Edward Gage and David J. Cooper

United States Department of Agriculture / Forest Service

Rocky Mountain Research Station

General Technical Report RMRS-GTR-286WWW

April 2013

Historical Range of Variation Assessment 

for Wetland and Riparian Ecosystems, 

U.S. Forest Service Rocky Mountain Region
Gage, Edward; Cooper, David J. 2013. Historical range of variation assessment for wetland 

and riparian ecosystems, U.S. Forest Service Rocky Mountain Region. Gen. Tech. Rep. 

RMRS-GTR-286WWW. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky 

Mountain Research Station. 239 p.

Abstract

This document provides an overview of historical range of variation concepts and explores 

their application to wetland and riparian ecosystems in the US Forest Service Rocky Mountain 

Region (Region 2), which includes National Forests and National Grasslands occurring in the 

states of Colorado, Wyoming, Nebraska, Kansas, and South Dakota. For each of five ecosystem 

types—riparian areas, fens, wet meadows, salt flats, and marshes—we review key structural and 

functional characteristics including geomorphic setting, principal ecological drivers, classification, 

and dominant vegetation. In addition, we discuss anthropogenic factors known to influence the 

abundance or condition of each main wetland type. 

Keywords: wetlands, wet meadows, fens, salt flats, riparian, marshes, Colorado, Wyoming, South 

Dakota, Kansas, Nebraska

Acknowledgments

We thank the numerous individuals who provided us information, data, or critical review of this 

document including Barry Johnston, Bernie Bornong, Carl Chambers, Dave Bradford, Deanna 

Reyher, Dennis Lowry, Gay Austin, Greg Schenbeck, Jeff Redder, Jim Maxwell, Joe Neal, Joe 

Rocchio, Kathy Carsey, Ken Kanaan, Les Dobson, Les Gonyer, Michele Girard, Ray Zubick, Dave 

Ode, Gerry Steinauer, George Jones, Kelly Kindscher, Dean Erhard, Mike Johnson, Joanna Lemly, 

Kathy Roche, Rod Chimner, John Proctor, and Mike Menaffe. Claudia Regan, Jim Maxwell, and 

Dan Binkley were particularly helpful in seeing this project through from start to end. Special thanks 

to Kate Dwire and Melinda Larson for assistance in final revisions and facilitating publication. We 

also thank Kathy Roche, Gay Austin, and five anonymous reviewers from the Ecological Society 

of America for providing insightful comments on draft versions of this document.
Historical Range of Variation Assessment for Wetland And 

Riparian Ecosystems, U.S. Forest Service Rocky Mountain Region

Edward Gage and David J. Cooper

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

Objectives of the Project . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1

Organization of This Report . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

Historical Range of Variation Concepts . . . . . . . . . . . . . . . . . . . . . . . . 5

Wetland Definitions and Applicable Laws . . . . . . . . . . . . . . . . . . . . . . 5

2. Data and Information Sources . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

Instrumental Records . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8

Paleoclimatic, Paleohydrologic, and Paleoecological Studies . . . . 10

Classification of Region 2 Wetlands and Riparian Communities . . 11

Historical Photographs and Accounts . . . . . . . . . . . . . . . . . . . . . . . . 12

3. Wetland and Riparian Ecosystems of Region 2: Structural 

and Functional Characteristics and Ecological and 

Historical Setting . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

Overview of the Principal Wetland and Riparian Types in 

the Region . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 14

Vegetation of Region 2 Wetlands and Riparian Areas . . . . . . . . . . . 17

Climate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 19

Geomorphology and Geology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26

Hydrologic Drivers of Wetland and Riparian Ecosystem 

Structure and Function . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 28

Disturbance Regimes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 30

Ecoregions of Region 2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31

Human History of the Region . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 31

Overview of Anthropogenic Impacts to Wetlands and 

Riparian Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 38

4. Historical Range of Variation for Region 2 Riparian Ecosystems . . . 49

Definitions and Concepts of Riparian . . . . . . . . . . . . . . . . . . . . . . . . . 49

Geomorphic Setting and Principal Ecological Drivers . . . . . . . . . . . 49

Riparian Classification and Gradients . . . . . . . . . . . . . . . . . . . . . . . . 57

HRV of Riparian Ecosystems in Region 2 . . . . . . . . . . . . . . . . . . . . . 64

Management Opportunities and Constraints . . . . . . . . . . . . . . . . . . . 90

5. Historical Range of Variation for Region 2 Marshes . . . . . . . . . . . . . . 95

Definitions and Concepts of Marshes . . . . . . . . . . . . . . . . . . . . . . . . 95

Geomorphic Setting and Principal Ecological Drivers . . . . . . . . . . . 95

Marsh Classification and Gradients . . . . . . . . . . . . . . . . . . . . . . . . . 103

Marsh Vegetation in the Region . . . . . . . . . . . . . . . . . . . . . . . . . . . . 106

HRV of Marshes in Region 2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 108

Management Opportunities and Constraints . . . . . . . . . . . . . . . . . . 117
6. Historical Range of Variation for Region 2 Fens . . . . . . . . . . . . . . . . 120

Definitions and Concepts of Fens . . . . . . . . . . . . . . . . . . . . . . . . . . . 120

Geomorphic Setting and Principal Ecological Drivers . . . . . . . . . . 121

Fen Classification and Gradients . . . . . . . . . . . . . . . . . . . . . . . . . . . 138

Fen Vegetation in the Region . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 142

HRV of Fens in Region 2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 146

Management Opportunities and Constraints . . . . . . . . . . . . . . . . . . 161

7. Historical Range of Variation for Region 2 Salt Flats . . . . . . . . . . . . 166

Definitions and Concepts of Salt Flats . . . . . . . . . . . . . . . . . . . . . . . 166

Geomorphic Setting and Principal Ecological Drivers . . . . . . . . . . 166

Salt Flat Classification and Gradients . . . . . . . . . . . . . . . . . . . . . . . 174

Salt Flat Vegetation in the Region . . . . . . . . . . . . . . . . . . . . . . . . . . . 176

HRV of Salt Flats in Region 2 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 178

Management Opportunities and Constraints . . . . . . . . . . . . . . . . . . 181

8. Historical Range of Variation for Region 2 Wet Meadows . . . . . . . . . 184

Definitions and Concepts of Wet Meadows . . . . . . . . . . . . . . . . . . . 184

Geomorphic Setting and PrincipalEcological Drivers . . . . . . . . . . 184

Wet Meadow Classification and Gradients . . . . . . . . . . . . . . . . . . . 190

Wet Meadow Vegetation in the Region . . . . . . . . . . . . . . . . . . . . . . . 192

HRV of Wet Meadows in Region 2 . . . . . . . . . . . . . . . . . . . . . . . . . . . 195

9. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 203

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 205
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 1

1. Introduction

Objectives of the Project________________________

Although they comprise only a small percentage of the region’s land area, wetland and riparian ecosystems are critical components of Rocky Mountain and 

Great Plains landscapes. These ecologically diverse ecosystems are found at all 

elevations and latitudes and provide a number of important economic and ecological functions (Gregory and others 1991; Mitsch and Gosselink 2007; Patten 

1998). Critical as wildlife habitat (Brown and others 1996; Davidson and Knight 

2001; Haukos 1992; Nelson and others 1984) and as local and regional centers 

of biodiversity (Naiman and others 1993; Nilsson and Svedmark 2002; Pollack 

1998), these ecosystems support many biogeochemical, physical, and ecological 

processes not found elsewhere on western landscapes.

As wetland and riparian functions and values have become better understood, 

a variety of laws and regulations have been promulgated that are aimed at promoting wetland and riparian conservation. However, wetland and riparian areas 

have historically been heavily impacted by anthropogenic activities throughout 

North America (Brinson and Malvarez 2002; Graf 1999; Office of Technology 

Assessment 1984; Shafroth and others 1998; Tiner 1984). Demand for water, 

fertile land, and forage for livestock in the arid and semi-arid West has already 

affected many aquatic, riparian, and wetland areas; and pressures will likely increase with time, threatening the integrity and long-term viability of these vital 

ecosystems and the biota they support (Baron and others 2002; Pringle 2000).

As a principal land manager in the region, the U.S. Department of Agriculture, 

Forest Service (USFS) sits at the center of this conflict. Charged with simultaneously providing natural resources goods and services and conserving biodiversity 

and ecosystem integrity, the USFS must balance these often conflicting goals—a 

task that is made more difficult by the complexity of the social, economic, and 

ecological systems involved (Kaufmann and others 1994; Landres and others 

1999). Recognizing this complexity, managers have increasingly looked to the 

scientific community to help inform and guide management and planning activities (Dale and others 2000; Thomas 1996).

Over the last several decades, a paradigm shift has occurred in the field of 

ecology, from a focus on stability and equilibrium, to an emphasis on concepts 

of variability, change, and dynamism (Holling 1996; Holling and Meffe 1996). 

Scientists have increasingly stressed the importance of disturbances to ecological systems (Turner and others 1996, 1998), and in response, managers have 

begun to apply such concepts to both individual species conservation efforts 

and broad-scale land management (Dale and others 2000). New conceptual and 

quantitative tools have been developed to relate ecological pattern to process at 
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Figure 1. National Forests and Grasslands in the U.S. Forest Service Rocky Mountain Region.

different spatial and temporal scales (Turner 1989), often with an explicit focus 

on system variability. One concept that is increasingly relied upon to guide land 

management activities is the historical range of variability (HRV). Our primary 

objective in this assessment is to apply this concept, defined and described below, to riparian and wetland ecosystems of the USFS Rocky Mountain Region 

(Region 2).

Specifically, our goals are to: (1) outline the types of wetland and riparian ecosystems that occur on Region 2 National Forests and Grasslands; (2) identify and 

characterize the key variables driving vegetation patterns and functional attributes 

of wetlands and riparian systems at a range of spatial and temporal scales; and 

(3) use available data and information to make inferences regarding the HRV of 

key variables influencing wetlands and riparian areas prior to the mid-nineteenth 

Century when extensive Euro-American settlement in the region began. Unlike 

most previous HRV assessments prepared for Region 2 (e.g., Dillon and others 

2005; Meyer and others 2003; Veblen and Donnegan 2005), the broad geographic 

scope of this assessment which encompasses 17 National Forests and 7 National 

Grasslands in 5 states (Figure 1) and high ecological and functional diversity 

of ecosystems in the assessment area preclude a fine-grained, spatially explicit 

analysis for each Forest and Grassland. Rather, our intent is (1) to provide a 

broad conceptual framework based on the best available scientific information 
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Table 1—Nested hierarchy of key spatial and temporal scales, ecological drivers, and ecological response variables relevant to the 

structure and function of a hypothetical wetland. Our assessment is primarily focused on drivers and response variables.

Scale

Level Time (yr) Space (m2) Key ecological drivers Potential response variables

Leaf 0.001 to 1 0.01 to 0.1 Photosynthetically active Respiration photosynthesis

radiation; osmotic pressure; decomposition

tissue nutrient concentrations

Plant 0.1 to 100 0.1 to 10 Soil water availability; Plant productivity; root, shoot,

nutrient concentrations; and propagule production

competition; herbivory

Wetland 1 to 1000 1 to 1000 Water table depth; local Species diversity and

hydrology; herbivory productivity; decomposition;

N mineralization; peat 

accumulation rate

Watershed 100 to 10,000 1000 to 10,000 Mass wasting; bedrock Distribution and abundance of

porosity; glacial history; different wetland types

fire history; beaver; 

stream discharge;

sediment production

Ecoregion > 10,000 > 100,000 Lithology; tectonics; Adaptation; extinction;

Milankovitch cycles species radiation

from which managers or researchers can develop reasonable assessments of the 

HRV of wetlands or riparian systems in particular planning areas and (2) to provide a starting point for further site-specific research. There are many potential 

variables we could focus on, each operating at different spatial and temporal 

scales (Table 1), but our general emphasis is on variables operating at the scale 

of individual wetlands to watersheds.

