elements of food, cover, nesting and breeding habitat, and movement/migration
corridors for organisms.
• Water, sediment, and nutrients exported to the river from ephemeral tributaries support
riparian communities of mainstem rivers; the riparian communities profoundly influence river
attributes through shading and allochthonous inputs of organic matter, detritus, wood, and
invertebrates to the river.
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• Regional ground-water aquifers are in part recharged through infiltration of water to the
streambed of ephemeral stream channels during wet years; the regional aquifer supplies a
varying but critical portion of baseflow for perennial river reaches.
B.6 Case Study: Vernal Pools
B.6.1 Abstract
Vernal pools are shallow, seasonal wetlands that accumulate water during colder, wetter months and
gradually dry down during warmer, dryer months. Despite differences in geology, climate, and biological
communities, some common findings about the hydrologic connectivity of vernal pools in different
regions include evidence for temporary or permanent outlets, frequent filling and spilling of higher
pools into lower elevation swales and stream channels, and conditions supporting subsurface flows
through pools without perched aquifers to nearby streams. Insects and amphibians that can live in
streams or permanent pools opportunistically use glaciated vernal pools as alternative breeding habitat,
refuge from predators or environmental stressors, hunting or foraging habitat, or stepping-stone
corridors for dispersal and migration. Nonglaciated vernal pools in western states are reservoirs of
biodiversity and can be connected genetically to other locations and aquatic habitats through wind- and
animal-mediated dispersal.
B.6.2 Introduction
The term "vernal pool" is broadly used to describe shallow, fishless pools situated on bedrock or low-
permeability soils that lack continuous surface-water connection to permanent water bodies but have a
seasonal period of inundation on which aquatic species depend for completion of their life cycles
(Zedler, 2003). This case study reviews evidence for physical and biological relationships between
vernal pools and downstream waters in the western United States (western vernal pools) and glaciated
areas of northeastern and midwestern states (northern vernal pools), where vernal pools are
particularly abundant (Zedler, 2003).
B.6.2.1 Geography and Geology
8.6.2.1.1 Western vernal pools
Zedler (1987) used the term vernal pool to describe basin/swale systems in California's Mediterranean
climate that flood in winter, host diverse communities of aquatic plants and animals in early spring,
transition to terrestrial ecosystems in late spring, and desiccate during hot, dry summer months.
Western vernal pools are seasonal wetlands associated with topographic depressions; soils with poor
drainage; mild, wet winters; and hot, dry summers in western North America from southeastern Oregon
to northern Baja California, Mexico (Bauder and McMillan, 1998). Locally, wetlands that fit this
definition might be known by other names, such as the upland playas in Oregon (Clausnitzer and
Huddleston, 2002).
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Historically, vernal pools covered 518 km2, or 5-6% of the total land surface in southern California and
northern Baja, but losses in that area have been substantial (Bauder and McMillan, 1998). Pools occur on
impermeable or slowly permeable soils or bedrock (Smith and Verrill, 1998) that limit percolation and
thus produce surficial aquifers that perch above regional ground-water aquifers. Pool-forming soil
layers in this region include clay-rich soils, silica-cemented hardpans (duripans), volcanic mudflows, or
bedrock (Weitkamp etal., 1996; Hobson and Dahlgren, 1998; Smith and Verrill, 1998; Rains etal., 2006).
Because their hydrology and ecology are so tightly coupled with the local and regional geologic
processes that formed them, western vernal pools typically occur within "vernal pool landscapes"
(Smith and Verrill, 1998), or complexes of pools in which swales connect pools to each other and to
seasonal streams (Weitkamp etal., 1996; Rains etal., 2008).
8.6.2.1.2 Northern vernal pools
The geologic formations underlying northern vernal pool landscapes were formed by the movement of
glaciers across the northeastern and north-central states approximately 12,000 years ago. Retreating
glaciers scoured basins in rock ledges and mountaintops, or left behind large pieces of ice that later
collapsed to form topographic depressions containing deposits of gravel, sand, or mud (Colburn, 2004).
Although not all vernal pools in these areas were formed by glaciers, the soils, geology, and evolutionary
history of plants and animals in northern vernal pools have been profoundly affected by glacial events.
Like western vernal pools, northern vernal pools are significantly grouped or clustered (Brooks, 2005).
Grant (2005) found that pools in Massachusetts are more likely to occur in more porous substrates
(alluvial, fine grained, or sand/gravel soils) than glacial till or impermeable bedrock, increasing their
hydrologic connection to shallow ground water.
Unlike western vernal pools, which typically occur in open grasslands, most northern vernal pools are
detrital wetlands fully contained within forest ecosystems that depend on the pulse of organic matter
from leaf fall that coincides with initial filling of temporary pools in these regions.
B.6.2.2 Temporal Dynamics
Zedler (1987) identified four distinct ecosystem phases in the annual hydrologic cycle of western vernal
pools, which we have generalized here (with additional citations) to describe the temporal dynamics of
northern vernal pools as well:
• Wetting or newly flooded phase: Rainwater, snow, runoff, or snowmelt infiltrate upper layers of
permeable soil and, when topsoils are saturated, collect in pool basins formed by impervious
rock, clay, or till layers (aquitards or aquicludes; Rains et al., 2008). In early spring, perennial
plants sprout and stored seeds germinate in wet soils. Aquatic invertebrate communities
develop from resting eggs and seed banks (Colburn, 2004).
• Aquatic phase: Soils are saturated and pools hold standing water, in many locations filled to
capacity. In some western vernal pools, surface and subsurface flows from upland pools through
swales feed downgradient pools, connecting pools at a site and extending the aquatic phase of
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the pool complex (Weitkamp et al., 1996; Hanes and Stromberg, 1998). Pools are colonized by
dispersing insects and breeding amphibians.
• Terrestrial phase: Evapotranspiration rates increase and pool water recedes, although soils
remain saturated. In western pools, aquatic plants flower and seed. Aquatic animals disperse or
become dormant Terrestrial plant communities persist.
• Dry phase: Pools and soils dry to moisture levels similar to uplands, and many plants senesce or
die. Summer rains produce no new ponding or plant growth.
In the western United States, vernal complexes saturate and begin to pool during winter rains, reach
maximum depth by early spring, and lose all standing water by late spring (Zedler, 1987). The timing of
filling and drying of northern vernal pools varies, depending on pool type. Colburn (2004) proposed five
hydrologic classes for northern vernal pools, based on time of filling and average duration of flooding:
(1) short-cycle, spring-filling pools that stay wet for 3-4 months; (2) long-cycle, spring-filling pools that
stay wet for 5-8 months; (3) short-cycle, fall-filling pools that stay wet for 7-9 months; (4) long-cycle,
fall-filling pools that stay wet for 9-11 months; and (5) semipermanent pools that stay wet for 36-120
months. Many northern vernal pools do not dry down completely, but retain areas of saturated sediment
or standing water in part of the basin. Such pools are considered "incompletely dry," to differentiate
them from pools that are "continuously flooded" or "dry."
B.6.2.3 Ecology
Vernal pool ecosystems support large breeding populations of amphibians, aquatic invertebrates, and
aquatic or semiaquatic plants, including many rare or endemic taxa (King et al., 1996; Zedler, 2003;
Colburn, 2004; Calhoun and DeMaynadier, 2007). The annual cycle of basin flooding and drying plays an
important role in structuring biological communities in vernal pools. The wet phase prevents
establishment of upland plant species in pool basins, while the dry phase limits colonization by aquatic
and semiaquatic plant and animal species that occur in permanent wetlands, ponds, or streams (Keeley
and Zedler, 1998; Bauder, 2000). Despite their cyclical nature, vernal pool habitats are species rich and
highly productive, in part because they provide relatively predator-free breeding habitat for
invertebrates and amphibians (Keeley and Zedler, 1998; Calhoun et al., 2003). Many resident species are
locally adapted to the timing and duration of inundation, soil properties, and spatial distribution of
vernal pools in a specific geographic subregion. Other species that are widespread across regions and
aquatic habitat types (including streams or lakes) use inundated pools periodically for refuge,
reproduction, or feeding (King et al., 1996; Williams, 1996; Colburn, 2004).
B.6.3 Evidence
B.6.3.1 Physical Connections
Vernal pools are primarily precipitation fed and typically lack permanent inflows from or outflows to
streams or other water bodies. They can be connected temporarily, however, to permanent waters by
surface or shallow subsurface flow (flow through) or ground-water exchange (recharge; Weitkamp et al.,
1996; Brooks, 2005; Rains et al., 2008). Hydrologic connectivity is typically limited to flow through in
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vernal pools formed by perching layers; ground-water exchange can occur in vernal pool systems
without perching layers (Brooks, 2005).
B.6.3.1.1 Western vernal pools
Rains et al. (2006; 2008) examined the hydrology and biogeochemistry of two vernal pool complexes in
the northern end of California's Central Valley (Smith and Verrill, 1998). The 2006 study evaluates water
balance and the relative importance of direct precipitation, evaporation, surface flow, and shallow
subsurface flow in a hardpan vernal pool complex (Rains et al., 2006). The 2008 study contrasts the role
of geology and soil type—specifically, clay-rich versus hardpan soils—in controlling vernal pool
hydroperiod, hydrodynamics, and water chemistry (Rains etal., 2008). Clay-rich and hardpan
complexes are common vernal pool types in California's Central Valley (Smith and Verrill, 1998). In both
studies, study sites were pool complexes located in the upper portion of the watersheds. Within each
complex, upland (feeder) pools were connected to lower (collector) pools by ephemeral swales, and the
lowest pool was connected by swale to a seasonal stream.
Results showed that high and low pools were connected via surface flows 10-60% of the time; surface
water flowed through swales connecting low-elevation pools to streams during 60% of the inundation
period (Table B-l). Underlying geology and soil type influenced ponding rates and inundation periods:
In water year 2003, pools in clay-rich soils accumulated water at the onset of rainfall and held water
longer than pools in hardpan soils, which have higher soil infiltration rates (Table B-l; Rains et al.,
2008). Horizontal subsurface flows reduced the number and volume of higher elevation surface flows
into hardpan pools, relative to the clay-rich pools. Most water discharging from the swale to the seasonal
stream at the hardpan site was perched ground water that had flowed around, rather than through, the
pool basins. In both soil types, however, vernal pool basins, swales, and seasonal streams were shown to
be part of a single surface-water and shallow ground-water system connected to the river network when
precipitation exceeds storage capacity of the system (Rains etal., 2006; Rains etal., 2008). Pyke (2004)
reported that a complex of 38 vernal pools north of Sacramento was filled to capacity in 10 of 11 years
from November 1999 to June 2001. A direct precipitation-evaporation model for another hardpan
complex near this Sacramento site showed that direct precipitation could fill pools beyond capacity in
most years (Hanes and Stromberg, 1998). Pools located at the lower end of a complex (and thus more
likely to be directly connected to streams) can receive surface water through stepping-stone spillage in
addition to direct input from precipitation; thus, they can remain wetted longer than upper pools. For
example, Bauder (2005) found that "collector" pools with no outlet held water longer than headwater
pools with no inlet. Collectively, these findings suggest that filling and overflow of vernal pools are not
rare phenomena. Filling and spilling also can occur in other vernal pool types because all vernal pools
are underlain by aquitards (Rains et al., 2008).
B.6.3.1.2 Northern vernal pools
Northern vernal pools include both perched and ground water-connected aquifers (Brooks, 2004; Boone
etal., 2006). As in western vernal pools, rainfall or snowmeltin excess of pool capacity is lost to surface
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Table B-l. California vernal pool inundation and hydrologic connectivity. Summarized from Rains et al.
(2008)
Soil; hydrology
Fine-grained, clay-
rich soils; perched
surface water
Coarse-grained,
hard pan soils;
perched surface
water and ground
water
Inundation period
(days/water year3)
200-205
150-154
Flow-through paths
(pool-pool and pool-
stream)
surface only
surface and
horizontal
subsurface
Surface flows
between high- and
low-elevation poolsb
120 (60%)
15 (10%)
Surface flows
between lowest
elevation pool and
stream network13
120-123 (60%)
90-92 (60%)
aOctober 1 2002-September 30, 2003.
b Units = days/water year3, % of inundation period.
runoff or subsurface flows into shallow, nearby ground water (Brooks, 2005). Studies of surface and
subsurface inflows and outflows were not found in the literature. Brooks (2004) reports that
precipitation and potential evapotranspiration alone could not account for large observed water losses
in four vernal pools he studied for 10 years. These losses could have been due to inaccurate estimates of
precipitation or evapotranspiration (both of which were significantly related to water depth) or to
surface overflow and soil infiltration, which were not measured. In a separate study, Boone et al. (2006)
used a classic water-budget model to predict vernal pool hydroperiods in Minnesota and found that,
although precipitation and evapotranspiration were good predictors of pool inundation in most cases,
errors in model estimates for a few pools suggested that surface outflows or infiltration might have been
occurring at some sites.
Individually small, temporary storage of heavy rainfall and snowmelt in vernal pool systems (pools plus
soils) can attenuate flooding, provide a reservoir for nearby vegetation during the spring growth period,
and increase nutrient availability (Hobson and Dahlgren, 1998).