Previous HRV Assessments in Region 2

Most previous HRV assessments in Region 2 have been prepared for individual 

National Forests and have been limited to upland forested ecosystem types (Dillon 

and others 2005; Kulakowski and Veblen 2006; Meyer and others 2003; Veblen 

and Donnegan 2005). Our overall objectives are similar to these assessments, 

but our focus is on a much broader geographic area and on multiple ecosystem 

types that differ widely in composition, structure, and function. These fundamental differences in scope among assessment efforts require the use of different 

conceptual and analytical approaches.

Even when applied in a qualitative rather than a quantitative statistical context, 

reference to a specific variable of interest is required for the HRV concept to 

have meaning. A variety of ecological attributes were examined in previous HRV 

analyses, spanning multiple spatial and temporal scales (Table 2). While our approach is similar, most of the specific ecological drivers and response variables 
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Table 2—Examples of ecological variables examined in previous HRV assessments (left column) 

and those variables discussed in this assessment (right column). Variables of particular 

importance to specific ecosystem type(s) are indicated in parentheses (M = marsh, WM = 

wet meadow; F = fen; R = riparian; SF = salt flat). Definitions and descriptions of ecosystem 

types are presented later in the document.

Examples of ecosystem variables Potential ecosystem variables of

examined for upland HRV interest in wetland/riparian HRV

Snag density (Dillon et al. 2005) Species composition (M, WM, SF, F, R)

Stand structure (Dillon et al. 2005) Age-class distribution (R, WM)

Tree species and genetic diversity Proportion woody cover (WM, F, R)

(Dillon et al. 2005)

Tree density (Dillon et al. 2005) Woody plant size/age distribution (WM, R)

Density and size of canopy gaps Wetland abundance (M, WM, SF, F, R)

(Dillon et al. 2005)

Density and cover of understory plants Landscape diversity of wetland types

(Dillon et al. 2005) (M, WM, SF, F, R)

Age-class structure of stands Flood frequency, magnitude, & timing

(Dillon et al. 2005) (M, WM, SF, F, R)

Size-class structure of stands Course woody debris (R)

(Dillon et al. 2005)

Course woody debris (Dillon et al. 2005) Beaver abundance and use (R)

Seedling and sapling density (Dillon et al. 2005) Sediment budget (M, WM, SF, F, R)

Fire (Dillon et al. 2005; Veblen and Carbon accumulation/loss dynamics (F)

Donnegen 2003)

Disease (Dillon et al. 2005; Veblen and Hydrologic variability (M, WM, SF, F, R)

Donnegen 2003)

Insects (Dillon et al. 2005; Veblen and Spatial and temporal patterns of woody

Donnegen 2003) plant establishment (WM, R)

Wind (Dillon et al. 2005; Veblen and Ground and surface water chemistry, 

Donnegen 2003) nutrient fluxes (M, WM, SF, F, R) 

Wetland functions (groundwater recharge, 

discharge, biodiversity, storm water 

detention, etc.) (M, WM, SF, F, R)

for wetland and riparian ecosystems differ from those of upland forests. Even 

among wetland types, the most important variables may differ. For instance, shallow groundwater inflow is a key variable driving vegetation dynamics in fens, 

wet meadows, and some marshes (Galatowitsch and van der Valk 1998; Winter 

1999; Winter and Rosenberry 1995), however it is not a critical driver of Great 

Plains riparian communities along ephemeral streams, which are influenced more 

by infrequent floods (Friedman and Lee 2002). Species composition can differ 

almost completely between two nearby fens within the same small watershed, 

but can be identical between cattail (Typha spp.) marshes hundreds of miles 

distant from one another. Variables examined in this and previous assessments 

are contrasted in Table 2 and are discussed in greater detail in chapters specific 

to each wetland type.
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Organization of This Report _____________________

A major challenge in preparing an assessment of this scope is to provide a concise 

yet thorough framework in which to explore the ecological and historical factors shaping the HRV of different ecosystem types. In this chapter, we provide 

an overview of our objectives and an introduction to some of the basic HRV 

concepts and definitions used in earlier assessments and those we have applied 

to wetlands and riparian ecosystems in this document. In Chapter 2, we outline 

the primary resources and analytical approaches we used in assessing the HRV 

for wetland and riparian ecosystems. Many topics, such as broad-scale climatic 

trends and the human history of different parts of the study region are presented 

in Chapter 3, along with a review of wetland and riparian classifications. Subsequent chapters present an HRV assessment for each of the five major wetland 

and riparian ecosystem types that occur in USFS Region 2: riparian ecosystems, 

marshes, fens, salt flats, and wet meadows. In Chapter 8, we provide a synthesis 

of our findings and present suggestions for future research.

Historical Range of Variation Concepts____________

The underlying premise of the HRV approach is that by managing ecosystems 

for the range of processes that have characterized ecosystems in the past, land 

managers can maximize the likelihood of preserving into the future the widest 

compliment of species and important ecological functions (Landres and others 

1999; Veblen 2003). Several closely related terms have been applied by previous 

authors, including “range of natural variability,” “reference variability” (Manley 

and others 1995),“natural variability” (Landres and others 1999), and “historic 

range of variability” (Dillon and others 2005; Meyer and others 2003; Morgan 

and others 1994; Veblen 2003). Common to all of these definitions is an explicit 

recognition of the dynamic nature of ecological systems and the necessity of 

analyzing ecosystems at multiple scales. Our HRV definition closely follows 

earlier ones: the spatial and temporal variation in key ecosystem processes and 

characteristics prior to major alteration by Euro-Americans—approximately 

the mid- to late-1800s for most of the region. Spatial and temporal scales must 

be explicitly discussed because there are differences in the timing and nature 

of initial Euro-American alteration of different parts of the region and among 

wetland and riparian ecosystem types. The most appropriate scales of analysis 

differ for each wetland type and are discussed within individual chapters.

Wetland Definitions and Applicable Laws__________

While few people have difficulty defining what constitutes a forest, there is no 

universal definition of a riparian area or wetland. Because of their high ecological diversity; special legal status under certain federal, state, and local laws; and 

unique place at the intersection of several scientific disciplines, a number of 

regulatory and scientific definitions have been developed. Regulatory definitions of wetlands, most notably the one used by U.S. Army Corps of Engineers 
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Table 3—Principal wetland definitions used by Federal agencies.

U.S. Fish and Wildlife Service, National Wetlands Inventory program (Cowardin et al. 1979)

“Wetlands are lands transitional between terrestrial and aquatic systems where the water table is 

usually at or near the surface or the land is covered by shallow water...wetlands must have one or 

more of the following three attributes: (1) at least periodically, the land supports predominately hydrophytes; (2) the substrate is predominately undrained hydric soil; and (3) the substrate is nonsoil 

and is saturated with water or covered with shallow water at some time during the growing season 

of each year” 

U.S. Army Corps of Engineers, section 404 manual (1987)

“...those areas that are inundated or saturated by surface or groundwater at a frequency and duration sufficient to support, and that under normal circumstances do support, a prevalence of vegetation typically adapted for life in saturated soil conditions. Wetlands generally include swamps, 

marshes, bogs, and similar areas.” [33 CFR323.2(c); 1984]

National Research Council (1995)

“A wetland is an ecosystem that depends on constant or recurrent, shallow inundation or saturation at or near the surface of the substrate. The minimum essential characteristics of a wetland are 

recurrent, sustained inundation or saturation at or near the surface and the presence of physical, 

chemical, and biological features reflective of recurrent, sustained inundation or saturation. Common diagnostic features of wetlands are hydric soils and Hydrophytic vegetation. These features 

will be present except where specific physiochemical, biotic, or anthropogenic factors have removed them or prevented their development.” 

USFS regulations: National Forest Management Act of 1976 – Section: 6(g)(3)(E)(iii)

Sec (E) insure that timber will be harvested from National Forest System lands only where:(i) soil, 

slope, or other watershed conditions will not be irreversibly damaged; (iii) protection is provided 

for streams, stream-banks, shorelines, lakes, wetlands, and other bodies of water from detrimental 

changes in water temperatures, blockages of water courses, and deposits of sediment, where harvests are likely to seriously and adversely affect water conditions or fish habitat.

in regulating wetlands under Section 404 of the Clean Water Act (Tiner 1999), 

are geared towards developing consistent and uniform criteria for field delineation. Precise wetland boundaries are difficult to identify for a number of reasons 

(Choesin and Boerner 2002; Kirkman and others 1998). Interannual and decadalscale variance in water availability can shrink or enlarge the area of saturated 

or flooded soils within the same site through time. The timing and duration of 

soil saturation may vary among years. Rare floods or plant colonization patterns may change landforms and their vegetation patterns. Plant species may 

have individualistic responses to environmental conditions producing complex 

distribution patterns (Tiner 1991).

More science-oriented definitions have been developed for wetlands and riparian 

areas, such as the one developed by the National Research Council (National 

Research Council 1995), that are broader in scope, emphasizing the unique suite 

of hydrological, chemical, and biological characteristics that differentiate wetlands 

from uplands. The definition developed by the U.S. Fish and Wildlife Service for 

use in its National Wetlands Inventory (NWI) program is commonly used and 

contains elements found in both the Corps of Engineers and National Research 

Council definitions (Cowardin and others 1979). All definitions recognize, to 

one degree or another, the key role of hydrologic processes in wetland formation 

and the resulting suite of soil and vegetation characteristics (Table 3).
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A variety of regulatory and scientific definitions of riparian ecosystems have 

also been created. The term riparian is derived from the Latin riparius meaning 

stream bank (DeBano and Schmidt 2004). Some definitions retain this geographic 

perspective, including one used by the USFS that defines riparian areas as “geographically delineated areas with distinctive resource values and characteristics, 

that are comprised of aquatic and riparian ecosystems, floodplains, and wetlands. 

They include all areas within 100 feet from the edge of perennial streams or other 

water bodies...A riparian area is a transition between the aquatic ecosystem and 

the adjacent terrestrial ecosystem and is identified by soil characteristics and 

distinctive vegetation communities that require free and unbound water.” (USDA 

FS 2000). A significant point regarding this definition is that, while referencing certain soil and vegetation characteristics, all areas within a 100-foot buffer 

around a perennial surface water feature are considered riparian, regardless of 

ecological characteristics. Under this definition, all of the five ecosystem types 

covered in this report would be considered riparian if they occurred within this 

buffer. While the NWI definition classifies riparian areas as one type of wetland, 

the USFS definition considers wetlands such as marshes or fens to be a subset of 

the term riparian. An additional definition, formulated by the National Research 

Council, defines riparian areas as: “transitional between terrestrial and aquatic 

ecosystems and ... distinguished by gradient(s) in biophysical conditions, ecological processes, and biota. They are areas through which surface and subsurface 

hydrology connect water bodies with their adjacent uplands” (National Research 

Council 2002). Each of these definitions has value in a specific context.
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2. Data and Information Sources

We utilized a wide range of information sources to make inferences on the HRV 

of key ecosystem parameters. Broadly, these can be classified into several categories: (1) contemporary studies examining the distribution, structure, or function 

of wetlands and riparian areas; (2) instrumental records (e.g., staff gauge and 

weather station data) and proxy evidence for past climates, hydrologic regimes, 

or vegetation; (3) studies documenting the effects of natural or anthropogenic 

disturbances on wetland and riparian ecosystems; (4) historical photographs or 

explorer accounts; and (5) discussions with scientists or USFS staff. Our report 

is largely a review of existing literature; however, we also analyzed primary 

vegetation and geospatial datasets to develop a better understanding of wetlands 

and riparian areas in the region.