B.6.3.2 Biological Connections
Dispersal of vernal pool organisms can be active or passive and occurs at multiple scales: local scale
(among nearby pools), neighborhood scale (among pools in a geographic cluster or complex), or
regional (outside of the complex, to other ecosystem types; Compton et al., 2007). Examples of active
regional dispersal include insect flight or juvenile dispersal by amphibians. Passive transport is of
particular interest for regional-scale dispersal, as it enables plants and low-vagility animals such as
microcrustaceans to move long distances. Examples of passive transport to and from unidirectional
wetlands and pools include water-mediated dispersal of larvae (Hulsmans et al., 2007); transport of
diapausing (dormant) eggs by waterbirds (Figuerola et al., 2005; Frisch et al., 2007) or flying insects
(Van De Meutter et al., 2007); and wind-mediated dispersal of dormant eggs, larvae, and adult
zooplankton from dry rock pools (Vanschoenwinkel et al., 2009).
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Western vernal pools are highly productive ecosystems that have evolved in what Zedler (2003)
describes as a "balance between isolation and connectedness." Pacific vernal pool landscapes are tightly
coupled with variable climate, soils, and geologic formations in the western United States, producing
diverse habitats for organisms with different life-history strategies (Bauder and McMillan, 1998).
Seasonal wetlands in this region might have served as evolutionary refuges since Mesozoic times (King
et al., 1996). As a result, present-day vernal pool communities have a large proportion of passively
dispersing, endemic (i.e., restricted to small geographic area) species in genera that are widely
distributed across continents and aquatic systems (King et al., 1996; Keeley and Zedler, 1998; Zedler,
2003). This apparent paradox is explained by the fact that individuals transported passively over long
distances have colonized, and through time have become locally adapted to, different vernal pool
landscapes, creating new endemic species from the rootstock of ancient lineages. As a result, Pacific
vernal pools are now rich reservoirs of genetic and species diversity connected to other locations and
aquatic habitats through continuing dispersal. The existence and connectivity of such reserves are
especially important at a time when changing climatic conditions are likely to increase intermittency of
stream flows and decrease duration of wetland inundations in other areas.
Western vernal pools also support generalist invertebrate communities, including crustaceans and
insects that are widely distributed in permanent wetlands, ponds, lakes, and streams (Zedler, 1987;
2003). Invertebrates and zooplankton can be flushed from vernal pools into streams or other water
bodies during periods of overflow, carried by animal vectors (including humans), or dispersed by wind.
Wind-mediated dispersal can be of particular importance in seasonal wetlands: during the dry phase,
dry soils containing large numbers of transportable seeds, resting eggs, cysts, diapausing larvae, and
adults are picked up and blown away (Vanschoenwinkel et al., 2009). The maximum distance such
propagules can travel is not known, but, from currently available literature, pool-pool or pool-stream
transport is clearly possible, and the potential for long-distance transport also exists.
Food webs in northern vernal pools include highly fecund amphibians and insects that convert detrital
organic matter inputs into biomass that subsidizes terrestrial and aquatic ecosystems in other parts of
the watershed (Semlitsch and Bodie, 1998; Brooks, 2000; Gibbons etal., 2006). Northern vernal pools
can provide alternative breeding habitat, refuge from predators or environmental stressors, hunting or
foraging habitat, or stepping-stone corridors for dispersal and migration. For example, Gahl et al. (2009)
reports that bullfrog (Rana catesbeiand) densities per unit wetland perimeter were greater in two small
seasonal pools than in a larger, permanent breeding wetland. Regular use of seasonal pools by bullfrogs
throughout this study offers compelling evidence for the role of seasonal pools as a component of their
nonbreeding habitat. Spotted turtles (Clemmysguttatd) used seasonal pools for foraging, basking, and
mating at two sites in Massachusetts (Milam and Melvin, 2001). Many insects and amphibians found in
streams, lakes, or riparian/floodplain wetlands are facultative users of vernal pool habitats (Table 4-2).
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B.6.4 Vernal Pools: Synthesis and Implications
The key findings from this case study are as follows:
• In the aquatic phase, some western vernal pools are filled to capacity in most years, creating
conditions under which water flows from pools into swales and stream channels.
• Documented evidence of surface flows connecting western vernal pool complexes to the river
network via swales and seasonal streams is available in the literature.
• Indirect evidence indicates that surface and subsurface flows connect northern pools without
perched aquifers to shallow ground water and thus to nearby streams.
• Many insects and amphibians that can live in streams or pools that are more permanent
opportunistically use northern vernal pools as alternative breeding habitat, refuge from
predators or environmental stressors, hunting or foraging habitat, or stepping-stone corridors
for dispersal and migration.
• Nonglaciated vernal pools in western states have achieved a long-term "balance between
isolation and connectedness" and have functioned as refuges for plant and animal diversity since
the Mesozoic era. They are current reservoirs of biodiversity connected genetically to other
locations and aquatic habitats through continuing dispersal.
Direct evidence supports the existence of seasonal hydrologic connections and indirect evidence
supports the movement of organisms between western vernal pool complexes and streams. Indirect
evidence supports the existence of hydrologic and biological connections between northern vernal pools
and river networks, with potential for storing water during the wet season, and providing alternative
breeding habitat or food resources for stream organisms.
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References
Alexander, R. B., R. A. Smith, and G. E. Schwarz. 2000. Effect of stream channel size on the delivery of
nitrogen to the Gulf of Mexico. Nature 403:758-761.
Alexander, R. B., R. A. Smith, G. E. Schwarz, E. W. Boyer, J. V. Nolan, and J. W. Brakebill. 2008. Differences
in phosphorus and nitrogen delivery to the Gulf of Mexico from the Mississippi River basin.
Environmental Science & Technology 42:822-830.
Amezaga, J. M., L. Santamaria, and A. J. Green. 2002. Biotic wetland connectivity—supporting a new
approach for wetland policy. Acta Oecologica 23:213-222.
Amoros, C., and G. Bornette. 2002. Connectivity and biocomplexity in waterbodies of riverine
floodplains. Freshwater Biology 47:761-776.
Anderson, M. T., and L. H. Woosley, Jr. 2005. Water availability for the Western United States—key
scientific challenges. U.S. Geological Survey Circular 1261, U.S. Department of the Interior, U.S.
Geological Survey, Reston, VA.
Armour, C. L., D. A. Duff, and W. Elmore. 1991. The effects of livestock grazing on riparian and stream
ecosystems. Fisheries 16:7-11.
Aronson, M. F. J., and S. Galatowitsch. 2008. Long-term vegetation development of restored prairie
pothole wetlands. Wetlands 28:883-895.
Arthington, A. H., S. R. Balcombe, G. A. Wilson, M. C. Thorns, and J. Marshall. 2005. Spatial and temporal
variation in fish-assemblage structure in isolated waterholes during the 2001 dry season of an arid-
zone floodplain river, Cooper Creek, Australia. Marine and Freshwater Research 56:25-35.
Ash worth, A. C. 1999. Climate change in North Dakota since the last glaciation—review of the
paleontological record. Proceedings of the North Dakota Academy of Science 53:171-176.
Ash worth, W. 2006. Ogallala blue: Water and life on the High Plains. The Countryman Press, Woodstock,
VT.
Atwood, G. 1994. Geomorphology applied to flooding problems of closed-basin lakes... specifically Great
Salt Lake, Utah. Geomorphology 10:197-219.
Baillie, M., J. F. Hogan, B. Ekwurzel, A. K. Wahi, and C. J. Eastoe. 2007. Quantifying water sources to a
semiarid riparian ecosystem, San Pedro River, Arizona. Journal of Geophysical Research:
Biogeosciences 112:G03S02.
Banner, B. K., A. J. Stahl, and W. K. Dodds. 2009. Stream discharge and riparian land use influence in-
stream concentrations and loads of phosphorus from Central Plains watersheds. Environmental
Management 44:552-565.
Baranyi, C., T. Hein, C. Holarek, S. Keckeis, and F. Schiemer. 2002. Zooplankton biomass and community
structure in a Danube River floodplain system: Effects of hydrology. Freshwater Biology 47:473-
482.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Batt, B. D.}., M. G. Anderson, C. D. Anderson, and F. D. Caswell. 1989. The use of prairie potholes by North
American ducks. Pages 204-227 in Northern prairie wetlands. A. G. van der Valk, editor. Iowa State
University Press, Ames, IA.
Battaglin, W. S., E. M. Thurman, S. J. Kalkhoff, and S. D. Porter. 2003. Herbicides and transformation
products in surface waters of the midwestern United States. Journal of the American Water
Resources Association 39:743-756.
Bauder, E. T. 2000. Inundation effects on small-scale plant distributions in San Diego, California vernal
pools. Aquatic Ecology 34:43-61.
Bauder, E. T. 2005. The effects of an unpredictable precipitation regime on vernal pool hydrology.
Freshwater Biology 50:2129-2135.
Bauder, E. T., and S. McMillan. 1998. Current distribution and historical extent of vernal pools in
Southern California and Northern Baja California, Mexico. Pages 56-70 in Ecology, conservation, and
management of vernal pool ecosystems—proceedings from a 1996 conference. C. W. Witham, E. T.
Bauder, D. Belk, W. R. Ferren, Jr., and R. Ornduff, editors. California Native Plant Society, Sacramento,
CA.
Bennett, S. H., and J. B. Nelson. 1991. Distribution and status of Carolina bays in South Carolina.
Nongame and Heritage Trust Publication No. 1, South Carolina Wildlife and Marine Resources
Department, Columbia, SC.
Bergey, E. A., W. J. Matthews, and J. E. Fry. 2008. Springs in time: Fish fauna and habitat changes in
springs over a 20-year interval. Aquatic Conservation: Marine and Freshwater Ecosystems 18:829-
838.
Bernot, M. J., D. J. Sobota, R. 0. Hall, P. J. Mulholland, W. K. Dodds, J. R. Webster, J. L. Tank, L. R. Ashkenas,
L. W. Cooper, C. N. Dahm, S. V. Gregory, N. B. Grimm, S. K. Hamilton, S. L. Johnson, W. H. McDowell, J.
L. Meyer, B. Peterson, G. C. Poole, H. M. Valett, C. Arango, J. J. Beaulieu, A. J. Burgin, C. Crenshaw, A. M.
Helton, L. Johnson, J. Merriam, B. R. Niederlehner, J. M. O'Brien, J. D. Potter, R. W. Sheibley, S. M.
Thomas, and K. Wilson. 2010. Inter-regional comparison of land-use effects on stream metabolism.
Freshwater Biology 55:1874-1890.
Bischel, H. N., J. E. Lawrence, B. J. Halaburka, M. H. Plumlee, A. S. Bawazir, J. P. King, J. E. McCray, V. H.
Resh, and R. G. Luthy. 2013. Renewing urban streams with recycled water for streamflow
augmentation: Hydrologic, water quality, and ecosystem services management. Environmental
Engineering Science 30:455-479.
Blann, K. L., J. L. Anderson, G. R. Sands, and B. Vondracek. 2009. Effects of agricultural drainage on
aquatic ecosystems: A review. Critical Reviews in Environmental Science and Technology 39:909-
1001.
Blasch, K. W., and J. R. Bryson. 2007. Distinguishing sources of ground water recharge by using 82H and
5180. Ground Water 45:294-308.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Bliley, D.}., and D. E. Pettry. 1979. Carolina bays on the Eastern Shore of Virginia. Soil Science Society of
America Journal 43:558-564.
Blinn, D. W., and N. L. Poff. 2005. Colorado River Basin. Pages 483-526 in Rivers of North America. A. C.
Benke and C. E. Gushing, editors. Elseviar Academic Press, Amsterdam, The Netherlands.
Boone, R. B., C. M. Johnson, and L. B. Johnson. 2006. Simulating vernal pool hydrology in central
Minnesota, USA. Wetlands 26:581-592.
Borchert, J. R. 1950. The climate of the central North American grassland. Annals of the Association of
American Geographers 40:1-39.
Boughton, E. H., P. F. Quintana-Ascencio, P. J. Bohlen, D. G. Jenkins, and R. Pickert 2010. Land-use and
isolation interact to affect wetland plant assemblages. Ecography 33:461-470.
Boulton, A. J., and P. S. Lake. 1992. The ecology of two intermittent streams in Victoria, Australia.
Freshwater Biology 27:99-121.
Braaten, P. J., and C. S. Guy. 1999. Relations between physicochemical factors and abundance of fishes in
tributary confluences of the lower channelized Missouri River. Transactions of the American
Fisheries Society 128:1213-1221.
Brant, H. A., C. H. Jagoe, J. W. Snodgrass, A. L. Bryan, and J. C. Gariboldi. 2002. Potential risk to wood
storks (Mycteria americand) from mercury in Carolina bay fish. Environmental Pollution 120:405-
413.
Brix, H. 1994. Use of constructed wetlands in water pollution control: Historical development, present
status, and future perspectives. Water Science and Technology 30:209-223.
Brooks, B. W., T. M. Riley, and R. D. Taylor. 2006. Water quality of effluent-dominated ecosystems:
Ecotoxicological, hydrological, and management considerations. Hydrobiologia 556:365-379.
Brooks, P. D., and M. M. Lemon. 2007. Spatial variability in dissolved organic matter and inorganic
nitrogen concentrations in a semiarid stream, San Pedro River, Arizona. Journal of Geophysical
Research: Biogeosciences 112:G03S05.
Brooks, R. T. 2000. Annual and seasonal variation and the effects of hydroperiod on benthic
macroinvertebrates of seasonal forest ("vernal") ponds in central Massachusetts, USA. Wetlands
20:707-715.
Brooks, R. T. 2004. Weather-related effects on woodland vernal pool hydrology and hydroperiod.
Wetlands 24:104-114.
Brooks, R. T. 2005. A review of basin morphology and pool hydrology of isolated ponded wetlands:
Implications for seasonal forest pools of the northeastern United States. Wetlands Ecology and
Management 13:335-348.
Brookshire, D. S., D. C. Goodrich, M. D. Dixon, A. Brand, K. Benedict, K. Lansey, J. Thacher, C. Broadbent, S.
Stewart, M. Mclntosh, and K. Doosun. 2010. Ecosystem services and reallocation choices: A
framework for preserving semi-arid regions in the southwest. Journal of Contemporary Water
Research and Education 144:60-74.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Brown, A. V., and W. J. Matthews. 1995. Stream ecosystems of the central United States. Pages 89-116 in
River and stream ecosystems. C. E. Gushing, K. W. Cummings, and G. W. Minshall, editors. Elsevier
Science, Amsterdam, The Netherlands.
Bull, W. B., and K. M. Scott. 1974. Impact of mining gravel from urban stream beds in the southwestern
United States. Geology 2:171-174.
Bunn, S. E., P. M. Davies, and M. Winning. 2003. Sources of organic carbon supporting the food web of an
arid zone floodplain river. Freshwater Biology 48:619-635.
Bunn, S. E., M. C. Thorns, S. K. Hamilton, and S. J. Capon. 2006. Flow variability in dryland rivers: Boom,
bust and the bits in between. River Research and Applications 22:179-186.
Cabezas, A., F. A. Comin, and D. E. Walling. 2009. Changing patterns of organic carbon and nitrogen
accretion on the Middle Ebro floodplain (NE Spain). Ecological Engineering 35:1547-1558.
Caldwell, P. V., M. J. Vepraskas, and J. D. Gregory. 2007a. Physical properties of natural organic soils in
Carolina bays of the southeastern United States. Soil Science Society of America Journal 71:1051-
1057.
Caldwell, P. V., M. J. Vepraskas, R. W. Skaggs, and J. D. Gregory. 2007b. Simulating the water budgets of
natural Carolina bay wetlands. Wetlands 27:1112-1123.
Calhoun, A. J. K., and P. G. DeMaynadier, editors. 2007. Science and conservation of vernal pools in
northeastern North America: Ecology and conservation of seasonal wetlands in northeastern North
America. CRC Press, New York, NY, USA.
Calhoun, A. J. K., T. E. Walls, S. S. Stockwell, and M. McCollough. 2003. Evaluating vernal pools as a basis
for conservation strategies: A Maine case study. Wetlands 23:70-81.
Callegary, J. B., J. M. Leenhouts, N. V. Paretti, and C. A. Jones. 2007. Rapid estimation of recharge potential
in ephemeral-stream channels using electromagnetic methods, and measurements of cahnnel and
vegetation characteristics. Journal of Hydrology 344:17-31.
Carroll, R., G. Pohll, J. Tracy, T. Winter, and R. Smith. 2005. Simulation of a semipermanent wetland basin
in the Cottonwood Lake area, east-central North Dakota. Journal of Hydrologic Engineering 10:70-
84.
Chou, R. Y. M., L. C. Ferrington, Jr., B. L. Hayford, and H. M. Smith. 1999. Composition and phenology of
Chironomidae (Diptera) from an intermittent stream in Kansas. Archiv fur Hydrobiologie 147:35-
64.
Clark, W. R. 2000. Ecology of muskrats in prairie wetlands. Pages 287-313 in Prairie wetland ecology:
The contribution of the marsh ecology research program. H. R. Murkin, A. G. van der Valk, and W. R.
Clark, editors. Iowa State University Press, Ames, IA.
Clausnitzer, D., and J. H. Huddleston. 2002. Wetland determination of a southeast Oregon vernal pool
and management implications. Wetlands 22:677-685.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Goes, A. L, and D. R. Pool. 2005. Ephemeral-stream channel and basin-floor infiltration and recharge in
the Sierra Vista subwatershed of the Upper San Pedro Basin, southeastern Arizona. USGS Open-File
Report 2005-1023, U.S. Department of the Interior, U.S. Geological Survey, Washington, DC.
Colburn, E. A. 2004. Vernal pools: Natural history and conservation. McDonald and Woodward
Publishing Company, Blacksburg, VA.
Compton, B. W., K. McGarigal, S. A. Cushman, and L. R. Gamble. 2007. A resistant-kernel model of
connectivity for amphibians that breed in vernal pools. Conservation Biology 21:788-799.
Constantz,}., A. E. Stewart, R. Niswonger, and L. Sarma. 2002. Analysis of temperature profiles for
investigating stream losses beneath ephemeral channels. Water Resources Research 38:1316.
Covich, A. P., S. C. Fritz, P. J. Lamb, R. D. Marzolf, W. J. Matthews, K. A. Poiani, E. E. Prepas, M. D. Richman,
and T. C. Winter. 1997. Potential effects of climate change on aquatic ecosystems of the Great Plains
of North America. Hydrological Processes 11:993-1021.
Cowardin, L. M., D. S. Gilmer, and L. M. Mechlin. 1981. Characteristics of central North Dakota wetlands
determined from sample aerial photographs and ground study. Wildlife Society Bulletin 9:280-288.
Crook, D. A., and B. M. Gillanders. 2006. Use of otolith chemical signatures to estimate carp recruitment
sources in the mid-Murray River, Australia. River Research and Applications 22:871-879.
Cross, F. B., and R. E. Moss. 1987. Historic changes in fish communities and aquatic habitats in plains
streams of Kansas. Pages 155-165 in Community and evolutionary ecology of North American
stream fishes. W. J. Matthews and D. C. Heins, editors. University of Oklahoma Press, Norman, OK.
Crumpton, W. G., and L. G. Goldsborough. 1998. Nitrogen transformation and fate in prairie wetlands.
Great Plains Science 8:57-72.
Cummins, K. W., and M. J. Klug. 1979. Feeding ecology of stream invertebrates. Annual Review of Ecology
and Systematics 10:147-172.
da Silva, H. P., A. C. Petry, and C. J. da Silva. 2010. Fish communities of the Pantanal wetland in Brazil:
Evaluating the effects of the upper Paraguay river flood pulse on baia Caicara fish fauna. Aquatic
Ecology 44:275-288.
Dahl, T. E. 1990. Wetlands Losses in the United States 1780's to 1980's. U.S. Department of the Interior,
U.S. Fish and Wildlife Service, Washington, DC.
Dahm, C. N., J. R. Cleverly, J. E. Allred Coonrod, J. R. Thibault, D. E. McDonnell, and D. J. Gilroy. 2002.
Evapotranspiration at the land/water interface in a semi-arid drainage basin. Freshwater Biology
47:831-843.
Deacon, J. E. 1961. Fish populations, following a drought, in the Neosho and Marais de Cygnes Rivers of
Kansas. Museum of Natural History 13:359-427.
Dickinson, J. E., J. R. Kennedy, D. R. Pool, J. T. Cordova, J. T. Parker, J. P. Macy, and B. Thomas. 2010.
Hydrogeologic framework of the middle San Pedro watershed, southeastern Arizona. USGS Scientific
Investigations Report 2010-5126, prepared in cooperation with the Arizona Department of Water
Resources, U.S. Department of the Interior, U.S. Geological Survey, Reston, VA.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Dodds, W. K., J. M. Blair, G. M. Henerbry, J. K. Koelliker, R. Ramundo, and C. M. Tate. 1996a. Nitrogen
transport from tallgrass prairie watersheds. Journal of Environmental Quality 25:973-981.
Dodds, W. K., M. A. Evans-White, N. M. Gerlanc, L. Gray, D. A. Gudder, M. J. Kemp, A. L. Lopez, D. Stagliano,
E. A. Strauss, J. L. Tank, M. R. Whiles, and W. M. Wollheim. 2000. Quantification of the nitrogen cycle
in a prairie stream. Ecosystems 3:574-589.
Dodds, W. K., K. Gido, M. R. Whiles, K. M. Fritz, and W. J. Matthews. 2004. Life on the edge: The ecology of
Great Plains prairie streams. BioScience 54:205-216.
Dodds, W. K., R. E. Hutson, A. C. Eichem, M. A. Evans, D. A. Gudder, K. M. Fritz, and L. Gray. 1996b. The
relationship of floods, drying, flow and light to primary production and producer biomass in a
prairie stream. Hydrobiologia 333:151-159.
Dodds, W. K., and R. M. Oakes. 2004. A technique for establishing reference nutrient concentrations
across watersheds affected by humans. Limnology and Oceanography-Methods 2:333-341.
Dodds, W. K., and R. M. Oakes. 2006. Controls on nutrients across a prairie stream watershed: Land use
and riparian cover effects. Environmental Management 37:634-646.
Dodds, W. K., and R. M. Oakes. 2008. Headwater influences on downstream water quality. Environmental
Management 41:367-377.
Donald, D. B., J. Syrgiannis, F. Hunter, and G. Weiss. 1999. Agricultural pesticides threaten the ecological
integrity of northern prairie wetlands. The Science of the Total Environment 231:173-181.
Du, B., J. G.Arnold, A. Saleh, and D. B. Jaynes. 2005. Development and application of SWAT to landscapes
with tiles and potholes. Transactions of the American Society of Agricultural Engineers 48:1121-
1133.
Durham, B. W., and G. R. Wilde. 2006. Influence of stream discharge on reproductive success of a prairie
stream fish assemblage. Transactions of the American Fisheries Society 135:1644-1653.
Durham, B. W., and G. R. Wilde. 2008. Composition and abundance of drifting fish larvae in the Canadian
River, Texas. Journal of Freshwater Ecology 23:273-280.
Eby, L. A., W. F. Pagan, and W. L. Minckley. 2003. Variability and dynamics of a desert stream community.
Ecological Applications 13:1566-1579.
Edwards, A. L., and A. S. Weakley. 2001. Population biology and management of rare plants in depression
wetlands of the southeastern coastal plain, USA. Natural Areas Journal 21:12-35.
England, A. S., and W. F. Laudenslayer. 1995. Birds of the California Desert. Pages 337-372 in The
California desert: An introduction to natural resources and man's impact, Volume 2. J. Latting and P.
G. Rowlands, editors. June Latting Books, Riverside, CA.
Euliss, N. H., Jr., and D. M. Mushet. 2004. Impacts of water development on aquatic macroinvertebrates,
amphibians, and plants in wetlands of a semi-arid landscape. Aquatic Ecosystem Health &
Management 7:73-84.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Euliss, N. H., J. W. Labaugh, L. H. Fredrickson, D. M. Mushet, M. R. K. Laubhan, G. A. Swanson, T. C. Winter,
D. 0. Rosenberry, and R. D. Nelson. 2004. The wetland continuum: A conceptual framework for
interpreting biological studies. Wetlands 24:448-458.
Evans-White, M. A., W. K. Dodds, L. J. Gray, and K. M. Fritz. 2001. A comparison of the trophic ecology of
the crayfishes (Orconectes nais (Faxon) and Orconectes neglectus (Faxon)) and the central
stoneroller minnow [Campostoma anomalum (Rafinesque)): Omnivory in a tallgrass prairie stream.
Hydrobiologia 462:131-144.
Evans-White, M. A., W. K. Dodds, and M. R. Whiles. 2003. Ecosystem significance of crayfishes and
stonerollers in a prairie stream: Functional differences between co-occurring omnivores. Journal of
the North American Benthological Society 22:423-441.
Falke, J. A., K. R. Bestgen, and K. D. Fausch. 2010. Streamflow reductions and habitat drying affect
growth, survival, and recruitment of brassy minnow across a Great Plains landscape. Transactions of
the American Fisheries Society 139:1566-1583.
Falke, J. A., K. D. Fausch, R. Magelky, A. Aldred, D. S. Durnford, L. K. Riley, and R. Oad. 2011. The role of
groundwater pumping and drought in shaping ecological futures for stream fishes in a dryland river
basin of the western Great Plains, USA. Ecohydrology 4:682-697.
Falke, J. A., and K. B. Gido. 2006. Effects of reservoir connectivity on stream fish assemblages in the Great
Plains. Canadian Journal of Fisheries and Aquatic Sciences 63:480-493.
Farag, A. M., D. D. Harper, A. Senecal, and W. A. Hubert. 2010. Potential effects of coalbed natural gas
development on fish and aquatic resources. Pages 227-242 in Coalbed natural gas: Energy and
environment. K. R. Reddy, editor. Nova Science Publishers, New York, NY.