We gave the greatest emphasis to peer-reviewed research. Unpublished theses 

and dissertations, USFS technical reports, and Forest Planning documents were 

also used, and they sometimes provided the only information on specific topics 

or ecosystem types. When available, we relied upon research conducted within 

Region 2 National Forests and Grasslands or from similar sites in the region. 

However, studies of certain processes and wetland and riparian types, such as 

ephemeral streams and salt flat ecosystems, are relatively rare, so where necessary, we relied upon information gleaned from studies outside of the region. Each 

National Forest and Grassland is unique in its ecological and human history; HRV 

assessments made using data from studies conducted offsite should therefore be 

viewed as hypotheses to be tested using data collected on site (Veblen 2003).

Instrumental Records __________________________

Stream gauges

An extensive network of stream gauges is located in the region, providing important data on the surface water resources (Table 4). The earliest stream gauges 

were established along major rivers in Region 2 beginning in the late 1800s, with 

many additional gauges added during the Twentieth Century, principally by the 

U.S. Geological Survey. Real-time and historical stream flow records are available 

for hundreds of stream gauges in Colorado, Wyoming, South Dakota, Nebraska, 

and Kansas, although their utility for evaluating historical stream flow patterns 

is often limited by short periods of record that fail to capture the full range of 

hydrologic variability that is important to geomorphic and vegetation dynamics. 

Moreover, very few small watersheds have been gauged, and as a consequence, 

there are relatively few studies of long-term stream flow variability in ephemeral 

and intermittent ecosystems.
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Table 4—USGS stream gauges in Region 2 with the longest periods of record.

Years of Length of

USGS site number USGS site name availability record (yr) State

13011000 Snake River near Moran 1904-present 99 WY

6693000 North Platte River at North Platte 1895-1997 102 NE

8251500 Rio Grande near Lobatos 1900-present 102 CO

6714000 South Platte River at Denver 1895-present 107 CO

6729500 South Boulder Creek near Eldorado Springs 1888-1995 107 CO

6724000 St. Vrain Creek at Lyons 1888-1998 110 CO

8220000 Rio Grande near Del Norte 1890-present 112 CO

7096000 Arkansas River at Canon City 1889-present 113 CO

6752000 Cache La Poudre River near Fort Collins 1882-present 120 CO

7144300 Arkansas River at Wichita 1877-present 125 KS

6889000 Kansas River at Topeka 1869-present 133 KS

Stream gauge data are essential to understanding riparian vegetation dynamics, 

as flooding regimes are one of the principal disturbance types controlling river 

geomorphology (Kochel 1988; Wolman and Miller 1960), plant establishment 

(Cooper and others 2003a; Johnson 2000), community composition, and vegetation structure (Baker 1989, 1990; Friedman and Auble 1999). Gauge data provide 

some of the best information to evaluate hydrological alterations of rivers including flow diversions, augmentations, or impoundments and the resulting changes 

in ecological systems (Andersson and others 2000; Cooper and others 1999; 

Johnson 1998; Richter and others 1996; Richter and others 1998). Records can 

also help evaluate indirect effects to riparian and wetland ecosystems resulting 

from logging, fire, or land use change in a watershed (Cleaveland 2000; Dwire 

and Kauffman 2003; Kondolf and others 2002; Legleiter and others 2003).

The two most basic types of information about rivers that can be gained from 

gauge records are stage—water depth above an arbitrary datum— and discharge— 

the volume of water flowing past a point for some period of time (Dunne and 

Leopold 1978). The two are related as a function of stream channel morphology 

and are directly correlated to the probability of flooding, a key factor governing riparian vegetation patterns. Instantaneous peak discharges are often more 

strongly correlated with vegetation establishment patterns than mean flows, 

although in many alluvial rivers, it is the more frequent bankfull discharges that 

accomplish the most geomorphic work (Leopold and others 1964; Wolman and 

Miller 1960). Both types of data were obtained from the USGS for a number of 

gauges, particularly those with long records.

Groundwater measurements

Despite the importance of groundwater to many wetlands, long-term datasets 

from groundwater monitoring wells and piezometers are relatively rare for most 

wetland and riparian ecosystems (Winter and Rosenberry 1998; Winter and others 2001). Where available, these data are typically of short duration, although 

even these provide important insights into wetland functions (Cooper and others 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 10

2003c; Scott and others 1999; Stromberg and others 1996; Winter and LaBaugh 

2003). Because groundwater systems are often complex and poorly understood, 

our ability to confidently predict groundwater dynamics for wetlands, let alone 

assess historical groundwater dynamics, is limited. Available data are helpful for 

understanding the natural hydrologic functions of different wetland types and 

their susceptibility to anthropogenic change.

Weather stations

Weather stations, particularly those with long periods of record, provide data for 

understanding climatic variability at short and intermediate time scales; provide 

information needed to validate proxy models of past climatic conditions (Woodhouse and Overpeck 1998); and can provide information useful for interpreting 

the structure and function of contemporary wetlands and riparian ecosystems. 

To help characterize important differences across the region, we selected stations 

across a latitudinal and elevation gradient, focusing on stations with the longest 

records and those closest to USFS units. Like stream gauges, all weather stations 

postdate the arrival of Euro-Americans in the region, limiting the stations’ direct 

utility for studies of historical climates.

Paleoclimatic, Paleohydrologic, and 

Paleoecological Studies ________________________

The longest instrumental records for the western United States are relatively 

brief compared with the time scales necessary to understand key geological, 

geomorphic, climatic, and ecological processes shaping wetland and riparian 

ecosystems. For instance, direct geomorphic drivers of riparian plant establishment, such as large floods or landslide events, or disturbances like fire or insect 

outbreaks, which can indirectly affect wetland or riparian systems by altering 

water or sediment fluxes, may occur at time scales of centuries or longer, while 

the glacial advances and retreats that created landforms suitable to the development of many montane and subalpine fens and wet meadows occur over tens of 

thousands of years. In addition, climatic variability at shorter annual and decadal 

scales is an important driver of the hydrologic function and species composition 

of wetlands such as marshes (Bolen and others 1989; Winter and others 2001) 

and salt flats (Ungar 1998) and would have been key to vegetation dynamics 

prior to Euro-American settlement.

Researchers working in a range of disciplines have developed methods for 

reconstructing past climates, hydrologic regimes, and plant communities. 

Sources of proxy evidence are tree rings, pack rat middens, plant macrofossils, 

14C-dating of peat, microfossils such as pollen or phytoliths, lake sediments, lake 

levels, paleoflood indicators, and fossil dunes (Table 5). Collectively, these methods provide evidence for large climate shifts at a range of spatial and temporal 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 11

Table 5—Representative paleoecological and paleoclimatic studies relevant to this 

assessment.

Study type Reference

Pollen Fall 1992, Fall 1994, Denniston et al. 1999,

Dean et al. 1996, Baker 1969, Whitlock and

Millspaugh 2001, Whitlock 1993

Eolian activity Forman et al. 2001, Dean et al. 1996, 

Clarke and Rendell 2003

Macrofossil Baker 1969, Baker 2000

Packrat middens Pearson and Betancourt 2002, Cole et al. 1997

Dendrological reconstructions Yansa 1998, Woodhouse and Brown 2001, 

Grissino-Mayer and Swetnam 2000, 

Veblen et al. 2000, Sherriff et al. 2001, 

Mast et al. 1998, Woodhouse 2001, 

Graumlich et al. 2003

Speleotherms Denniston et al. 1999

Lake sediments Fritz 1996, Laird et al. 2003, Laird et al. 1996, 

Baker 1969

scales and, when interpreted in light of studies of modern ecosystems, provide 

a means of inferring past ecological function. As an example, pollen evidence 

has been commonly used to document vegetation responses to changing climates 

(Fall 1992, 1994; Meyer and others 2003; van Ahlefeldt 2000).

Classification of Region 2 Wetlands and 

Riparian Communities__________________________

Vegetation classifications are critical for identifying the types of plant communities occurring in any region and prioritizing conservation strategies. They also 

provide an effective way to understand the range of community types that occur 

with different hydrologic regimes and past disturbance regimes. Extensive work 

on vegetation classification has been done in Region 2, although most studies have 

been conducted for individual forests rather than for the entire region. An exception is the statewide wetlands classification produced by Colorado (Carsey and 

others 2003). To analyze the similarities and differences of wetland and riparian 

plant communities across Region 2, we gathered primary vegetation composition 

and cover data sets from the region and combined them into a relational database 

(MS Access, Microsoft, Inc.) (Table 6). These data were used as inputs into PC 

Ord (McCune and Mefford 1999), a multivariate statistics package for analysis.
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Historical Photographs and Accounts_____________

The written accounts and photographs of early explorers and settlers can provide 

useful, if somewhat biased, insights into changes in wetlands and riparian ecosystems. Fur trappers were pursuing beaver throughout the Central and Southern 

Rockies by the 1820s (Wohl 2001), yet they left few accounts of their travels. 

Important written accounts of landscapes from members of early exploring parties to the Great Plains and Rocky Mountains are Edwin James’ account of the 

Stephen Long expedition to Colorado in 1820 (Evans 1997) or George Custer’s 

1874 expedition through the Black Hills (Horsted and Grafe 2002). Photographers 

accompanying some of the early expeditions, including W. Illinsworth and W. 

H. Jackson, took many excellent photographs that provide visual evidence of 

some riparian and wetland areas in the early period of Euro-American settlement. 

Photographs and accounts were obtained from several sources, primarily the 

USGS photographic library in Lakewood, Colorado, the Denver Public Library’s 

Western History collection, and a number of publications (e.g., Bradford 1998; 

Horsted and Grafe 2002).

Table 6—Sources of plot data used in vegetation classification

Classification and ordinations Location and extent

Girard et al. 1997 Wyoming; Bighorn National Forest

Marriott and Faber-Langendoen 2000 South Dakota; Black Hills Ecoregion

CNHP & Cooper Colorado; entire state

Cooper and Andrus 1994 Wyoming; Wind River Range

Heidel and Laursen 2003a Wyoming; Medicine Bow National Forest

Jones 1992 Wyoming; Medicine Bow National Forest (riparian)

Walford et al. 1997 Wyoming; Shoshone National Forest

Jones and Walford 1995 Wyoming; Eastern Plains

NPS, VEG Bank Nebraska; Agate Fossils National Monument
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 13

3. Wetland and Riparian Ecosystems of 

Region 2: Structural and 

Functional Characteristics and 

Ecological and Historical Setting

Introduction __________________________________

The relative abundance of water on a seasonal, annual, and long-term basis differentiates wetland and riparian ecosystems from uplands. The source of water 

(precipitation, surface flow, or groundwater) and its spatial and temporal distribution comprise an ecosystem’s hydrologic regime, which influences a wide 

range of ecological attributes, from rates of nutrient cycling and productivity 

(Mitsch and Gosselink 2007; Pinay and Naiman 1991; Weltzin and others 2000) 

to floristic composition (Cooper and Andrus 1994; Seabloom and others 1998). 