Fausch, K. D., and K. R. Bestgen. 1997. Ecology of fishes indigenous to the central and southwestern
Great Plains. Pages 131-166 in Ecology and conservation of Great Plains vertebrates. F. L. Knopf and
F. B. Samson, editors. Springer-Verlag, New York, NY.
Fausch, K. D., and R. G. Bramblett. 1991. Disturbance and fish communities in intermittent tributaries of
a western Great Plains river. Copeia 1991:659-674.
Fausch, K. D., C. E. Torgersen, C. V. Baxter, and H. W. Li. 2002. Landscapes to riverscapes: Bridging the
gap between research and conservation of stream fishes. BioScience 52:483-498.
Fernald, A. F., P. J. Wigington, and D. Landers. 2001. Transient storage and hyporheic flow along the
Willamette River, Oregon: Field measurements and model estimates. Water Resources Research
37:1681-1694.
Ferrington, L. C. 1993. Endangered rivers: A case history of the Arkansas River in the Central Plains.
Aquatic Conservation: Marine and Freshwater Ecosystems 3:305-316.
Figuerola, J., A. J. Green, and T. C. Michot. 2005. Invertebrate eggs can fly: Evidence of waterfowl-
mediated gene flow in aquatic invertebrates. The American Naturalist 165:274-280.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Franssen, N. R., K. B. Gido, C. S. Guy, J. A. Tripe, S. J. Shrank, T. R. Strakosh, K. N. Bertrand, C. M. Franssen,
K. L. Pitts, and C. P. Paukert 2006. Effects of floods on fish assemblages in an intermittent prairie
stream. Freshwater Biology 51:2072-2086.
Friedman, J. M., W. R. Osterkamp, M. L. Scott, and G. T. Auble. 1998. Downstream effects of dams on
channel geometry and bottomland vegetation: Regional patterns in the Great Plains. Wetlands
18:619-633.
Frisch, D., A. J. Green, and J. Figuerola. 2007. High dispersal capacity of a broad spectrum of aquatic
invertebrates via waterbirds. Aquatic Sciences 69:568-574.
Fritz, K. M., and W. K. Dodds. 2002. Macroinvertebrate assemblage structure across a tallgrass prairie
stream landscape. Archivfur Hydrobiologie 154:79-102.
Fritz, K. M., and W. K. Dodds. 2004. Resistance and resilience of macroinvertebrate assemblages to
drying and flood in a tallgrass prairie stream system. Hydrobiologia 527:99-112.
Fritz, K. M., and W. K. Dodds. 2005. Harshness: Characterization of intermittent stream habitat over
space and time. Marine and Freshwater Research 56:13-23.
Fritz, K. M., J. A. Tripe, and C. S. Guy. 2002. Recovery of three fish species to flood and seasonal drying in
a tallgrass prairie stream. Transactions of the Kansas Academy of Science 105:209-219.
Gahl, M. K., A. J. K. Calhoun, and R. Graves. 2009. Facultative use of seasonal pools by American bullfrogs
(Rana catesbeiand). Wetlands 29:697-703.
Galat, D. L., C. R. Berry, Jr., E. J. Peters, and R. G. White. 2005. Missouri River basin. Pages 427-480 in
Rivers of North America. A. C. Benke and C. E. Gushing, editors. Elsevier Academic Press, Burlington,
MA.
Galat, D. L., and R. Lipkin. 2000. Restoring ecological integrity of great rivers: Historical hydrographs aid
in defining reference conditions for the Missouri River. Hydrobiologia 422/423:29-48.
Galatowitsch, S. M., and A. G. van der Valk. 1996. The vegetation of restored and natural prairie wetlands.
Ecological Applications 6:102-112.
Galbarczyk-Gasiorowska, L., M. Gasiorowski, and K. Szeroczynska. 2009. Reconstruction of human
influence during the last two centuries on two small oxbow lakes near Warsaw (Poland).
Hydrobiologia 631:173-183.
Gallardo, B., M. Garcia, A. Cabezas, E. Gonzalez, M. Gonzalez, C. Ciancarelli, and F. A. Comin. 2008.
Macroinvertebrate patterns along environmental gradients and hydrological connectivity within a
regulated river-floodplain. Aquatic Sciences 70:248-258.
Gelroth, J. V., and G. R. Marzolf. 1978. Primary production and leaf-litter decomposition in natural and
channelized portions of a Kansas stream. American Midland Naturalist 99:238-243.
Gelwick, F. P., and W. J. Matthews. 1997. Effects of algivorous minnows (Campostoma) on spatial and
temporal heterogeneity of stream periphyton. Oecologia 112:386-392.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Gibbons, J. W., and R. D. Semlitsch. 1991. Guide to amphibians and reptiles of the Savannah River Site.
University of Georgia Press, Athens, GA.
Gibbons, J. W., C. T. Winne, D. E. Scott, J. D. Willson, X. Glaudas, K. M. Andrews, B. D. Todd, L. A. Fedewa, L.
Wilkinson, R. N. Tsaliagos, S. J. Harper, J. L. Greene, T. D. Tuberville, B. S. Metts, M. E. Dorcast, J. P.
Nestor, C. A. Young, T. Akre, R. N. Reed, K. A. Buhlmann, J. Norman, D. A. Croshaw, C. Hagen, and B. B.
Rothermel. 2006. Remarkable amphibian biomass and abundance in an isolated wetland:
Implications for wetland conservation. Conservation Biology 20:1457-1465.
Ginting, D., J. F. Moncrief, and S. C. Gupta. 2000. Runoff, solids, and contaminant losses into surface tile
inlets draining lacustrine depressions. Journal of Environmental Quality 29:551-560.
Gleason, R. A., N. H. Euliss, D. Hubbard, and W. Duffy. 2003. Effects of sediment load on emergence of
aquatic invertebrates and plants from wetland soil egg and seed banks. Wetlands 23:26-34.
Gleason, R. A., N. H. Euliss, D. E. Hubbard, and W. G. Duffy. 2004. Invertebrate egg banks of restored,
natural, and drained wetlands in the prairie pothole region of the United States. Wetlands 24:562-
572.
Gleason, R. A., M. K. Laubhan, and N. H. Euliss, Jr. 2008. Ecosystem services derived from wetland
conservation practices in the United States prairie pothole region with an emphasis on the U.S.
Department of Agriculture Conservation Reserve and Wetlands Reserve Programs. USGS
Professional Paper 1745, U.S. Department of the Interior, U.S. Geological Survey, Reston, VA.
Gleason, R. A., B. A. Tangen, M. K. Laubhan, K. E. Kermes, and N. H. Euliss. 2007. Estimating water storage
capacity of existing and potentially restorable wetland depressions in a subbasin of the Red River of
the North. USGS Open-File Report 2007-1159, U.S. Department of the Interior, U.S. Geological
Survey, Reston, VA.
Glinska-Lewczuk, K. 2009. Water quality dynamics of oxbow lakes in young glacial landscape of NE
Poland in relation to their hydrological connectivity. Ecological Engineering 35:25-37.
Golladay, S. W. 1997. Suspended particulate organic matter concentration and export in streams. Journal
of the North American Benthological Society 16:122-131.
Golladay, S. W., and C. L. Hax. 1995. Effects of an engineered flow disturbance on meiofauna in a north
Texas prairie stream. Journal of the North American Benthological Society 14:404-413.
Goodrich, D. C., A. Chehbouni, B. Goff, R. Mac Nish, T. Maddock, M. S. Moran, J. Shuttleworth, D. G.
Williams, C. Watts, L. H. Hipps, D. I. Cooper, J. Schieldge, Y. H. Kerr, H. Arias, M. Kirkland, R. Carlos, P.
Cayrol, W. Kepner, B. Jones, R. Avissar, A. Begue, J.-M. Bonnefond, G. Boulet, B. Branan, J. P. Brunei, L.
C. Chen, T. Clarke, M. R. Davis, H. DeBruin, G. Dedieu, E. Elguero, W. E. Eichinger, J. Everitt, J.
Garatuza-Payan, V. L. Gempko, H. Gupta, C. Harlow, 0. Hartogensis, M. Helfert, C. Holifield, D. Hymer,
A. Kahle, T. Reefer, S. Krishnamoorthy, J.-P. Lhomme, J.-P. Lagouarde, D. Lo Seen, D. Luquet, R.
Marsett, B. Monteny, W. Ni, Y. Nouvellon, R. Pinker, C. Peters, D. Pool, J. Qi, S. Rambal, J. Rodriguez, F.
Santiago, E. Sano, S. M. Schaeffer, M. Schulte, R. Scott, X. Shao, K. A. Snyder, S. Sorooshian, C. L.
Unkrich, M. Whitaker, and I. Yucel. 2000. Preface paper to the Semi-Arid Land-Surface-Atmosphere
(SALSA) program special issue. Agricultural and Forest Meteorology 105:3-20.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Goodrich, D. C., L. J. Lane, R. M. Shillito, S. N. Miller, K. H. Syed, and D. A. Woolhiser. 1997. Linearity of
basin response as a function of scale in a semiarid watershed. Water Resources Research 33:2951-
2965.
Goodrich, D. C., D. G. Williams, C. L. Unkrich, J. F. Hogan, R. L. Scott, K. R. Hultine, D. R. Pool, A. L. Goes, and
S. Miller. 2004. Comparison of methods to estimate ephemeral channel recharge, Walnut Gulch, San
Pedro River basin, Arizona. Pages 77-99 in Recharge and vadose zone processes: Alluvial basins of
the southwestern United States. F. M. Phillips, J. F. Hogan, and B. Scanlon, editors. American
Geophysical Union, Washington, DC.
Gornall, R. J., P. M. Hollingsworth, and C. D. Preston. 1998. Evidence for spatial structure and directional
gene flow in a population of an aquatic plant, Potamogeton coloratus. Heredity 80:414-421.
Graf, W. L. 1988. Fluvial processes in dryland rivers. Springer New York.
Grant, E. H. C. 2005. Correlates of vernal pool occurrence in the Massachusetts, USA landscape. Wetlands
25:480-487.
Grant, J. A., M. J. Brooks, and B. E. Taylor. 1998. New constraints on the evolution of Carolina bays from
ground-penetrating radar. Geomorphology 22:325-345.
Gray, L. J., and K. W. Johnson. 1988. Trophic structure of benthic macroinvertebrates in Kings Creek.
Transactions of the Kansas Academy of Science 91:178-184.
Gregoire, C., D. Elsaesser, D. Huguenot, J. Lange, T. Lebeau, A. Merli, R. Mose, E. Passeport, S. Payraudeau,
T. Schutz, R. Schulz, G. Tapia-Padilla, J. Tournebize, M. Trevisan, and A. Wanko. 2009. Mitigation of
agricultural nonpoint-source pesticide pollution in artificial wetland ecosystems. Environmental
Chemistry Letters 7:205-231.
Gregory, K. J. 1976. Drainage networks and climate. Pages 289-315 in Geomorphology and climate. E.
Derbyshire, editor. John Wiley & Sons, London, UK.
Gregory, S. V., F. J. Swanson, W. A. McKee, and K. W. Cummings. 1991. An ecosystem perspective of
riparian zones: Focus on links between land and water. BioScience 41:540-551.
Groffman, P. M., A. J. Gold, and R. C. Simmons. 1992. Nitrate dynamics in riparian forests: Microbial
studies. Journal of Environmental Quality 21:666-671.
Gurtz, M. E., G. R. Marzolf, D. L. Killingbeck, D. L. Smith, and J. V. McArthur. 1988. Hydrologic and riparian
influences on the import and storage of coarse particulate organic matter in a prairie stream.
Canadian Journal of Fisheries and Aquatic Sciences 45:655-665.
Gurtz, M. E., G. R. Marzolf, K. T. Killingbeck, D. L. Smith, and J. V. McArthur. 1982. Organic matter loading
and processing in a pristine stream draining a tallgrass prairie/riparian forest watershed. Kansas
Water Resources Research Institute Contribution 230, Manhattan, KS.
Hadley, R. F., M. R. Karlinger, A. W. Burns, and T. R. Eschner. 1987. Water development and associated
hydrologic changes in the Platte River, Nebraska, U.S.A. Regulated Rivers: Research & Mangement
1:331-341.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Hall, D. H., and R. J. Steidl. 2007. Movements, activity, and spacing of Sonoran mud turtles (Kinosternon
sonoriense) in interrupted mountain streams. Copeia 2007:403-412.
Hanes, T., and L. Stromberg. 1998. Hydrology of vernal pools on non-volcanic soils in the Sacramento
Valley. Pages 38-49 in Ecology, conservation, and management of vernal pool ecosystems—
proceedings from a 1996 Conference. C. W. Witham, E. T. Bauder, D. Belk, W. R. Ferren, Jr., and R.
Ornduff, editors. California Native Plant Society, Sacramento, CA.
Hanson, M. A., K. D. Zimmer, M. G. Butler, B. A. Tangen, B. R. Herwig, and N. H. Euliss. 2005. Biotic
interactions as determinants of ecosystem structure in prairie wetlands: An example using fish.
Wetlands 25:764-775.
Harrell, R. C., B. J. Davis, and T. C. Dorris. 1967. Stream order and species diversity of fishes in an
intermittent Oklahoma stream. American Midland Naturalist 78:428-436.
Harrington, G. A., P. G. Cook, and A. L. Herczeg. 2002. Spatial and temporal variability of ground water
recharge in central Australia: A tracer approach. Ground Water 40:518-528.
Harris, M. A., B. C. Kondratieff, and T. P. Boyle. 1999. Macroinvertebrate community structure of three
prairie streams. Journal of the Kansas Entomological Society 72:402-425.
Hastings, J. R. 1959. Vegetation change and arroyo cutting in southeastern Arizona. Journal of the
Arizona Academy of Science:60-67.
Hax, C. L., and S. W. Golladay. 1998. Flow disturbance of macroinvertebrates inhabiting sediments and
woody debris in a prairie stream. American Midland Naturalist 139:210-223.
Hayashi, M., G. van der Kamp, and D. L. Rudolph. 1998. Water and solute transfer between a prairie
wetland and adjacent uplands, 1. Water balance. Journal of Hydrology 207:42-55.
Hedman, E. R., and W. R. Osterkamp. 1982. Streamflow characteristics related to channel geometry of
streams in western United States. USGS Water Supply Paper 2193, U.S. Department of the Interior,
U.S. Geological Survey, Alexandria, VA.
Hein, T., C. Baranyi, G. J. Herndl, W. Wanek, and F. Schiemer. 2003. Allochthonous and autochthonous
particulate organic matter in floodplains of the River Danube: The importance of hydrological
connectivity. Freshwater Biology 48:220-232.
Hentges, V. A., and T. W. Stewart. 2010. Macroinvertebrate assemblages in Iowa prairie pothole wetlands
and relation to environmental features. Wetlands 30:501-511.
Herwig, B. R., K. D. Zimmer, M. A. Hanson, M. L. Konsti, J. A. Younk, R. W. Wright, S. R. Vaughn, and M. D.
Haustein. 2010. Factors influencing fish distributions in shallow lakes in prairie and prairie-
parkland regions of Minnesota, USA. Wetlands 30:609-619.
Hill, B. H., and T. J. Gardner. 1987a. Benthic metabolism in a perennial and an intermittent Texas prairie
stream. Southwestern Naturalist 32:305-311.
Hill, B. H., and T. J. Gardner. 1987b. Seston dynamics in two Texas prairie streams. American Midland
Naturalist 118:85-93.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Hill, B. H., T. J. Gardner, and 0. F. Ekisola. 1992. Predicatability of streamflow and particulate organic
matter concentration as indicators of stability in prairie streams. Hydrobiologia 242:7-18.
Hilty, J. A., W. Z. Lidicker Jr, and A. Merenlender. 2006. Corridor ecology: The science and practice of
linking landscapes for biodiversity conservation. Island Press, Washington, DC.
Hoagstrom, C. W., J. E. Brooks, and S. R. Davenport. 2010. A large-scale conservation perspective
considering endemic fishes of the North American plains. Biological Conservation 144:21-34.
Hobson, W. A., and R. A. Dahlgren. 1998. Soil forming processes in vernal pools of Northern California,
Chico area. Pages 24-37 in Ecology, conservation, and management of vernal pool ecosystems—
proceedings from a 1996 conference. C. W. Witham, E. T. Bauder, D. Belk, W. R. Ferren, Jr., and R.
Ornduff, editors. California Native Plant Society, Sacramento, CA.
Horowitz, A. J., K. A. Elrick, and E. Callender. 1988. The effect of mining on the sediment-trace element
geochemistry of cores from the Cheyenne River Arm of Lake Oahe, South Dakota, USA. Chemical
Geology 67:17-33.
Hubbard, D. E., and R. L. Linder. 1986. Spring runoff retention in prairie pothole wetlands. Journal of Soil
and Water Conservation 41:122-125.
Hubert, W. A., and K. M. Gordon. 2007. Great Plains fishes declining or threatened with extirpation in
Montana, Wyoming, or Colorado. Pages 3-13 in Status, distribution, and conservation of native
freshwater fishes of western North America. M. J. Brouder and J. A. Scheurer, editors. American
Fisheries Society Symposium 53, Bethesda, MD.
Hudson, P. L., D. R. Lenat, B. A. Caldwell, and D. Smith. 1990. Chironomidae of the southeastern United
States: A checklist of species and notes on biology, distribution, and habitat. Fish and Wildlife
Research 7, U.S. Department of the Interior, U.S. Fish and Wildlife Service, Washington, DC.
Hulsmans, A., K. Moreau, L. De Meester, B. J. Riddoch, and L. Brendonck. 2007. Direct and indirect
measures of dispersal in the fairy shrimp Branchipodopsis wolfi indicate a small-scale isolation-by-
distance pattern. Limnology and Oceanography 52:676-684.
Humphries, P., A. J. King, and J. D. Koehn. 1999. Fish, flows and flood plains: Links between freshwater
fishes and their environment in the Murray-Darling River system, Australia. Environmental Biology
of Fishes 56:129-151.
Hunsinger, T. W., and M. J. Lannoo. 2005. Notophthalmus viridescens, eastern newt. Pages 912-914 in
Amphibian declines: The conservation status of United States species. M. Lannoo, editor. University
of California Press, Berkeley, CA.
Huntzinger, T. L. 1995. Surface water: A critical resource of the Great Plains. Pages 253-273 in
Conservation of Great Plains ecosystems: Current science, future options. S. R. Johnson and A.
Bouzaher, editors. Kluwer Academic Publishers, Dordrecht, The Netherlands.
Jackson, C. R., and C. M. Pringle. 2010. Ecological benefits of reduced hydrologic connectivity in
intensively developed landscapes. BioScience 60:37-46.
Jewell, M. E. 1927. Aquatic biology of the prairie. Ecology 8:289-298.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
John, K. R. 1964. Survival offish in intermittent streams of the Chirichua Mountains, Arizona. Ecology
45:112-119.
Johnson, D. W. 1942. The Origin of Carolina Bays. Columbia Univeristy Press, New York, NY.
Johnson, S. L, and A. P. Covich. 1997. Scales of observation of riparian forests and distributions of
suspended detritus in a prairie river. Freshwater Biology 37:163-175.
Johnson, W. C. 1994. Woodland expansion in the Platte River, Nebraska: Patterns and causes. Ecological
Monographs 64:45-84.
Johnson, W. C., S. E. Boettcher, K. A. Poiani, and G. Guntenspergen. 2004. Influence of weather extremes
on the water levels of glaciated prairie wetlands. Wetlands 24:385-398.
Johnson, W. C., B. V. Millett, T. Gilmanov, R. A. Voldseth, G. R. Guntenspergen, and D. E. Naugle. 2005.
Vulnerability of northern prairie wetlands to climate change. BioScience 55:863-872.
Johnson, Z. B., and J. H. Kennedy. 2003. Macroinvertebrate assemblages of submerged woody debris in
the Elm Fork of the Trinity River, Texas. Journal of Freshwater Ecology 18:187-197.
Junk, W. J., P. B. Bayley, and R. E. Sparks. 1989. The flood pulse concept in river-floodplain systems. Pages
110-127 in Proceedings of the international large river symposium, Canadian Special Publication of
Fisheries and Aquatic Sciences 106. D. P. Dodge, editor., Ottawa, Canada.
Kahara, S. N., R. M. Mockler, K. F. Higgins, S. R. Chipps, and R. R. Johnson. 2009. Spatiotemporal patterns
of wetland occurrence in the prairie pothole region of eastern South Dakota. Wetlands 29:678-689.
Kantrud, H. A., G. L. Krapu, and G. A. Swanson. 1989. Prairie basin wetlands of the Dakotas: A community
profile. Biological Report 85(7.28), U.S. Department of the Interior, U.S. Fish and Wildlife Service and
U.S. Environmental Protection Agency, Washington, DC.
Keckeis, S., C. Baranyi, T. Hein, C. Holarek, P. Riedler, and F. Schiemer. 2003. The significance of
zooplankton grazing in a floodplain system of the River Danube. Journal of Plankton Research
25:243-253.
Keeley, J. E., and P. H. Zedler. 1998. Characterization and global distribution of vernal pools. Pages 1-14
in Ecology, conservation, and management of vernal pool ecosystems—proceedings from a 1996
conference. C. W. Witham, E. T. Bauder, D. Belk, W. R. Ferren, Jr., and R. Ornduff, editors. California
Native Plant Society, Sacramento, CA.
Keiper, J. B., W. E. Walton, and B. A. Foote. 2002. Biology and ecology of higher diptera from freshwater
wetlands. Annual Review of Entomology 47:207-232.
Kemp, M. J., and W. K. Dodds. 2001. Spatial and temporal patterns of nitrogen concentrations in pristine
and agriculturally-influenced prairie streams. Biogeochemistry 53:125-141.
Kemp, M. J., and W. K. Dodds. 2002. Comparisons of nitrification and denitrification in prairie and
agriculturally influenced streams. Ecological Applications 12:998-1009.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Kennedy,}., D. Goodrich, and C. Unkrich. 2013. Using the KINEROS2 modeling framework to evaluate the
increase in storm runoff from residential development in a semiarid environment. Journal of
Hydrologic Engineering 18:698-706.
Kennedy, J. R., and B. Gungle. 2010. Quantity and sources of base flow in the San Pedro River near
Tombstone, Arizona. USGS Scientific Investigations Report 2010-5200, U.S. Department of the
Interior, U.S. Geological Survey, Reston, VA.
Kepner, W. G., D. J. Semmens, S. D. Bassett, D. A. Mouat, and D. C. Goodrich. 2004. Scenario analysis for
the San Pedro River, analyzing hydrological consequences of a future environment. Environmental
Monitoring and Assessment 94:115-127.
Kepner, W. G., C. J. Watts, C. M. Edmonds, J. K. Maingi, S. E. Marsh, and G. Luna. 2000. A landscape
approach for detecting and evaluating change in a semi-arid environment. Environmental
Monitoring and Assessment 64:179-195.
King, J. L, M. A. Simovich, and R. C. Brusca. 1996. Species richness, endemism and ecology of crustacean
assemblages in northern California vernal pools. Hydrobiologia 328:85-116.
Kirkman, L. K, and R. R. Sharitz. 1994. Vegetation disturbance and maintenance of diversity in
intermittently flooded Carolina bays in South Carolina. Ecological Applications 4:177-188.
Knight, R. L., J. S. Bays, and F. R. Richardson. 1989. Floral composition, soil relations and hydrology of a
Carolina bay in South Carolina. Pages 219-234 in Freshwater wetlands and wildlife. R. R. Sharitz and
J. W. Gibbons, editors. U.S. Department of Energy, Office of Scientific and Technical Information, Oak
Ridge, TN.
Knosche, R. 2006. Organic sediment nutrient concentrations and their relationship with the hydrological
connectivity of floodplain waters (River Havel, NE Germany). Hydrobiologia 560:63-76.
Krapu, G. L., J. G. Raymond, C. P. Dwyer, K. M. Kraft, and L. M. Cowardin. 1997. Wetland use, settling
patterns, and recruitment in mallards. Journal of Wildlife Management 61:736-746.
LaBaugh, J. W., T. C. Winter, G. A. Swanson, D. 0. Rosenberry, R. D. Nelson, and N. H. Euliss. 1996. Changes
in atmospheric circulation patterns affect midcontinent wetlands sensitive to climate. Limnology
and Oceanography 41:864-870.
Labbe, T. R., and K. D. Fausch. 2000. Dynamics of intermittent stream habitat regulate persistence of a
threatened fish at multiple scales. Ecological Applications 10:1774-1791.
Lane, L. J., M. Hernandez, and M. H. Nichols. 1997. Processes controlling sediment yield from watersheds
as functions of spatial scale. Environmental Modelling and Software 12:355-369.
Lang, M., 0. McDonough, G. McCarty, R. Oesterling, and B. Wilen. 2012. Enhanced detection of wetland-
stream connectivity using LiDAR. Wetlands 32:461-473.
Lauenroth, W. K., I. C. Burke, and M. P. Gutmann. 1999. The structure and function of ecosystems in the
central North American grassland region. Great Plains Research 9:223-259.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Leeper, D. A., and B. E. Taylor. 1998. Insect emergence from a South Carolina (USA) temporary wetland
pond, with emphasis on the Chironomidae (Diptera). Journal of the North American Benthological
Society 17:54-72.
Lehtinen, R. M., and S. M. Galatowitsch. 2001. Colonization of restored wetlands by amphibians in
Minnesota. American Midland Naturalist 145:388-396.
Leibowitz, S. G., and K. C. Vining. 2003. Temporal connectivity in a prairie pothole complex. Wetlands
23:13-25.
Lekach,}., A. P. Shick, and A. Schlesinger. 1992. Bedload yield and in-channel provenance in a flash-flood
fluvial system. Pages 537-554 in Dynamics of gravel-bed rivers. P. Billi, R. D. Hey, C. R. Thorne, and P.
Tacconi, editors. John Wiley & Sons, New York, NY.