At fine scales, the hydrologic regime operates as a driver of ecological structure 

and function, while at broader scales, it shapes patterns of wetland and riparian 

ecosystem abundance and distribution. Additional climatic and geological factors 

operating across a range of spatial and temporal scales affect biogeochemical 

processes and vegetation, either acting independently or through their control 

of hydrologic processes. Collectively, climate, geology, and hydrology form the 

template upon which wetlands and riparian ecosystems develop; it is largely by 

examining how they influence contemporary systems that we form a basis for 

inferring the HRV of key ecological attributes.

Also important to developing an HRV assessment is a critical examination of the 

human history of the region, prior to and following the arrival of Euro-Americans. 

Though there is continued debate regarding the extent and magnitude of their 

influence on natural systems (Denevan 1992; Flores 2001; Vale 1998), Native 

Americans likely had some impact on Region 2 ecosystems, at least locally, although there is insufficient information from which to draw general conclusions. 

There is less uncertainty regarding the impact of Euro-Americans on wetland 

and riparian ecosystems—since the Eighteenth Century, wetlands and riparian 

ecosystems have been among the most intensely and systematically altered in 

North America (Dahl 1990). The wide variety of impacts, detailed later in this 

and subsequent chapters, include hydrologic alterations associated with dams 

and water diversions (Graf 1999; Nilsson and Berggren 2000; Pringle 2000), 

agricultural drainage (Dahl 1990), grazing (Belsky and others 1999; Fleischner 

1994; Patten 1998), and the widespread introduction of non-native species (Crooks 

2002; Mack and others 2000; Stein and Flack 1996).

The extent and magnitude of Euro-American impacts across Region 2 are variable. Many systems, such as those found in remote wilderness areas, may have 

seen little direct impacts, while others, such as lower montane or Great Plains 

ecosystems, may have been more significantly altered. Our assessment of where 

and to what degree departures from the HRV may have occurred will, therefore, 

involve an examination of both the region’s human and ecological history.
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The ecosystem types we evaluate in this assessment vary considerably from one 

another in their relative abundance, landscape distribution, vegetation composition and dynamics, and functional characteristics. Because of these differences, 

separate assessments are presented for each major ecosystem type—fens, wet 

meadows, marshes, salt flats, and riparian areas. The main goal in this chapter 

is to characterize the key process variables and attributes that distinguish these 

major ecosystem types and control their structure and function. We also provide 

a broad overview of topics common to each, including: (1) the contemporary 

and historical climatic, geologic, and hydrologic patterns in the region, (2) a 

description of principal plant communities associated with different wetland 

types, (3) the major anthropogenic impacts to wetland and riparian ecosystems, 

and (4) a review of the human and ecological history of Region 2 as it pertains 

to the assessment.

Overview of the Principal Wetland and Riparian 

Types in the Region____________________________

Though frequently managed and regulated as a single resource, the wetland and 

riparian ecosystem types described in this document differ widely in their functional and structural characteristics. A common problem is the often ambiguous 

terminology used to describe wetlands and riparian ecosystems (Mitsch and 

Gosselink 2007). Many terms are regional in application and have had the effect of unnecessarily separating out systems which may, in fact, be quite similar 

in function. For example, playas—precipitation-driven depressional wetlands 

occurring in the Great Plains—are often thought to be restricted to the “Playa 

Lakes region” of Texas, New Mexico, Kansas, New Mexico, and Southeast 

Colorado (Nelson and others 1984). However, similar ecosystems can be found 

in eastern Wyoming, the rainwater basin of Nebraska, and unglaciated portions 

of the South Dakota mixed grass prairie (Smith 2003).

Likewise, some commonly used terms lump very different ecosystems together 

based on criteria with little ecological significance. For instance, riparian areas 

are often defined based on arbitrary geographic criteria, like absolute distance 

from a stream channel, resulting in the combination of communities with primarily surface water hydrologic regimes with those supported principally by 

groundwater. Though there may be some common species between these two 

ecosystems, it is likely that each would respond differently to a major change in 

the watershed such as dam construction, water diversion, or timber harvest. Given 

the goals of this assessment, it is critical to correctly identify ecosystems that 

function similarly. Since there are few data available that pertain directly to the 

historical variability of vegetation in wetland and riparian ecosystems, our HRV 

assessments are based largely on our understanding of how key drivers affect 

the structure and function of contemporary ecosystems, and how anthropogenic 

factors may have altered these drivers.
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The five main ecosystem types that form the basis of this assessment possess 

certain hydrologic characteristics, which in turn drive additional differences in 

floristic composition, vegetation structure, and biogeochemical functioning. 

The following section outlines some of these differences, distinctions that will 

be addressed in greater detail in subsequent chapters.

Riparian areas

Riparian ecosystems are diverse in landform, hydrologic regime, vegetation, and 

ecological functioning and they occur in all portions of Region 2. Included are 

ecosystems found along steep, small-order headwater streams in the mountains, 

intermittent and ephemeral streams in the foothills and plains, and areas along 

broad, low-gradient alluvial rivers on the plains (National Research Council 

2002). The principal characteristic unifying riparian ecosystems is the presence 

of unidirectional moving water, which has the potential to erode and transport 

sediment. The frequency, magnitude, and energy of floods, which vary widely 

due to differences in basin size, topography, and climatic regime, affect all ecological processes from nutrient cycling to plant establishment (Adair and others 2004; Cooper and others 2003b; Karrenberg and others 2002). Hydrologic 

regimes of riparian areas vary widely across the region, from relatively stable, 

groundwater-driven flows in the Sandhills of Nebraska (Bio/West 1986; Winter 

1999) to infrequent and unpredictable flash floods associated with intermittent 

and ephemeral streams on the Great Plains (Friedman and Lee 2002) and lowerelevation mountain regions. Despite this complexity, there are definite patterns 

in vegetation structure and ecosystem function across elevation gradients, which 

provide a basis for characterizing riparian HRV.

Fens

Fens form in a variety of landscape settings and show great floristic diversity 

(Bedford and Godwin 2003; Cooper 1996). However, all fens share one basic 

characteristic: they have stable groundwater-driven hydrologic regimes with 

high water tables that retard organic matter decomposition and promote peat 

accumulation. Depending upon the geological characteristics of the watershed 

supplying their water, they can vary considerably in water chemistry. Fens are 

among the most floristically unique ecosystems in the region, often supporting rare 

or endemic species (Chadde and others 1998; Cooper and Andrus 1994; Cooper 

and Sanderson 1997; Heidel and Laursen 2003a,b; Weber 2003). In contrast to 

riparian areas, there is little sediment moving into or out of fens. Fens may occur adjacent to riparian areas in mountain valley bottoms; however, they have 

independent water sources and function quite differently from riparian areas. 

Fens may support floating mats of vegetation, which is unique among wetlands 

within the region.
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Marshes

Marshes typically form in depressions and include such diverse systems as prairie 

potholes, playas, lacustrine fringe wetlands, and abandoned oxbow lakes on river 

flood plains. Marsh hydrology can be extremely variable and frequently includes 

both prolonged periods of inundation as well as extended dry periods, particularly during dry years (Winter and Rosenberry 1998; Winter and others 2001). 

A second key variable is water depth. Marshes periodically have deep standing 

water, which limits the species that occur to aquatic and wetland species tolerant 

of submersion. A third key variable is water and soil chemistry. Some marshes 

have significant flow-through of fresh water. In contrast, terminal basins, where 

water is lost primarily through evapotranspiration, may be highly saline, which 

affects species composition, litter decomposition, and productivity (Thormann 

and others 1999). Hydrologic variability, water depth, and salinity are key factors 

determining the species composition of marshes, both spatially within and among 

marsh complexes and temporally from wet to dry years (Seabloom and others 

1998; Smith and Haukos 2002; van der Valk and others 1994). Seed banks play a 

particularly important role in marsh vegetation dynamics (Smith and Kadlec 1983; 

van der Valk and Davis 1976; Wilson and others 1993) with large fluctuations in 

species composition commonly occurring over relatively short time scales. The 

strong water depth gradients also generate distinct vegetation zonation patterns 

in many marshes (Johnson and others 1987; Lenssen and others 1999).

Wet meadows

Wet meadows are widespread and likely cover more area than all other wetland 

types combined in the region. They occur from the alpine zone to the plains; 

yet despite their relative abundance, there are relatively few studies examining 

their hydrologic and edaphic characteristics or vegetation dynamics. Although 

wet meadows typically exhibit seasonally saturated soils, they lack the perennial high water tables of fens or the large seasonal and inter-annual water table 

fluctuations that are characteristic of marshes (Mitsch and Gosselink 2007). 

Many wet meadows are managed for livestock forage and hay production, and in 

agricultural areas, meadows frequently develop down slope of unlined irrigation 

canals. Over long periods of time, wet meadows may form as marshes or beaver 

ponds become filled with sediment (Naiman and others 1994).

Salt flats

Salt flats are a widespread wetland type found at low elevations and intermountain basins through the West, but are poorly understood. They generally form in 

closed basins in areas with heavy-textured soils or where evaporation from a high 

water table promotes the accumulation of salts. High salt concentrations create a 

difficult environment for most plants; consequently, plant cover and productivity 
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is generally low and species composition is limited to salt-tolerant species (i.e., 

halophytes) (Dodd and Coupland 1966; Ungar 1966; Ungar 1974b). Ironically, 

because salt flats are so marginal in terms of forage production or suitability for 

crops, they may have been spared many of the anthropogenic impacts that affect 

the other primary wetland types.

Vegetation of Region 2 Wetlands and 

Riparian Areas ________________________________

Wetlands and riparian areas support a variety of plant species and community 

types found nowhere else in Region 2 landscapes. For example, 183 of the nearly 

600 ecological system types (groups of plant community types that tend to cooccur in similar ecological settings) defined by Commer and others (2003) in 

their analysis of Region 2 vegetation are found in wetlands. Wetlands and riparian 

systems function as both local and regional centers of biodiversity (Naiman and 

others 1993; Stohlgren and others 1997) and support many rare or endemic species (Cooper and others 2002; Cooper and Sanderson 1997; Heidel and Laursen 

2003 a,b; Mellmann-Brown 2004).

Variation in a site’s hydrologic and disturbance regimes strongly influence vegetation composition among wetland types. Environmental characteristics, such 

as periodic deep inundation in marshes or sustained high water tables in fens, 

limit the species that can occur at any site. Only species with the appropriate 

physiological and life history adaptations are able to persist and compete in the 

distinct environmental conditions typical of each wetland type. Examples of 

adaptations commonly found in wetland species are aerenchyma, adventitious 

roots, pressured gas flow, and salt tolerance (Mitsch and Gosselink 2007; Rood 

and others 2003a). As a consequence, many species found in wetlands occur only 

in wetlands. Some species show a range of fidelity for wetland habitats and are 

referred to as facultative wetland species (Reed 1988). Even within a particular 

species, there may be regional differences in the frequency at which plants are 

found in wetlands (Tiner 1999).