Lenhart, C. F., K. N. Brooks, D. Heneley, and J. A. Magner. 2010. Spatial and temporal variation in
suspended sediment, organic matter, and turbidity in a Minnesota prairie river: Implications for
TMDLs. Environmental Monitoring and Assessment 165:435-447.
Lerner, D. N. 1986. Leaking pipes recharge ground water. Ground Water 24:654-662.
Levick, L, J. Fonseca, D. Goodrich, M. Hernandez, D. Semmens, R. Leidy, M. Scianni, P. Guertin, M. Tluczek,
and W. Kepner. 2008. The ecological and hydrological significance of ephemeral and intermittent
streams in the arid and semi-arid American Southwest. EPA/600/R-08/134 and ARS/233046, U.S.
Environmental Protection Agency, Office of Research and Development and U.S. Department of
Agriculture/Agricultural Research Service. Southwest Watershed Research Center, Washington, DC.
Lide, R. F., V. G. Meentemeyer, J. E. Pinder, and L. M. Beatty. 1995. Hydrology of a Carolina bay located on
the upper Coastal Plain of western South Carolina. Wetlands 15:47-57.
Lorenz, D. L., C. A. Sanocki, and M. J. Kocian. 2010. Techniques for estimating the magnitude and
frequency of peak flows on small streams in Minnesota based on through water year 2005. USGS
Scientific Investigations Report 2009-5250, U.S. Department of the Interior, U.S. Geological Survey,
in cooperation with the Minnesota Department of Transportation and the Minnesota Pollution
Control Agency, Reston, VA.
Louw, G., and M. Seely. 1982. Ecology of desert organisms.
Luttrell, G. R., A. A. Echelle, W. L. Fisher, and D. J. Eisenhour. 1999. Declining status of two species of the
Macrhybopsis aestivalis complex (Teleostei: Cyprinidae) in the Arkansas River basin and related
effects of reservoirs as barriers to dispersal. Copeia 1999:981-989.
Lytle, D. A., M. T. Bogan, and D. S. Finn. 2008. Evolution of aquatic insect behaviors across a gradient of
disturbance predictability. Proceedings of the Royal Society - Series B 275:453-462.
Machavaram, M. V., D. 0. Whittemore, M. E. Conrad, and N. L. Miller. 2006. Precipitation induced stream
flow: An event based chemical and isotopic study of a small stream in the Great Plains region of the
USA. Journal of Hydrology 330:470-480.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Macpherson, G. L, and M. Sophocleous. 2004. Fast ground-water mixing and basal recharge in an
unconfined alluvial aquifer, Konza LTER site, northeastern Kansas. Journal of Hydrology 286:271-
299.
Mahoney, D. L, M. A. Mort, and B. E. Taylor. 1990. Species richness of calanoid copepods, cladocerans
and other branchiopods in Carolina bay temporary ponds. American Midland Naturalist 123:244-
258.
Marron, D. C. 1989. The transport of mine tailings as suspended sediment in the Belle Fourche River,
west-central South Dakota, USA. International Association of Hydrologic Sciences 184:19-26.
Mathis, B.}., and T. C. Dorris. 1968. Community structure of benthic macroinvertebrates in an
intermittent stream receiving oil field brines. American Midland Naturalist 80:428-439.
Matthai, H. F. 1969. Floods of June 1965 in South Platte River basin, Colorado. USGS Water Supply Paper
1850-B., U.S. Department of the Interior, U.S. Geological Survey, Reston, VA.
Matthews, W. J. 1988. North American prairie streams as systems for ecological study. Journal of the
North American Benthological Society 7:387-409.
Matthews, W. J., J. F. Hoover, and W. B. Milstead. 1985. Fishes of Oklahoma springs. Southwestern
Naturalist 30:2 3-32.
Matthews, W. J., and E. Marsh-Matthews. 2007. Extirpation of red shiner in direct tributaries of Lake
Texoma (Oklahoma-Texas): A cautionary case history from a fragmented river-reservoir system.
Transactions of the American Fisheries Society 136:1041-1062.
Matthews, W. J., and H. W. Robinson. 1998. Influence of drainage connectivity, drainage area and
regional species richness on fishes of the interior highlands in Arkansas. American Midland
Naturalist 139:1-19.
Matthews, W. J., C. C. Vaughn, K. B. Gido, and E. Marsh-Matthews. 2005. Southern Plains rivers. Pages
283-325 in Rivers of North America. A. C. Benke and C. E. Gushing, editors. Elsevier Academic Press,
Burlington, MA.
Matthews, W. J., and E. G. Zimmerman. 1990. Potential effects of global warming on native fishes of the
southern Great Plains and the Southwest. Fisheries 15:26-32.
McArthur, J. V., M. E. Gurtz, C. M. Tate, and F. S. Gilliam. 1985a. The interaction of biological and
hydrological phenomena that mediate the qualities of water draining native tallgrass prairie on the
Konza Prairie Research Natural Area. Pages 478-482 in Perspectives on nonpoint source pollution,
proceedings of 1985 national conference. U.S. Environmental Protection Agency, Office of Water
Regulations and Standards, Washington, DC.
McArthur, J. V., G. R. Marzolf, and J. E. Urban. 1985b. Response of bacteria isolated from a pristine prairie
stream to concentration and source of soluble organic carbon. Applied and Environmental
Microbiology 49:238-241.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
McCoy, R. W., and D. C. Hales. 1974. A survey of eight streams in eastern South Dakota: Physical and
chemical characteristics, vascular plants, insects and fishes. Proceedings of the South Dakota
Academy of Sciences 53:202-219.
Meffe, G. K. 1984. Effects of abiotic disturbance on coexistence of predator-prey fish species. Ecology
65:1525-1534.
Meinzer, 0. E. 1923. Outline of ground-water hydrology. US Geology Survey Water Supply Paper 494,
U.S. Department of the Interior, U.S. Geological Survey, Washington, DC.
Meixner, T. A., A. K. Huth, P. D. Brooks, M. H. Conklin, N. B. Grimm, R. C. Bales, P. A. Haas, and J. R. Petti.
2007. Influence of shifting flow paths on nitrogen concentrations during monsoon floods, San Pedro
River, Arizona. Journal of Geophysical Research: Biogeosciences 112:G03S03.
Meschiatti, A. J., M. S. Arcifa, and N. Fenerich-Verani. 2000. Fish communities associated with
macrophytes in Brazilian floodplain lakes. Environmental Biology of Fishes 58:133-143.
Meyer, J. L, D. L. Strayer, J. B. Wallace, S. L. Eggert, G. S. Helfman, and N. E. Leonard. 2007. The
contribution of headwater streams to biodiversity in river networks. Journal of the American Water
Resources Association 43:86-103.
Milam, J. C., andS. M. Melvin. 2001. Density, habitat use, movements, and conservation of spotted turtles
[Clemmysguttutd) in Massachusetts. Journal of Herpetology 35:418-427.
Miller, A. M., and S. W. Golladay. 1996. Effects of spates and drying on macroinvertebrate assemblages of
an intermittent and perennial prairie stream. Journal of the North American Benthological Society
15:670-689.
Miller, B. A., W. G. Crumpton, and A. G. van der Valk. 2009. Spatial distribution of historical wetland
classes on the Des Moines Lobe, Iowa. Wetlands 29:1146-1152.
Miller, M. W., and T. D. Nudds. 1996. Prairie landscape change and flooding in the Mississippi River
Valley. Conservation Biology 10:847-853.
Miller, T. K., and L. J. Onesti. 1988. Interregional comparison of alluvial stream channel morphology:
Great Plains versus Central Lowlands. Water Resources Bulletin 24:1207-1217.
Miltner, R. J., D. White, and C. Yoder. 2004. The biotic integrity of streams in urban and suburbanizing
landscapes. Landscape and Urban Planning 69:87-100.
Minckley, W., and G. K. Meffre. 1987. Differential selection by flooding in stream-fish communities of the
arid American Southwest. Pages 93-104 in Community and evolutionary ecology of North American
stream fishes. W. J. Matthews and D. C. Heins, editors. University of Oklahoma Press, Norman, OK.
Mitsch, W. J. 1992. Landscape design and the role of created, restored, and natural riparian wetlands in
controlling nonpoint source pollution. Ecological Engineering 1:27-47.
Mueller, M. H., and A. G. van der Valk. 2002. The potential role of ducks in wetland seed dispersal.
Wetlands 22:170-178.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Mulholland, P.}., C. S. Fellows, J. L. Tank, N. B. Grimm, J. R. Webster, S. K. Hamilton, E. Marti, L. Ashkenas,
W. B. Bowden, W. K. Dodds, W. H. McDowell, M. J. Paul, and B. J. Peterson. 2001. Inter-biome
comparison of factors controlling stream metabolism. Freshwater Biology 46:1503-1517.
Mulhouse, J. M., and S. M. Galatowitsch. 2003. Revegetation of prairie pothole wetlands in the mid-
continental US: Twelve years post-reflooding. Plant Ecology 169:143-159.
Murdock, J. N., K. B. Gido, W. K. Dodds, K. N. Bertrand, and M. R. Whiles. 2010. Consumer return
chronology alters recovery trajectory of stream ecosystem structure and function following drought.
Ecology 91:1048-1062.
Murkin, H. R., and P. J. Caldwell. 2000. Avian use of prairie wetlands. Pages 249-286 in Prairie wetland
ecology: The contribution of the marsh ecology research program. H. R. Murkin, A. G. van der Valk,
and W. R. Clark, editors. Iowa State University Press, Ames, IA.
Naiman, R. J., H. Decamps, and M. E. McClain. 2005. Riparia: Ecology, conservation, and management of
streamside communities. Elsevier Academic Press, Burlington, MA.
Nanson, G. C., and J. C. Croke. 1992. A genetic classification of floodplains. Geomorphology 4:459-486.
National Research Council. 2002. Riparian areas: Functions and strategies for management. The National
Academies Press, Washington, DC.
Negrel, P., E. Petelet-Giraud, J. Barbier, and E. Gautier. 2003. Surface water-groundwater interactions in
an alluvial plain: Chemical and isotopic systematics. Journal of Hydrology 277:248-267.
Newman, M. C., and J. F. Schalles. 1990. The water chemistry of Carolina bays: A regional survey. Archiv
fur Hydrobiologie 118:147-168.
NHD. 2008. National Hydrography Dataset. U.S. Geological Survey.
O'Connor, B. L., Y. Hamada, E. E. Bowen, M. A. Grippo, H. M. Hartmann, T. L. Patton, R. A. Van Lonkhuyzen,
and A. E. Carr. 2014. Quantifying the sensitivity of ephemeral streams to land disturbance activities
in arid ecosystems at the watershed scale. Environmental Monitoring and Assessment 186:7075-
7095.
Obolewski, K., K. Glinska-Lewczuk, and S. Kobus. 2009. Effect of hydrological connectivity on the
molluscan community structure in oxbow lakes of the Lyna River. Oceanological and
Hydrobiological Studies 38:75-88.
Ohmart, R. D. 1995. Historical and present impacts of livestock grazing on fish and wildlife resources in
western riparian habitats. Pages 245-279 in Rangeland wildlife. P. R. Krausman, editor. The Society
for Range Management, Denver, CO.
Osterkamp, W. R., and J. E. Costa. 1987. Changes accompanying an extraordinary flood on a sand-bed
stream. Pages 201-224 in Catastrophic flooding. L. Mayer and D. Nash, editors. Allen & Unwin,
Boston, MA.
Osterkamp, W. R., and J. M. Friedman. 2000. The disparity between extreme rainfall events and rare
floods—with emphasis on the semi-arid American west. Hydrological Processes 14:2817-2829.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Ostrand, K. G., and D. E. Marks. 2000. Mortality of prairie stream fishes confined in an isolated pool.
Texas Journal of Science 52:255-258.
Ostrand, K. G., and G. R. Wilde. 2004. Changes in prairie stream fish assemblages restricted to isolated
streambed pools. Transactions of the American Fisheries Society 133:1329-1338.
Parker, G., C. Paola, K. X. Whipple, and D. Mohrig. 1998. Alluvial fans formed by channelized fluvial and
sheet flow. I: Theory. Journal of Hydraulic Engineering 124:985-995.
Pattenden, R. K., and D. A. Boag. 1989. Skewed sex ratio in a northern wintering population of Mallards.
Canadian Journal of Zoology 67:1084-1087.
Pavelis, G. A. 1987. Farm drainage in the United States: History, status, and prospects. USDA
Miscellaneous Publication 1455, U.S. Department of Agriculture, Economic Research Service,
Washington, DC.
Pease, A. A., J. J. Davis, M. S. Edwards, and T. F. Turner. 2006. Habitat and resource use by larval and
juvenile fishes in an arid-land river (Rio Grande, New Mexico). Freshwater Biology 51:475-486.
Perkin, J. S., and K. B. Gido. 2011. Stream fragmentation thresholds for a reproductive guild of Great
Plains fishes. Fisheries 36:371-383.
Peterson, J. T., and C. F. Rabeni. 1996. Natural thermal refugia for temperate warmwater stream fishes.