Overall variation in Region 2 wetland and riparian vegetation

An indirect ordination, shown in Figure 2, was created using detrended correspondence analysis (DCA) with the computer program PcOrd (McCune and 

Mefford 2006) and vegetation data from 5266 plots from wetland and riparian 

studies done in the region. Variation along axis 1 is due primarily to elevation 

and water table permanence. High-elevation sites with perennially high water 

tables occur on the left and low-elevation sites with varying water tables are on 

the right. Variation in Axis 2 is largely due to water chemistry, with saline sites 

near the top and freshwater sites at the bottom. Each plot is represented by one 

point and the centroids of key indicator species are shown in the ordination space.
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Figure 2—Ordination of Region 2 vegetation data. Common plant species are shown using the following abbreviations: 

CARNIG = Carex nigricans, DREADU = Drepanocladus aduncus, ELEQUI = Eleocharis quinqueflora, KALPOL = Kalmia 

polifolia, CARILL = Carex illota, TOMNIT = Tomenthypnum nitens, BETFON = Betula fontinalis, CARSCO = Carex scopulorum, 

SALPLA = Salix planifolia, CARVES = Carex vesicaria, SENTRI = Senecio triangularis, CARAQU = Carex aquatilis, CARUTR = 

Carex utriculata, CALCAN = Calamagrostis canadensis, SALMON = Salix monticola, SALGEY = Salix monticola, DESCES = 

Deschampsia cespitosa, PENFLO = Pentaphylloides floribunda, PICPUN = Picea pungens, BETGLA = Betula glandulosa, 

POPANG = Populus angustifolia, SALIRR = Salix irrorata, CARSIM = Carex simulata, JUNARC = Juncus arcticus, SALEXI = 

Salix exigua, CARNEB = Carex nebraskensis, CIRCAN = Cirsium canadensis, AGRGIG = Agrostis gigantea, BROINE = 

Bromopsis inermis, NEGACE = Negundo aceroides, ULMAME = Ulmus americanus, POPDEL = Populus deltoides, ELEANG = 

Eleagnus angustifolia, PHRAUS = Phragmites australis, SPAEUR = TAMRAM = Tamarix spp., PASSMI = Pascopyrum smithii, 

PUCAIR = Puccinellia airoides, SUACAL = Sueda calcioliformia, SARVER = Sarcobatus vermiculatus, BOLMAR = Bolboschoenus 

maritimus, DISSTR = Distichlis stricta, AMPNEV = Amphiscirpus nevadensis, SPOAIR = Sporobolus airoides, GLAMAR = Glaux 

maritimus, TRIMAR = Triglochin maritimum, SCHPUN = Schoenoplectus pungens, ELEPAL = Eleocharis palustris, SPAPEC = 

Spartina pectinatus, GLYGRA = Glyceria grandis, BECSYZ = Beckmannia syzygachne, TYPLAT = Typha latifolia, SCHLAC = 

Schoenoplectus lacustris, SPAEUR = Sparganium eurycarpum.
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The five major wetland types can be identified in the ordination space. Fens are 

on the far left, salt flats are on the top right, marshes are on the bottom right, 

wet meadows are on the bottom center, and riparian areas are elongated from 

the top right to the left. The riparian continuum includes springs and small headwater mountains streams with herbaceous communities dominated by Senecio 

triangularis, Salix geyeriana and other willow-dominated thickets along small 

to medium sized mountain rivers; Populus angustifolia and Picea pungensdominated forests along mid-elevation mountain rivers; Acer negundo forests 

in canyons and foothills; and on the plains, Salix irrorata and Populus deltoides

forests with Tamarix and Elaeagnus angustifolia.

A range of fen types occur, including basin types dominated by Carex utriculata

and Carex vesicaria; slope fens dominated by Eleocharis quinqueflora, Carex

aquatilis, and Carex illota; and wooded fens with Kalmia polifolia and Salix

planifolia. Wet meadows include communites dominated by Deschampsia cespitosa at high elevations; Juncus arcticus at mid- to low-elevations; and Carex

nebraskensis at lower elevations. A wide range of marshes occur and due to deep, 

prolonged inundation, many are monocultures or have very low floristic diversity. The deepest water sites have Schoenoplectus lacustris and Typha latifolia, 

while sites with more shallow water are dominated by Eleocharis palustris and 

Schoenoplectus pungens. The later species can also occupy marshes that are 

highly saline and that are found adjacent to salt flats. The salt flat communities 

are also species poor—often dominated by monocultures of halophytic species 

such as Distichlis spicata.

Climate ______________________________________

Issues of scale

As with the ecological parameters examined in this HRV assessment, climate 

variables require explicit spatial and temporal bounds for proper interpretation 

(Bradley 1999). Measures of central tendency and variability are dependent 

upon the scale of analysis. For instance, climatic variables such as global mean 

temperature are relatively stable at time scales of decades to centuries. However, 

the paleoclimatic record demonstrates significant variation over time scales of 

tens of thousands to millions of years. Similarly, notions of change are frequently 

scale dependent. Climatic patterns can be heterogeneous over relatively small 

spatial scales due to factors such as local topography.
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Spatial and temporal climate patterns and historical 

variability

Directly measured climate records provide important although temporally limited 

data for evaluating climatic variability. Assessments of historical climates from 

paleoclimatic and instrumental records are, for some wetland types, the primary 

information source for making inferences on the HRV for hydrologic and ecological 

variables since no reliable instrumental data are available. Unlike forests, where 

tree rings or stand structure can provide evidence of past disturbance regimes and 

vegetation dynamics, herbaceous-dominated ecosystems such as marshes or wet 

meadows generally leave little evidence of long-term dynamics. Additionally, 

relative to forests, few studies have succesfuly reconstructed historical conditions 

or dynamics in wetlands or riparian ecosystems.

Climate integrates the mean and variability of atmospheric conditions, including 

temperature, precipitation, humidity, and solar radiation, and exerts a key influence on the Earth’s biota. Over broad spatial and temporal scales, it is a major 

driver of geomorphic as well as ecological processes, influencing patterns of 

glaciations (Bradley 1999), erosion and sediment dynamics, and drainage network 

formation (Knighton 1998; Leopold and others 1964; Rosgen 1996). It is also 

a principal determinant of the distribution and functional characteristics of the 

wetland and riparian ecosystem types found in Region 2. The climatic variables 

of greatest significance to wetland and riparian ecosystems are temperature and 

precipitation, which vary along both latitude and elevation gradients. However, 

considerable variability can occur at a given latitude or elevation due to factors 

such as land form and land use (Kittel and others 2002).

As a fundamental thermodynamic control on rates of chemical, physical, and 

biological processes, temperature—its range, minimum, and maximum values—is 

an important factor that shapes such diverse phenomena as species ranges, biogeochemical cycling, erosion and mass wasting rates, and hydrologic cycling. 

Temperature regimes throughout Region 2 are continental in character, exhibiting greater summer and wintertime temperature fluctuations relative to maritime 

regions (Kittel and others 2002).

Mean annual temperature at a given elevation increases from north to south, 

although this is due more to differences in winter than in summer temperatures 

(Kittel and others 2002). The relationship between elevation and temperature, both 

minimum and maximum values, varies in an approximately linear fashion, with 

cooler temperatures observed at higher elevations (Figures 3, 5). In mountainous 

areas of Wyoming and Colorado, steep elevation gradients and complex topography create environmental heterogeneity that is not present in the Great Plains.
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Precipitation patterns may influence groundwater-driven wetlands by controlling 

aquifer recharge. Unlike temperature, only weak trends between precipitation and 

latitude occur. Elevation and topography are much stronger controls on precipitation due to orographic effects, which occur when moist air is lifted and cools as it 

moves over a mountain range (Bradley 1999). A higher amount of precipitation 

falls on the windward side of a mountain ridge, and precipitation is generally 

significantly lower on the lee side of mountains, resulting in the formation of a 

rain shadow. The effect of mountain ranges on spatial patterns of precipitation is 

evident in maps of precipitation in the region—significantly greater heterogeneity occurs in the mountainous portions of Wyoming and Colorado relative to the 

Great Plains (Figures 4, 5).

The seasonality of precipitation varies across the region due to regional synoptic 

weather patterns—an important factor influencing Region 2 wetlands and riparian 

areas. Monsoonal precipitation, typified by intense late summer thunderstorms, 

is important in the southern tier of National Forests (Kittel and others 2002) but 

is uncommon farther north. Monsoonal precipitation may produce the largest 

floods on some low-elevation streams (Fleener 1997) and recharge aquifers that 

sustain fens at higher elevations (Cooper 1990a).

Figure 3—PRISM mean annual temperature (°F) for Region 2 from the National Climatic Data Center’s U.S. 

National 1971-2000 Climate normals (data online: http://datagateway.nrcs.usda.gov/).
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Figure 4—PRISM mean annual precipitation (inches) from the National Climatic Data Center’s U.S. National 

1971-2000 Climate normals (Data online: http://datagateway.nrcs.usda.gov/).

Figure 5—Mean annual temperature and annual precipitation plotted for each USFS unit. Values from PRISM GIS layers 

(1971 to 2000 normals) were extracted for each USFS unit using zonal statistics in ArcGIS (Data online: http://datagateway.

nrcs.usda.gov/).
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Potential evapotranspiration (PET) refers to the potential rate of evaporation and 

transpiration under optimal soil moisture and plant cover conditions (Dunne and 

Leopold 1978). The PET/precipitation ratio provides a measure of a region’s 

aridity. Regions with a PET/precipitation ratio greater than 1.0 have a moisture 

deficit, while regions with a ratio less than 1.0 have a moisture excess. This ratio 

may influence seasonal soil water content, runoff, and hydrologic conditions 

needed to support wetlands. Evaporation is driven in large part by temperature 

and it shows a similar pattern at broad spatial scales. Because PET greatly exceeds 

precipitation for most of the region, wetlands are generally less abundant than in 

more humid areas such as the southeast United States (Kroes and Brinson 2004).

Drought, defined here as the prolonged and abnormal absence of precipitation, 

influences wetlands and riparian areas in many ways, including its influence on 

fire regimes in adjacent ecosystems (Donnegan and others 2001; Hessl and others 2004; Swetnam and Betancourt 1998) and its control on hydrologic regimes 

and, therefore, vegetation dynamics, in marshes and riparian areas (Hardy 1995; 

Rejmankova and others 1999; Winter and Rosenberry 1998).

Palmer Drought Severity Index (PDSI) integrates temperature and rainfall 

data to determine relative dryness (Barry and Chorley 2003) at a time scale 

of months. A PDSI value of 0 indicates average precipitation, negative values 

indicate increasing drought conditions, and positive values indicate wetter conditions. A PDSI value of -2 indicates moderate drought, while -4 indicates extreme 

drought. Since the PDSI is standardized to local climate, it can be applied to any 

region. Though drought is unpredictable, instrumental records and paleoclimatic 

reconstructions have documented numerous drought episodes and indicate that 

moderate to extreme drought recurs at decadal to millennial scales (Cook and 

others 1999).

PDSI records document three major decadal-scale moisture regimes on the 

Great Plains. The early part of the Twentieth Century, from approximately 1905 

to 1917 when a large influx of homesteaders immigrated to the region, was one 

of the wettest periods since the Sixteenth Century (Figure 6). Paleoclimatic 

reconstructions of PDSI values indicate that similar wet periods (pluvials) occurred in the Sixteenth, Seventeenth, and Nineteenth centuries (Fye and others 

2003) (Figure 7). There is only a weak correspondence between precipitation 

records and climatic reconstructions from Colorado and Wyoming and those 

from the Plains, suggesting different large-scale climatic drivers. The “dust bowl” 

drought of the 1930s has no analog for size or severity in the historical record. 