North American Journal of Fisheries Management 16:738-746.
Phillips, F. M., J. F. Hogan, and B. R. Scanlon. 2004. Introduction and overview, in groundwater recharge
in a desert environment: The southwestern United States. Pages 1-14 in Water science and
applications series. J. F. Hogan, F. M. Phillips, and B. R. Scanlon, editors. American Geophysical
Union., Washington, DC.
Phillips, P. J., J. M. Denver, R. J. Shedlock, and P. A. Hamilton. 1993. Effect of forested wetlands on nitrate
concentrations in ground water and surface water on the Delmarva peninsula. Wetlands 13:75-83.
Phillips, P. J., and R. J. Shedlock. 1993. Hydrology and chemistry of groundwater and seasonal ponds in
the Atlantic coastal-plain in Delaware, USA. Journal of Hydrology 141:157-178.
Platania, S. P., and C. S. Altenbach. 1998. Reproductive strategies and egg types of seven Rio Grande
basin cyprinids. Copeia 1998:559-569.
Plummer, L. N., S. K. Anderholm, W. E. Sanford, and E. Busenberg. 2004. Hydrochemical tracers in the
middle Rio Grande basin, USA: 1. Conceptualization of groundwater flow. Hydrogeology Journal
12:359-388.
Poff, N. L. 1996. A hydrogeography of unregulated streams in the United States and an examination of
scale dependence in some hydrological descriptors. Freshwater Biology 36:71-91.
Poff, N. L., J. D. Allan, M. B. Bain, J. R. Karr, K. L. Prestegaard, B. D. Richter, R. E. Sparks, and J. C.
Stromberg. 1997. The natural flow regime: A paradigm for river conservation and restoration.
BioScience 47:769-784.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Pool, D. R. 2005. Variations in climate and ephemeral channel recharge in southeastern Arizona, United
States. Water Resources Research 41:1-24.
Pool, D. R., and J. E. Dickinson. 2007. Ground-water flow model of the Sierra Vista subwatershed and
Sonoran portions of the Upper San Pedro basin, southeastern Arizona, United States, and Northern
Sonora, Mexico. USGS Scientific Investigations Report 2006-5228, U.S. Department of the Interior,
U.S. Geological Survey prepared in cooperation with the Upper San Pedro Partnership and Bureau of
Land Management, Reston, VA.
Powell, B. F., and R. J. Steidl. 2002. Habitat selection by riparian songbirds breeding in southern Arizona.
The Journal of Wildlife Management:1096-1103.
Powell, D. M., R. Brazier, A. Parsons, J. Wainwright, and M. Nichols. 2007. Sediment transfer and storage
in dryland headwater streams. Geomorphology 88:152-166.
Power, M. E., and A. J. Stewart. 1987. Disturbance and recovery of an algal assemblage following flooding
in an Oklahoma stream. American Midland Naturalist 117:333-345.
Prophet, C. W., and J. D. Ransom. 1974. Summer stream metabolism values for Cedar Creek, Kansas.
Southwestern Naturalist 19:305-308.
Prouty, W. F. 1952. Carolina bays and their origin. Bulletin of the Geological Society of America 63:167-
224.
Pyke, C. R. 2004. Simulating vernal pool hydrologic regimes for two locations in California, USA.
Ecological Modelling 173:109-127.
Pyzoha, J. E., T. J. Callahan, G. Sun, C. C. Trettin, and M. Miwa. 2008. A conceptual hydrologic model for a
forested Carolina bay depressional wetland on the Coastal Plain of South Carolina, USA. Hydrological
Processes 22:2689-2698.
Rabeni, C. F. 1996. Prairie legacies - fish and aquatic resources. Pages 111-124 in Prairie conservation:
Preserving North America's most endangered ecosystem. F. B. Samson and F. L. Knopf, editors.
Island Press, Washington, DC.
Rains, M. C., R. A. Dahlgren, G. E. Fogg, T. Harter, and R. J. Williamson. 2008. Geological control of physical
and chemical hydrology in California vernal pools. Wetlands 28:347-362.
Rains, M. C., G. E. Fogg, T. Harter, R. A. Dahlgren, and R. J. Williamson. 2006. The role of perched aquifers
in hydrological connectivity and biogeochemical processes in vernal pool landscapes, Central Valley,
California. Hydrological Processes 20:1157-1175.
Reckendorfer, W., C. Baranyi, A. Funk, and F. Schiemer. 2006. Floodplain restoration by reinforcing
hydrological connectivity: Expected effects on aquatic mollusc communities. Journal of Applied
Ecology 43:474-484.
Richardson, C. J., and J. W. Gibbons. 1993. Pocosins, Carolina bays, and mountain bogs. Pages 257-310 in
Biodiversity of the southeastern United States. W. H. Martin, S. G. Boyce, and A. C. Echternacht,
editors. John Wiley & Sons, New York, NY.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Rinne, J. N., and D. Miller. 2006. Hydrology, geomorphology and management: Implications for
sustainability of native southwestern fishes. Reviews in Fisheries Science 14:91-110.
Roach, W. J., J. B. Heffernan, N. B. Grimm, J. R. Arrowsmith, C. Eisinger, and T. Rychener. 2008.
Unintended consequences of urbanization for aquatic ecosystems: A case study from the Arizona
desert. BioScience 58:715-727.
Rosenberry, D. 0., and T. C. Winter. 1997. Dynamics of water-table fluctuations in an upland between
two prairie-pothole wetlands in North Dakota. Journal of Hydrology 191:266-289.
Ross, T. E. 1987. A comprehensive bibleography of the Carolina bays literature. Journal of Elisha Mitchell
Scientific Society 103:28-42.
Routman, E. 1993. Population structure and genetic diversity of metamorphic and paedomorphic
populations of the tiger salamander, Ambystoma tigrinum. Journal of Evolutionary Biology 6:329-
357.
Royer, T. V., M. B. David, and L. E. Gentry. 2006. Timing of riverine export of nitrate and phosphorus from
agricultural watersheds in Illinois: Implications for reducing nutrient loading to the Mississippi
River. Environmental Science & Technology 40:4126-4131.
Rulifson, R. A., and B. L. Wall. 2006. Fish and blue crab passage through water control structures of a
coastal bay lake. North American Journal of Fisheries Management 26:317-326.
Samson, F., and F. Knopf. 1994. Prairie conservation in North America. BioScience 44:418-421.
Savage, H., Jr. 1982. The mysterious Carolina bays. University of South Carolina Press, Columbia, SC.
Sawin, R. S., R. C. Buchanan, and W. Lebsack. 1999. Flint Hills springs. Transactions of the Kansas
Academy of Science 102:1-31.
Scanlon, B. R., C. C. Faunt, L. Longuevergne, R. C. Reedy, W. M. Alley, V. L. McGuire, and P. B. McMahon.
2012. Groundwater depletion and sustainability of irrigation in the US High Plains and Central
Valley. Proceedings of the National Academy of Sciences 109:9320-9325.
Schalles, J. F., and D. J. Shure. 1989. Hydrology, community structure, and productivity patterns of a
dystrophic Carolina bay wetland. Ecological Monographs 59:365-385.
Schlosser, I. J. 1987. A conceptual framework for fish communities in small warmwater streams. Pages
17-24 in Community and evolutionary ecology of North American stream fishes. W. J. Matthews and
D. C. Heins, editors. University of Oklahoma Press, Norman, OK.
Schlosser, I. J. 1991. Stream fish ecology: A landscape perspective. BioScience:704-712.
Schoof, R. R., J. W. Naney, and W. M. Boxley. 1978. Hydrologic effects of the Tonkawa Creek flood
abatement program. Water Resources Bulletin 14:629-639.
Schrank, S. J., C. S. Guy, M. R. Whiles, and B. L. Brock. 2001. Influence of instream and landscape-level
factors on the distribution of Topeka shiners Notropis topeka in Kansas streams. Copeia 2001:413-
421.
Schueler, T. R. 1994. The importance of imperviousness. Watershed protection techniques 1:100-111.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Schumm, S. A., and R. F. Hadley. 1957. Arroyos and the semiarid cycle of erosion [Wyoming and New
Mexico]. American Journal of Science 255:161-174.
Seager, R., M. Ting, I. Held, Y. Kushnir, J. Lu, G. Vecchi, H.-P. Huang, N. Harnik, A. Leetmaa, and N.-C. Lau.
2007. Model projections of an imminent transition to a more arid climate in southwestern North
America. Science 316:1181-1184.
Semlitsch, R. D., and J. R. Bodie. 1998. Are small, isolated wetlands expendable? Conservation Biology
12:1129-1133.
Sharitz, R. R. 2003. Carolina bay wetlands: Unique habitats of the southeastern United States. Wetlands
23:550-562.
Sharitz, R. R., and J. W. Gibbons. 1982. The ecology of southeastern shrub bogs (pocosins) and Carolina
bays: A community profile. FWS/OBS-82/04, U.S. Department of the Interior, U.S. Fish and Wildlife
Services Program, Washington, DC.
Shaw, J. R., and D. J. Cooper. 2008. Linkages among watersheds, stream reaches, and riparian vegetation
in dryland ephemeral stream networks. Journal of Hydrology 350:68-82.
Shoup, D. E., and D. H. Wahl. 2009. Fish diversity and abundance in relation to interannual and lake-
specific variation in abiotic characteristics of floodplain lakes of the lower Kaskaskia River, Illinois.
Transactions of the American Fisheries Society 138:1076-1092.
Smith, C. L, and C. R. Powell. 1971. The summer fish communities of Brier Creek, Marshall County,
Oklahoma. American Museum Novitates 2458:1-30.
Smith, D. W., and W. L. Verrill. 1998. Vernal pool-soil-landform relationships in the Central Valley,
California. Pages 15-23 in Ecology, conservation, and management of vernal pool ecosystems—
proceedings from a 1996 conference. C. W. Witham, E. T. Bauder, D. Belk, W. R. Ferren, Jr., and R.
Ornduff, editors. California Native Plant Society, Sacramento, CA.
Smith, S. V, W. H. Renwick, J. D. Bartley, and R. W. Buddemeier. 2002. Distribution and significance of
small, artificial water bodies across the United States landscape. The Science of the Total
Environment 299:21-36.
Snodgrass, J. W., A. L. Bryan, and J. Burger. 2000a. Development of expectations of larval amphibian
assemblage structure in southeastern depression wetlands. Ecological Applications 10:1219-1229.
Snodgrass, J. W., A. L. Bryan, R. F. Lide, and G. M. Smith. 1996. Factors affecting the occurrence and
structure offish assemblages in isolated wetlands of the upper Coastal Plain, USA. Canadian Journal
of Fisheries and Aquatic Sciences 53:443-454.
Snodgrass, J. W., C. H. Jagoe, A. L. Bryan, H. A. Brant, and J. Burger. 2000b. Effects of trophic status and
wetland morphology, hydroperiod, and water chemistry on mercury concentrations in fish.
Canadian Journal of Fisheries and Aquatic Sciences 57:171-180.
Soons, M. B. 2006. Wind dispersal in freshwater wetlands: Knowledge for conservation and restoration.
Applied Vegetation Science 9:271-278.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Sophocleous, M. 2005. Groundwater recharge and sustainability in the High Plains aquifer in Kansas,
USA. Hydrogeology Journal 13:351-365.
Sophocleous, M. 2010. Review: Groundwater management practices, challenges, and innovations in the
High Plains aquifer, USA—lessons and recommended actions. Hydrogeology Journal 18:559-575.
Stagliano, D. M., and M. R. Whiles. 2002. Macroinvertebrate production and trophic structure in a
tallgrass prairie headwater stream. Journal of the North American Benthological Society 21:97-113.
Stanford, J. A., and J. V. Ward. 1988. The hyporheic habitat of river ecosystems. Nature 335:64-66.
Stanley, E. H., S. G. Fisher, and N. B. Grimm. 1997. Ecosystem expansion and contraction in streams.
BioScience 47:427-435.
Stolt, M. H., and M. C. Rabenhorst. 1987a. Carolina bays on the eastern shore of Maryland: 1. Soil
characterization and classification. Soil Science Society of America Journal 51:394-398.
Stolt, M. H., and M. C. Rabenhorst. 1987b. Carolina bays on the eastern shore of Maryland: 2. Distribution
and origin. Soil Science Society of America Journal 51:399-405.
Strand, M., and R. W. Merritt. 1999. Impacts of livestock grazing activities on stream insect communities
and the riverine environment. American Entomologist 45:13-29.
Stromberg, J., R. Tiller, and B. Richter. 1996. Effects of groundwater decline on riparian vegetation of
semiarid regions: the San Pedro, Arizona. Ecological Applications:113-131.
Stromberg, J. C., K. J. Bagstad, J. M. Leenhouts, S. J. Lite, and E. Makings. 2005. Effects of stream flow
intermittency on riparian vegetation of a semiarid region river (San Pedro River, Arizona). River
Research and Applications 21:925-938.
Stromberg, J. C., and B. J. Tellman. 2009. Ecology and conservation of the San Pedro River. University of
Arizona Press, Tucson, AZ.