However, a “megadrought” in the late Sixteenth Century that was concentrated 

over the Southwest appears to have surpassed the 1930s drought in severity and 

duration (Acuna-Soto and others 2002; Fye and others 2003).
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Paleoclimatic studies have utilized a variety of climate proxies, including tree 

rings (Gray and others 2003; Woodhouse 2003), dune accretion (Clarke and 

Rendell 2003; Dean and others 1996), pollen records (Fall 1992; Fredlund 1995; 

Laird and others 1996), lake levels and sediments (Arnow and Stephens 1990; 

Fritz and others 2000), and ice cores (Soon and Baliunas 2003) to document 

climatic variability for times scales that range from millennia to centuries and 

decades. Climate variability may have broad impacts on species ranges, extinction, endemism (Weber 2003), community composition (Whitlock and Millspaugh 

2001), disturbance regimes such as fire and insect outbreaks in upland ecosystem 

(Grissino-Mayer and Swetnam 2000; Sherriff and others 2001; Swetnam and 

Betancourt 1998), and flooding in riparian ecosystems (Graumlich and others 

2003; Knox 2000; Woodhouse 2001).

Figure 6—Palmer drought severity index (PDSI) estimates (gray lines), 5 year running averages (blue line), and overall 

averages (red lines) reconstructed from regional networks of tree rings for five locations in Region 2. (Data obtained online 

from the NOAA Paleoclimatology Program, http://www.ngdc.noaa.gov/paleo/drought/drght_data.html).
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The El Niño Southern Oscillation (ENSO) is characterized by an increase in the 

sea surface temperature in the eastern equatorial Pacific Ocean and occurs on 

average every three to seven years (Barry and Chorley 2003). ENSO patterns have 

been correlated with anomalous climatic conditions across the world, including 

areas in USFS Region 2 (Cayan and others 1999; Hidalgo and Dracup 2003), and 

are correlated with a wide range of regional climatic, hydrologic, and ecological 

factors, including streamflow variability (Andrews and others 2004), changes 

in fire frequency and magnitude (Hessl and others 2004), and the occurrence 

of extreme weather events (Cayan and others 1999). Drought characteristics, at 

least in the central United States, appear to be weakly correlated with ENSO 

events and more strongly correlated with the interdecadal climate mode called the 

Pacific Decadal Oscillation (PDO) (Englehart and Douglas 2003). Considering 

the overall importance of climate to patterns of wetland distribution and function, understanding large-scale climate drivers like ENSO and PDO can provide 

insight into the natural variability of wetlands.

Figure 7—Summer Palmer Drought Severity Indices (PDSI) from 1400 

to 2000 from three locations in the southern, central, and northern Great 

Plains along longitude 102.5°W at latitudes 35°N, 40°N and 45°N. Values 

to 1978 are tree-ring reconstructed data while those from 1979 to 1990 

are instrumental PDSI data for the same grid points. (Source: http://lwf.

ncdc.noaa.gov/oa/climate/research/2002/sep/st105dv00pcp200209.

html#wgppaleo).
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Geomorphology and Geology ___________________

Wetland and riparian ecosystems are influenced by a variety of geologic and 

geomorphic factors operating over multiple spatial and temporal scales. Over 

broad geological time scales, processes of uplift and erosion, influenced by lithology and climate, are largely responsible for the physiographic characteristics 

of contemporary landscapes. Differences between the tectonically active landscapes of the Rocky Mountains and those of the Great Plains broadly determine 

the landscape distribution of different wetland types. The mineralogy and grain 

structure of geological strata influences processes such as infiltration, sediment 

yield (Knighton 1998), and groundwater chemistry (Winter 2001) and varies at 

a range of scales.

Although some geological processes, such as mountain uplift, are practically 

invariable over the time scales that are relevant to management (decades to centuries), others, such as fluvial geomorphic processes or slope failures, are more 

highly dynamic over shorter temporal scales and can influence the formation and 

function of wetland and riparian ecosystems. Many geologic and geomorphic 

processes such as glaciation, although not presently active, are critical for understanding where and in what form wetlands and riparian ecosystems are found.

Glacial processes have significantly influenced wetland distribution and function 

in portions of Region 2. In North America, some of the highest wetland densities occur in areas affected by past continental glacial advances, most notably 

the prairie pothole region of eastern South and North Dakota, Minnesota, and 

southern Canada (van der Valk 1989). Although no Forest or Grassland in Region 

2 was covered by continental glaciers, localized ice caps and valley glaciers were 

widespread in the mountains in the past, and many small glaciers still occur at 

high elevations in Wyoming and Colorado (Figure 8). Although presently small 

and shrinking, glaciers were larger and more active during the Pleistocene and 

had lasting effects on the physiographic characteristics of many montane and 

subalpine valleys. Many wetlands occur in landforms formed by glacial retreat 

such as kettle basins—small depressions caused by the melting of buried ice 

masses (Winter and Woo 1990). In addition, in many mountain valleys, glaciers 

have indirectly influenced valley form by depositing end moraines and flattening 

valley gradients, thereby retarding the runoff of water. Glaciers also influenced 

areas downstream, shaping the form of many river floodplains and terrace systems by paleo-flood events associated with glacial melting (Knighton 1998).

Extensive sand dune complexes are located in several parts of the region, including in the Nebraska National Forest, which includes a portion of the Nebraska 

Sandhills. Smaller dune systems occur in Wyoming, Kansas, and Colorado. Because of the high permeability and hydraulic conductivity of sand substrates, 

little to no surface runoff occurs during precipitation events. Consequently, water 

storage is high and sites may support extensive groundwater flow systems. These 

sites can support wetlands where local or regional groundwater systems intersect 

the ground surface, typically in interdunal areas (Drda 1998; Novacek 1989). 
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The hydrologic characteristics of streams formed in these settings are strongly 

driven by groundwater inflows and are unusual for Region 2 in their flow stability. 

The Nebraska Sandhills and similar eolian features are also important barometers 

of historical climate change. Stratigraphic and radiocarbon studies conducted 

in Kansas, Colorado, and Nebraska have demonstrated multiple episodes of 

increased dune activity resulting from prolonged drought conditions over the 

last millennium (Clarke and Rendell 2003; Forman and Maat 1990; Muhs and 

Holliday 1995).

Bedrock geology can influence wetlands and riparian areas by influencing sediment yield and composition. Igneous rock types, such as granite, yield relatively 

low amounts of coarse sediment relative to sedimentary rock types such as sandstones or limestone (Knighton 1998). Differences in permeability among different rock strata or the presence of faults or rock fractures can create preferential 

flow pathways resulting in the formation of seeps and springs. These lithologic 

differences influence patterns of groundwater recharge and discharge and may 

affect water table dynamics in wetlands and riparian ecosystems. An example 

is the variable surface water flows in streams flowing through limestone in the 

Black Hills region of South Dakota (Driscoll and others 2000).

Figure 8—Approximate extent of mountain glaciation during the 

Pleistocene (source: Windell et al. 1986).
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Lithology can also influence the concentrations of elements and compounds 

that control wetland water alkalinity and that affect plant growth (e.g., calcium, 

magnesium, and potassium). This variation is driven by different rates and 

modes of weathering between rock types. Calcareous rocks, such as limestone, 

contain calcium carbonate, which reacts with the weak acids typically present 

in rainfall, groundwater, and snowmelt. This chemical weathering can release 

minerals into solution, concentrations of which can subsequently be modified 

by hydrologic mixing with groundwater, precipitation, or surface water. Water 

chemistry variability forms an important environmental gradient that influences 

vegetation patterns in fens (Cooper 1996; Demars and others 1997; Johnson and 

Steingraeber 2003).

The specific geological factors important to ecosystem formation and function 

vary among different wetland and riparian types. Mass wasting events such as 

landslides can influence riparian ecosystems by damming streams and increasing 

sediment flux but may have little effect on plains marsh or salt flat ecosystems.

Hydrologic Drivers of Wetland and Riparian 

Ecosystem Structure and Function _______________

Hydrologic processes control most wetland and riparian functions. Riparian plant 

establishment is linked to the frequency and magnitude of flooding (Auble and 

others 1994; Baker 1990; Cooper and others 2003a; Friedman and Lee 2002; 

Johnson 2000; Scott and others 1997). Water depth and duration controls species 

composition in marshes (Haukos and Smith 2001; Kantrud and others 1989b; 

Neill 1990; van der Valk and others 1994), while water table dynamics influence carbon accumulation rates in fens (Chimner and Cooper 2003; Chimner 

and others 2002). Because of its fundamental importance, many classification 

systems, such as that developed by the National Wetlands Inventory (Cowardin 

and others 1979), use hydrologic regime in differentiating wetland types. The 

primary ecosystem types examined in this assessment can be defined solely in 

terms of their characteristic hydrologic regimes.

Several hydrologic variables are key drivers of wetland and riparian ecosystem 

dynamics. They include whether soils are saturated, deeply inundated, or subject 

to erosive, high energy floods; and the timing of water availability, referring to 

both the seasonality and length of saturation or inundation. These factors act as a 

sieve on the pool of potential species capable of occupying the system, allowing 

only those species with the required physiological, morphological, or life history 

adaptations to establish and be recruited (Keddy 2000). Lastly, the source of water 

supporting a wetland or riparian area (i.e., whether it is principally precipitation, 

surface, or groundwater) and the flow paths taken before and after entering the 

system can affect a variety of important ecological functions. For instance, the 
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availability of mineral ions such as Ca2+, and pH influences species composition in many fens (Bedford 1996; Boeye and others 1994; Jeglum 1971; 

Tahvanainen and others 2003). Likewise, marsh water chemistry varies depending on the influence of local and regional groundwater flow patterns (LaBaugh 

1987; Nicholson 1995; Stewart and Kantrud 1971).

The amplitude and frequency of water table variation is an important difference 

between wetland and riparian ecosystem types. Inundation can result in plant 

mortality (Amlin and Rood 2001; Friedman and Auble 1999), as can dry conditions resulting from low water tables (Shafroth and others 2000). In flooded areas, 

periodic drawdowns can expose substrates and allow new plant establishment 

either from the soil seed bank, as occurs in many marshes (Haukos and Smith 

1994; Smith and Kadlec 1983; Wilson and others 1993), or through seed rain, 

as occurs in many riparian areas (Auble and Scott 1998; Karrenberg and others 

2002). Water table depth variation differs among wetland types. For example, 

fens have relatively low inter-annual water table depth variation compared to 

marshes. Likewise, fen water tables are more stable over annual and decadal 

scales—a necessary condition for peat accumulation. Precipitation events that 

allow playas on the southern plains to fill with water are highly unpredictable 

from year to year, and species composition is strongly controlled by seed bank 

dynamics (Haukos and Smith 1993). In contrast, less hydrodynamically variable 

fens are typically dominated by long-lived clonal perennial species such as Carex

aquatilis (Chadde and others 1998; Cooper 1996; Cooper and Andrus 1994).

Hydrologic processes supporting wetlands and riparian areas can be complex. 

Some wetland complexes have multiple, nested groundwater flow systems (Winter 

and others 2001). Wetlands occur where local variation in subsurface hydraulic 

conductivity and permeability form a perched water table, or where a change in 

slope causes groundwater to discharge to the surface (Winter and Woo 1990). 

Wetlands dominated by groundwater versus surface water and precipitation may 

respond differently to climatic variability (Winter 2000). Wetlands that are dependent primarily on precipitation for their water supply, such as playas, are highly 

vulnerable; while those that are dependent primarily on discharge from regional 

groundwater flow systems, such as fens, are the least vulnerable (Winter 2000).

The affinity or tolerance of individual species for particular hydrologic conditions is highly variable both within and among wetland types. The wide variety 

of physiological and life history adaptations exhibited by wetland plants, coupled 

with the hydrologic variability characteristic of most wetlands and riparian areas, 

drives the “individualistic” (sensu Gleason 1926) response of wetland communities to natural and anthropogenic changes (van der Valk 1981).
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Disturbance Regimes __________________________

Disturbance regime also influences wetland and riparian area distribution and 

ecological structure and function. Various definitions of disturbance have been 

developed, some more easily applied to wetland and riparian ecosystems than 

others. In this assessment, we follow Keddy (2000), who defined disturbance 

as “a short-lived event that causes a measurable change in the properties of an 

ecological community.” Two key parts of the definition are the duration of the 

event—short relative to the lifespan of the dominant species in the systems—and 

the presence of a measurable change in some ecological characteristic. Among 

the most commonly examined response variables are plant mortality or reductions 

in biomass. There are several attributes that together comprise an ecosystem’s 

disturbance regime, including the extent, magnitude, frequency, predictability, 

and return rate of disturbance (Table 7). Each component influences ecological 

structure and function and varies in character among disturbance agents and ecosystem types. For example, the frequency and predictability of large-magnitude 

floods along intermittent and ephemeral Great Plains streams is quite different 

than for headwater streams in mountain areas. This is important because floods 

of equal duration or frequency, but of different intensity, may have different 

ecological effects. Disturbance types particularly important to wetlands and 

riparian areas include water level variation, bank erosion, sediment deposition, 

fire, herbivory, ice flows, and wave action (Auble and Scott 1998; Bendix and 

Hupp 2000; Katz 2001; Keddy 2000; Kellogg and Bridgham 2004; Pettit and 

Froend 2001; Scott and others 1997).

Table 7—Commonly measured components of disturbance regimes (source: Sousa 1985)

Disturbance regime

characteristic Definition and example applications

Aerial extent The size of the disturbed area. It is generally the easiest component 

to measure and can be expressed in either absolute (m2 or acres) or 

relative terms (% of wetland).

Magnitude Two central components: 

 Intensity: an objective measure of the strength of the disturbance agent 

(depth or duration of flooding, shear stress, temperature of fire).

 Severity: A measure of the ecological or physical response to the 

disturbance agent (% mortality and % biomass).

Frequency The number of disturbance per unit time. Can be expressed differently 

based on the spatial scale used. Two commonly used approaches are:

 Random point frequency: the mean number of disturbance events per 

unit time at any random point. Often reported as the recurrence or 

return interval.

 Regional frequency: the total number of disturbances that occur in a 

geographical area per unit time.

Predictability Typically measured as the variance on the mean time between 

disturbances.

Rotation rate The mean time required to disturb the entire area in question.
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Ecoregions of Region 2_________________________

Because of the large geographic scope of this HRV assessment, the ecoregional 

classification originally developed by Bailey (1980) provides an effective structure 

to examine historical and current patterns of wetland and riparian distribution, 

form, and function. The classification is hierarchical in nature and recognizes 

that, since climate ultimately governs energy and moisture gradients, differences 

in continental and regional climatic regimes and gross physiography act as the 

primary controls over more localized ecosystems (Bailey 1980).

At the broadest scale, Bailey’s classification identifies three ecoregional levels: 

Domains, Division, and Provinces. Domains are subcontinental areas of broad 

climatic similarity, while Divisions are subdivisions of domains delineated on 

the basis of water deficit, which differentiates, for example, the “Dry Domain” 

into semi-arid, steppe, or arid desert, which are important controls of many biological and physical processes. Provinces identify climatic sub-zones controlled 

primarily by continental weather patterns such as length of dry season and duration of cold temperatures as well as similar soil orders and overlap considerably with areas of similar potential natural vegetation such as those mapped by 

Kuchler (Kuchler 1964). At finer scales, geomorphic processes, topography, and 

stratigraphy, which influence hydrologic function, soil-forming processes, and 

the potential natural community, are used to delineate Sections. These represent 

broad areas of similar sub-regional climate, geomorphic process, stratigraphy, 

geologic origin, and topography. Fifteen Sections have been identified in Region 

2. Although finer hierarchical scales have been identified, because of the large 

geographic scope of this assessment, Sections will be the smallest level discussed. 

Characteristic wetland and riparian types present in each Section are discussed 

in subsequent chapters.

Human History of the Region ____________________

Native American history as it pertains to HRV—Although the magnitude of 

anthropogenic changes to ecological systems during the Twentieth Century has 

no historical precedent, humans have long been effective at altering their environment. Compared to much of Africa, Asia, and Europe, humans are relatively 

recent components of North American ecosystems, with conventional wisdom 

suggesting the arrival of humans via a Bering Sea land bridge approximately 

13,500 years before present (BP) (Fiedel 1999). Paleo-Indian hunters may have 

contributed to the large-scale extinction of large mammal species at the end of 

the Pleistocene (Alroy 2001; Haynes 2002). How humans impacted ecological systems remains a point of controversy among anthropologists, historians, 

and ecologists. The influence of Native Americans on ecosystems is a critical 

component of all HRV assessments since these assessments typically assume 

that the variability of processes and conditions being described occurred under 

little or no human influence prior to the arrival of Euro-Americans (Landres and 

others 1999).
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There are two competing views about Native American influence on ecosystems. 

One emphasizes the importance of Native Americans in altering their environments (Kay 1994), for example, through effects on fire regimes (Delcourt and 

Delcourt 1997; Denevan 1992; Mann 2005; Pyne 1982). Native American burning 

has been demonstrated to be at least locally important in most temperate North 

American regions, from the pine barrens of the Northeast (Lorimer and White 

2003) to the American Southwest (Fulé and others 1997), and from the forests 

of Great Lakes region (Clark and Royall 1995) to the Black Hills (Brown and 

Sieg 1996).

A competing viewpoint argues that Native American alteration of ecosystems is 

overstated, and that with certain exceptions, Native Americans had minor impacts 

on pre-settlement landscapes. The so-called “myth of the humanized landscape” 

(Vale 1998) has been argued most effectively for remote western landscapes where 

aboriginal population densities and technologies are presumed to have been too 

low to have resulted in widespread or persistent ecological impacts (Baker and 

Ehle 2001; Vale 1998). An additional question, important to this assessment, is 

whether the most common forms of Native American influence, such as altered 

fire regimes, were important drivers of wetland and riparian area structure and 

function.

There is little evidence of direct aboriginal impacts on wetland and riparian ecosystems in the present Region 2 area. Unlike Native cultures in Central America, 

Mexico or the Eastern United States, Native Americans on the Great Plains and 

Rocky Mountains were largely nomadic, with cultures based largely on big 

game hunting. Unlike woodland cultures or the ancestral Puebloan people in the 

Southwest, Native Americans in the western Great Plains and Rocky Mountains 

practiced no significant agriculture. Instead, they tracked large animals and 

moved from favorable over-wintering locations from the base of the Rockies to 

the high country and back again seasonally (West 1998). This migration pattern 

was practiced for thousands of years, as supported by archeological evidence of 

large game drives throughout the region (Benedict and Olsen 1978). Though these 

original inhabitants would undoubtedly have utilized streams as water sources 

and travel ways, the difficulties of a hunting and gathering lifestyle probably 

demanded that groups remain relatively small, thereby minimizing the overall 

impacts to wetlands and riparian systems.

The arrival of Europeans on the continent had immediate impacts on Native 

peoples, even those located far from newly established colonies (Mann 2011). 

Cholera, small pox, measles, and venereal diseases decimated native populations. 

Denevan (1992) estimated that from 1492 to 1800, the North American native 

population dropped from 3.8 million to approximately 1 million—a decline 

of nearly 75%. A second destabilizing factor that influenced Native societies 

was the introduction of new technologies, most notably guns, iron-works, and 

horses (West 1998). These greatly altered traditional ways of life and catalyzed 

significant changes in the ways Native tribes interacted among themselves, with 

Euro-Americans, and with their environment (Carlson 1998; Flores 2001).
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Though accompanying the earliest Spanish expeditions during the Sixteenth 

Century, it wasn’t until after the Pueblo revolt of 1680 that large numbers of 

horses began to appear among Native Americans (Beck and Haas 1989). It is 

difficult to overstate the effect of horses on Native American economic and 

social systems. As the number of horses available through trade or raiding increased during the Eighteenth Century, several Indian tribes embraced the new 

technology. An example is the western or plains Sioux, comprising two broad 

divisions, the Nakota and Lakota. From their original territory in the Great Lakes 

region, they emigrated to the Minnesota prairies, then on to the Missouri River 

Valley. By the late Eighteenth Century, they had expanded west and south into 

the major bison ranges, and by 1830, they had become the preeminent power in 

the northern plains (West 1998).

In 1680, the Cheyenne abandoned their homes on the upper Mississippi and 

began a long westward migration toward the Minnesota River, then to the James 

River in present day North Dakota. Nearly a century later, in response to attacks 

from other Native American groups and the ravages of old world diseases, they 

began moving westward again, drawn toward the plains by horses and bison. 

They moved quickly through the Dakotas, rapidly acquiring horses as they went, 

thus, completing their transformation to a nomadic culture centered on the accumulation and trade of horses and bison (West 1998).

The Ute Indians historically ranged across present day western Colorado and 

much of the northern Colorado Plateau. Separate bands occupied the mountains 

and the more arid regions to the south and west, with each group migrating 

seasonally to follow food resources particular to their environment. In the late 

Seventeenth and early Eighteenth centuries, Utes obtained horses and livestock 

from the Pueblo peoples of northern New Mexico (Beck and Haas 1989), initiating a significant cultural transformation. In the years and decades that followed, 

the Utes became increasingly involved in raiding and trading horses, becoming 

a critical conduit for the dispersion of the horse northward.

Though of obvious historical significance, it is unclear how the changes resulting from Native American adaptation to horses would have affected wetlands 

or riparian ecosystem. West (1998) estimated that six horses per person were 

minimally needed to sustain the nomadic plains tribes with nearly a dozen required for a more secure existence. It follows that even relatively small groups 

would have had sizable herds in need of forage, which would have been most 

abundant along river bottoms, in riparian zones. Large groups would have had 

even greater potential to degrade sites, particularly when stays were long, such 

as when groups over-wintered. One trader noted in his accounts a gathering of 

approximately 6000 Lakota and Cheyenne Indians along the Arkansas River, 

which could have included over 25,000 horses (West 1998). The demand for 

firewood and forage year after year would likely have affected the structure of 

cottonwood stands, inhibiting seedling establishment and recruitment.
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The greater hunting efficiency allowed by the horse led to increased participation 

by Native Americans in the commercial trade of buffalo robes (Flores 2001). 

However, this period of relative abundance was short lived. Several factors led to 

the nearly complete depopulation of bison on the Great Plains. Although estimates 

of the original Great Plains bison population widely vary from approximately 

30 million animals (Flores 2001) to nearly 70 million (Beck and Haas 1989), by 

1890, only a few hundred remained. With the disappearance of bison, the complex social, economic, and ecological systems that had characterized the plains 

came to an end, with an uncertain but probably measurable impact on wetland 

and riparian ecosystems.