Sun, G., T. J. Callahan, J. E. Pyzoha, and C. C. Trettin. 2006. Modeling the climatic and subsurface
stratigraphy controls on the hydrology of a Carolina bay wetland in South Carolina, USA. Wetlands
26:567-580.
Sutter, R. D., and R. Krai. 1994. The ecology, status, and conservation of two non-alluvial wetland
communities in the south Atlantic and eastern Gulf Coastal Plain, USA. Biological Conservation
68:235-243.
Swanson, C. D., and R. W. Bachmann. 1976. Model of algal exports in some Iowa streams. Ecology
57:1076-1080.
Swanson, G. A., T. C. Winter, V. A. Adomaitis, and J. W. LaBaugh. 1988. Chemical characteristics of prairie
lakes in south-central North Dakota—their potential for impacting fish and wildlife. U.S. Fish and
Wildlife Technical Report 18, U.S. Department of the Interior, U.S. Fish and Wildlife Service,
Washington, DC.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Tang, C., I. Machida, S. Shindo, A. Kondoh, and Y. Sakura. 2001. Chemical and isotopic methods for
confirming the roles of wadis in regional groundwater recharge in a regional arid environment: A
case study in Al Ain, UAE. Hydrological Processes 15:2195-2202.
Tate, C. M. 1990. Patterns and controls of nitrogen in tallgrass prairie streams. Ecology 71:2007-2018.
Taylor, C. M., M. R. Winston, and W. J. Matthews. 1993. Fish species-environment and abundance
relationships in a Great Plains river system. Ecography 16:16-23.
Thorp, J. H., and A. D. Belong. 2002. Dominance of autochthonous autotrophic carbon in food webs of
heterotrophic rivers. Oikos 96:543-550.
Tiner, R. W. 2003. Geographically isolated wetlands of the United States. Wetlands 23:494-516.
Townsend-Small, A., D. E. Pataki, H. Liu, Z. Li, Q. Wu, and B. Thomas. 2013. Increasing summer river
discharge in southern California, USA, linked to urbanization. Geophysical Research Letters
40:4643-4647.
Transeau, E. N. 1905. Forest centers of eastern America. American Naturalist 39:875-889.
Transeau, E. N. 1935. The prairie peninsula. Ecology 16:423-437.
Turner, D. S., and M. D. List. 2007. Habitat mapping and conservation analysis to identify critical streams
for Arizona's native fish. Aquatic Conservation: Marine and Freshwater Ecosystems 17:737-748.
Turner, D. S., and H. E. Richter. 2011. Wet/dry mapping: Using citizen scientists to monitor the extent of
perennial surface flow in dryland regions. Environmental Management 47:497-505.
Turner, K. W., B. B. Wolfe, and T. W. D. Edwards. 2010. Characterizing the role of hydrological processes
on lake water balances in the Old Crow Flats, Yukon Territory, Canada, using water isotope tracers.
Journal of Hydrology 386:103-117.
U.S. EPA. 2006. Wadeable streams assessment: A collaborative survey of the nation's streams. EPA 841-
B-06-002, U.S. Environmental Protection Agency, Office of Water, Washington, DC.
USDA. 2002. Management and techniques for riparian restoration, roads field guide, Volume 1. General
Technical Report RMRS-GTR-1-2, U.S. Department of Agriculture, U.S. Forest Service, Rocky
Mountain Research Station, Fort Collins, CO.
USGS. 2006. National Hydrography Dataset.
Van De Meutter, F., L. De Meester, and R. Stoks. 2007. Metacommunity structure of pond
macroinvertebrates: Effects of dispersal mode and generation time. Ecology 88:1687-1695.
van der Kamp, G., and M. Hayashi. 1998. The groundwater recharge function of small wetlands in the
semi-arid northern prairies. Great Plains Research 8:39-56.
van der Kamp, G., and M. Hayashi. 2009. Groundwater-wetland ecosystem interaction in the semiarid
glaciated plains of North America. Hydrogeology Journal 17:203-214.
van der Valk, A. G. 2006. The biology of freshwater wetlands. Oxford University Press, New York, NY.
van der Valk, A. G., and C. B. Davis. 1978. The role of seed banks in the vegetation dynamics of prairie
glacial marshes. Ecology 59:322-335.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
van der Valk, A. G., and R. L. Pederson. 2003. The SWANCC decision and its implications for prairie
potholes. Wetlands 23:590-596.
Van Devender, T. R. 2002. The Sonoran desert tortoise: Natural history, biology, and conservation.
University of Arizona Press, Tucson, AZ.
van Digglen, R. 2006. Landscape: Spatial interactions. Pages 31-44 in Restoration Ecology. J. Andel and J.
Aronson, editors. Blackwell, Oxford, UK.
Van Haveren, B. P. 1986. Management of instream flows through runoff detention and retention. Water
Resources Bulletin 22:399-404.
Van Riper, C., and K. L. Cole. 2004. The Colorado Plateau: Cultural, biological, and physical research.
University of Arizona Press.
Vannote, R. L., G. W. Minshall, K. W. Cummins, J. R. Sedell, and C. E. Gushing. 1980. The river continuum
concept. Canadian Journal of Fisheries and Aquatic Sciences 37:130-137.
Vanschoenwinkel, B., S. Gielen, M. Seaman, and L. Brendonck. 2009. Wind mediated dispersal of
freshwater invertebrates in a rock pool metacommunity: Differences in dispersal capacities and
modes. Hydrobiologia 635:363-372.
Vining, K. C. 2002. Simulation of streamflow and wetland storage, Starkweather Coulee subbasin, North
Dakota. Water years 1981-98. U.S.G.S. Water-Resources Investigations Report 02-4113, U.S.
Department of the Interior, U.S. Geological Survey in cooperation with the North Dakota State Water
Commission, Bismarck, ND.
Vining, K. C. 2004. Simulation of runoff and wetland storage in the Hamden and Lonetree watershed
sites within the Red River of the North basin, North Dakota and Minnesota. USGS Scientific
Investigations Report 2004-5168, U.S. Department of the Interior, U.S. Geological Survey, Reston, VA.
Vivoni, E., R. S. Bowman, R. L. Wyckoff, R. T. Jakubowski, and K. E. Richards. 2006. Analysis of a monsoon
flood event in an ephemeral tributary and its downstream hydrologic effects. Water Resources
Research 42:W03404.
Wagner, F. H. 1978. Livestock grazing and the livestock industry. Pages 121-145 in Wildlife and America.
H. P. Brokaw, editor. Council on Environmental Quality, U.S. Government Printing Office,
Washington, DC.
Wahi, A. K., J. F. Hogan, B. Ekwurzel, M. N. Baillie, and C. J. Eastoe. 2008. Geochemical quantification of
semiarid mountain recharge. Ground Water 46:414-425.
Weakley, A. S., and M. P. Schafale. 1991. Classification of pocosins of the Carolina Coastal Plain. Wetlands
11:355-375.
Webb, R. H., and J. L. Betancourt. 1992. Climatic variability and flood frequency of the Santa Cruz River,
Pima County, Arizona. USGS Water-Supply Paper 2379, U.S. Department of the Interior, U.S.
Geological Survey, Reston, VA.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Weitkamp, W. A., R. C. Graham, M. A. Anderson, and C. Amrhein. 1996. Pedogenesis of a vernal pool
entisol-alfisol-vertisol catena in southern California. Soil Science Society of America Journal 60:316-
323.
Wells, B. W., and S. G. Boyce. 1953. Carolina bays: Additional data on their origin, age and history. Journal
of Elisha Mitchell Scientific Society 69:119-141.
Whigham, D. F., and T. E. Jordan. 2003. Isolated wetlands and water quality. Wetlands 23:541-549.
Whiles, M. R., and W. K. Dodds. 2002. Relationships between stream size, suspended particles, and filter-
feeding macroinvertebrates in a Great Plains drainage network. Journal of Environmental Quality
31:1589-1600.
Whiting, D. P., M. R. Whiles, and M. L. Stone. 2011. Patterns of macroinvertebrate production, trophic
structure, and energy flow along a tallgrass prairie stream continuum. Limnology and Oceanography
56:887-898.
Wiley, M. J., L. L. Osborne, and R. W. Larimore. 1990. Longitudinal structure of an agricultural prairie
river system and its relationship to current stream ecosystem theory. Canadian Journal of Fisheries
and Aquatic Sciences 47:373-384.
Williams, D. D. 1996. Environmental constraints in temporary fresh waters and their consequences for
the insect fauna. Journal of the North American Benthological Society 15:634-650.
Williams, D. D. 2005. The biology of temporary waters. Oxford University Press, Oxford, UK.
Williams, G. P., and M. G. Wolman. 1984. Downstream effects of dams on alluvial rivers. USGS
Professional Paper 1286, U.S. Department of the Interior, U.S. Geological Survey, Washington, DC.
Wilson, J. L., and H. Guan. 2004. Mountain-block hydrology and mountain-front recharge. Pages 113-137
in Groundwater recharge in a desert environment: The southwestern United States. F. M. Phillips, J.
Hogan, and B. R. Scanlon, editors. American Geophysical Union, Washington, DC.
Winemiller, K. 0., S. Tarim, D. Shormann, and J. B. Cotner. 2000. Fish assemblage structure in relation to
environmental variation among Brazos River oxbow lakes. Transactions of the American Fisheries
Society 129:451-468.
Winston, M. R., C. M. Taylor, and J. Pigg. 1991. Upstream extirpation of four minnow species due to
damming of a prairie stream. Transactions of the American Fisheries Society 120:98-105.
Winter, T. C. 2007. The role of groundwater in generating streamflow in headwater areas and in
maintaining base flow. Journal of the American Water Resources Association 43:15-2 5.
Winter, T. C., J. W. Harvey, 0. L. Franke, and W. M. Alley. 1998. Ground water and surface water: A single
resource. USGS Circular 1139, U.S. Department of the Interior, U.S. Geological Survey, Denver, CO.
Winter, T. C., and J. W. LaBaugh. 2003. Hydrologic considerations in defining isolated wetlands. Wetlands
23:532-540.
Winter, T. C., and D. 0. Rosenberry. 1995. The interaction of groundwater with prairie pothole wetlands
in the Cottonwood Lake area, east-central North Dakota, 1979-1990. Wetlands 15:193-211.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
-------
Winter, T. C., and D. 0. Rosenberry. 1998. Hydrology of prairie pothole wetlands during drought and
deluge: A 17-year study of the cottonwood lake wetland complex in North Dakota in the perspective
of longer term measured and proxy hydrological records. Climatic Change 40:189-209.
Wohl, E., D. Egenhoff, and K. Larkin. 2009. Vanishing riverscapes: A review of historical channel change
on the western Great Plains. Pages 131-142 in Management and restoration of fluvial systems with
broad historical changes and human impacts. Geological Society of America Special Paper 451. L. A.
James, S. L. Rathburn, and G. R. Whittecar, editors. Geological Society of America, Boulder, CO.
Wolock, D. M., T. C. Winter, and G. McMahon. 2004. Delineation and evaluation of hydrologic-landscape
regions in the United States using geographic information system tools and multivariate statistical
analysis. Environmental Management 34:S71-S88.
Woolhiser, D. A., T. 0. Reefer, and K. T. Redmond. 1993. Southern Oscillation effects on daily
precipitation in the southwestern United-States. Water Resources Research 29:1287-1295.
Yuan, F., and S. Miyamoto. 2008. Characteristics of oxygen-18 and deuterium composition in waters
from the Pecos River in American Southwest. Chemical Geology 255:220-230.
Zale, A. V., D. M. Leslie Jr., W. L. Fisher, and S. G. Merrifield. 1989. The physicochemistry, flora, and fauna
of intermittent prairie streams: A review of the literature. U.S. Fish and Wildlife Service Biological
Report 89(5), U.S. Department of the Interior, U.S. Fish and Wildlife Service, Washington, DC.
Zedler, P. H. 1987. The ecology of southern California vernal pools: A community profile. U.S. Fish and
Wildlife Service Biological Report 85(7.11), U.S. Department of the Interior, U.S. Fish and Wildlife
Service, Washington, DC.
Zedler, P. H. 2003. Vernal pools and the concept of "isolated wetlands". Wetlands 23:597-607.
Zeug, S. C., D. Peretti, and K. 0. Winemiller. 2009. Movement into floodplain habitats by gizzard shad
(Dorosoma cepedianum] revealed by dietary and stable isotope analyses. Environmental Biology of
Fishes 84:307-314.
Zeug, S. C., and K. 0. Winemiller. 2008. Relationships between hydrology, spatial heterogeneity, and fish
recruitment dynamics in a temperate floodplain river. River Research and Applications 24:90-102.
Zeug, S. C., K. 0. Winemiller, and S. Tarim. 2005. Response of Brazos River oxbow fish assemblages to
patterns of hydrologic connectivity and environmental variability. Transactions of the American
Fisheries Society 134:1389-1399.
Zimmer, K. D., M. A. Hanson, and M. G. Butler. 2001. Effects of fathead minnow colonization and removal
on a prairie wetland ecosystem. Ecosystems 4:346-357.
Stream and Wetland Connectivity: January 2015
A Review and Synthesis 3
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