The fur trade—Until the middle part of the Nineteenth Century, when gold was 

discovered in California and later the Colorado Rockies, the most significant 

European and American presence in the Rocky Mountains and Great Plains 

was associated with the fur trade. Even at its peak, the number of trappers was 

relatively low; however, trappers were catalysts for enormous social, economic, 

and ecological change, with a legacy lasting to the present day (Beck and Haas 

1989; Sandoz 1978).

Although hides from a variety of animals, including bison, deer, river otter, 

muskrat, raccoon, and fox were traded, it was the trapping and trade of beaver 

pelts that drove much of the early Euro-American exploration and exploitation 

of the region (Sandoz 1978). Demand for beaver pelts was driven primarily by 

the western European hat-making industry, although beaver were also valued 

for castoreum, an alkaloid produced by their perineal scent glands used in the 

manufacture of perfumes and medicines (Wishart 1992). Native Americans were 

a major source of beaver pelts for the Canadian and Great Lakes fur trade, which 

flourished beginning in the late Seventeenth Century (Sandoz 1978). However, it 

was not until favorable reports from Lewis and Clark’s 1804 to 1806 expedition 

that significant numbers of trappers moved to the Rocky Mountains (Wishart 

1992).

In 1807, Manuel Lisa organized the first American trading and trapping expedition to the northern Rockies, ascending the Missouri River and establishing 

a trading post at the junction of the Bighorn and Yellowstone Rivers (Sandoz 

1978). Two years later, Lisa took on new partners, forming the St. Louis Missouri 

Fur Company, and later, the Missouri Fur Company. Their activities, along with 

those of the American Fur Company (founded in 1808 by John Jacob Astor), 

were generally confined to the Missouri River watershed, but likely would have 

affected streams on what are now the Shoshone and Bighorn National Forests.
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Extensive exploration and trapping in the Central and Southern Rockies, including 

many future National Forests, developed later. In the 1830s, trappers such as the 

legendary Kit Carson worked their way up the Arkansas and Rio Grande valleys 

of Colorado from settlements in Taos and Santa Fe, undoubtedly reaching into 

several present day National Forests. By the late 1830s, two factors drastically 

reduced the trade in beaver pelts: (1) fashion trends in men’s hats had shifted 

from beaver to silk, causing a plunge in the market value of beaver pelts; and 

(2) more importantly ecologically, the relentless competition of the late 1820s 

and early 1830s decimated beaver populations, first in more accessible areas of 

the central Rocky Mountains, then in other parts of the range (Wishart 1992).

In many portions of the West, large numbers of beaver were trapped. William 

Ashley wrote of Jed Smith’s visit to a trading post on the upper Columbia River 

that, “...it appeared from the account, that they had taken in the last four years 

within that district eighty thousand beaver, equal to one hundred and sixty 

thousand pounds of furs.” Later, Ashley noted, “You can form some idea of the 

quantity of beaver that country once possessed, when I tell you that some of our 

hunters had taken upwards of one hundred in the last spring hunt out of streams 

which had been trapped, as I am informed, every season for the last four years” 

(Dale 1941). Although difficult to identify precisely, the ecological impacts of 

such intense levels of trapping would likely have been great considering the 

large role beaver play in regulating the geomorphic and ecological attributes of 

riparian systems (Brown and others 1996; Butler and Malanson 1995; Naiman 

and others 1988). Beaver populations have recovered in many areas, but total 

population estimates across the species’ range are estimated to be significantly 

less than prior to the arrival of Euro-Americans on the continent.

Military, scientific, and commercial surveys—The earliest formal exploration by Europeans of the region was undertaken by the Spanish. The purposes 

of their travels varied but included practical objectives like identifying mineral 

resources and finding ways to link their California and New Mexican settlements, 

as well as political ones, such as subjugating rebellious natives and defending 

their claims against the incursions of British and French interests (White 1994). 

Although explorers such as Vargas (1694), Ulibarri (1706), Villasur (1720), 

Rivera (1765), and Dominguez (1776) entered into present day Colorado (Noel 

and others 1994), they recorded minimal information on the natural ecosystems 

they passed through.

Formal American government exploration was similarly driven by a mix of 

political and economic interests. Both Lewis and Clark’s and Zebulon Pike’s 

expeditions were commissioned to explore and assert United States control of 

the new territories acquired under the Louisiana Purchase (Noel and others 1994). 
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In addition to increasing knowledge of the region’s geography and its suitability 

for settlement, these expeditions and others that followed, such as Stephen Long’s 

1820 expedition up the Platte River drainage to the Colorado Front Range, made 

some of the first detailed observations and collections of natural history information (Evans 1997). The 1840s and 1850s brought a number of Government 

commissioned expeditions that greatly expanded American consciousness of 

and enthusiasm for the West (Gilbert 1966; Goetzmann 1960). Among the most 

famous were John C. Fremont’s expeditions, the accounts of which helped spark 

the flow of emigrants to the west coast (Beck and Haas 1989).

Although of historical interest, the value of these explorer’s narratives for our 

assessment is limited. The amount of territory actually traversed by these early 

exploring parties was small relative to the total land area of the region. Since 

exploring parties generally followed rivers, numerous accounts reference riparian 

communities. However, the language in these descriptions is frequently vague 

or focused on more practical issues such as the quantity and quality of forage. 

Perhaps the most significant limitation of these narratives is there is no means 

to extrapolate to wider areas. For instance, differing accounts of the amount of 

tree and shrub cover along plains streams can be found, making more general 

assessments regarding vegetation difficult.

Euro-American settlers—The timing of Euro-American settlement varied across 

the region, but generally was late relative to areas on the west coast. Beginning 

in the 1840s, the number of Euro-American emigrants moving through the Great 

Plains and over the Rockies increased dramatically, particularly following the 

discovery of gold in California. However, most only passed through the region 

and probably had a relatively minor impact on the area.

The first wagon train of pioneers en route to the west coast crossed the continental divide at South Pass in 1841, one year prior to the departure of John C. 

Fremont’s expedition along the same route (Beck and Haas 1989). This route, 

the Oregon-California trail, became the most important passage for emigrants 

traveling west, and it was also utilized by Mormons making their way toward 

Utah (Beck and Haas 1989). During the next 18 years, over 300,000 persons 

traveling with at least 1.5 million oxen, cattle, sheep, and horses traveled up 

the Platte Valley (West 1995). Though this route had been employed by Native 

Americans, trappers, and traders for decades, the greatly increased number of 

travelers severely taxed resources along the river valleys—a problem exacerbated 

by the arrival of drought conditions in the late 1840s (West 1998). The cumulative impacts to riparian communities from harvesting fire wood and grazing 

were likely significant but fairly localized. Since most travelers did not range 

far from the main trail, their impacts to current Region 2 National Forests were 

likely negligible.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 37

More significant changes followed the discovery of gold in the South Platte River 

along Colorado’s Front Range in 1858. An estimated 100,000 people traveled to 

Colorado the following year, beginning a period of rapid transformation for the 

region (Noel and others 1994). Within 20 years, mining camps were common 

throughout the region. Reports of gold in the Black Hills of South Dakota in 

1874 by General George Custer’s expedition caused a similarly large influx of 

prospectors (Froiland 1990; Horsted and Grafe 2002), with the populations of 

both states dramatically increasing over the next several decades. Two types of 

mining operations occurred: placer mines, which recovered minerals in surface 

deposits such as stream alluvium, and lode or hard rock mines (Wohl 2001). Both 

kinds of operations evolved over time and resulted in different kinds of impacts 

to wetlands and riparian ecosystems.

To supply the burgeoning population of miners, numerous towns and cities 

developed on the plains. Boulder and Denver prospered as agricultural and 

trade centers and rapidly expanded as the number of miners entering the region 

increased (Noel and others 1994). To meet the new demand for water by cities 

and agriculture, the first of many irrigation and water control projects in the region was developed. The appropriation of water for municipal and agricultural 

uses proceeded rapidly, with ditches established as early as 1859 in several 

Front Range locations (Jones and Cech 2009; Wohl 2001). The first significant 

reservoir was completed in 1869 to supply water to the expanding network of 

ditches in the region, and by 1882, irrigators had established the first major diversion over the continental divide (Wohl 2001). These trends would continue 

throughout the Nineteenth and early Twentieth Centuries, as demand for water 

and storage increased. Collectively, these developments would become a major 

factor influencing wetland and riparian ecosystems.

Ranching also developed rapidly during the Nineteenth Century. Driven by local 

demand for meat from local populations as well as the possibility of high profits 

in the eastern markets, ranchers introduced thousands of head of cattle to present 

day National Forests (White 1994). Large herds of longhorn cattle were brought 

from Texas and released to graze in meadows and, each fall, were rounded up 

and driven to more favorable winter range in the lower valleys. Large herds of 

sheep were common in the Forests bordering New Mexico, and were introduced 

more widely through the region in the 1890s leading to significant conflict with 

cattle ranchers (Sypolt 1974).

The development of the railroads in the western United States precipitated additional changes to the region. Between 1855 and 1900, over 26,000 miles of track 

were constructed—nearly a 3,000 percent increase in total rail length from 1855 

(Robertson 1991). This expansion mirrored the rapid population growth occurring in the region and hastened the end of Native American presence outside of 

reservations. In addition to the social and economic impacts, railroad construction also impacted riparian areas. Rail lines were frequently built parallel to the 

major rivers, requiring that armoring such as rip-rap be applied to protect the 

new lines, thereby altering the natural channel dimensions (Wohl 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 38

Early concerns about the sustainability of resource use in the West led Congress to 

pass the General Revision Act in 1891, which allowed the President to proclaim 

forest reserves (Bradford 1998). Several forest reserves were subsequently created making them precursors to modern National Forests. It wasn’t until 1905 

that Congress officially established the United States Forest Service (USFS), 

transferring the forest reserves from the General Land Office to the Department 

of Agriculture. During the dustbowl drought of the 1930s, Congress passed the 

Bankhead-Jones Farm Tenant Act, authorizing the Soil Conservation Service to 

purchase sub-marginal farm land. Responsibility for these lands shifted among 

several different Federal agencies until 1954, when it was transferred to the 

USFS, resulting in the formation of the new National Grasslands.

Overview of Anthropogenic Impacts to Wetlands and 

Riparian Ecosystems___________________________

Wetlands and riparian areas are among the most heavily disturbed ecosystems 

in North America. It has been estimated that since the Late Eighteenth Century, 

35 to 40 percent of the wetlands have been lost in South Dakota, Nebraska, and 

Wyoming, with wetland losses in Colorado and Kansas approaching 50 pecent 

(Dahl 1990). Since much of the wetland losses have resulted from agricultural 

activities such as draining and tilling, a minor historical land use for most Forest Service units, it is likely that these estimated rates of wetland loss are higher 

than has occurred on USFS managed lands.

A variety of anthropogenic factors have altered the structure and function of 

wetlands and riparian ecosystems, including many found in Region 2. Several 

general concepts that have been developed are useful for describing how anthropogenic changes may have impacted ecological systems. Ecological integrity has 

been defined as the ability of a system to support and maintain a “...balanced, 

integrated, adaptive community of organisms having a species composition, 

diversity, and functional organization comparable to the natural habitat of the 

region” (Karr and Dudley 1981). Alternatively, ecological condition has been used 

to describe the extent to which a given site departs from full ecological integrity 

and can be defined as the relative ability of a wetland to support and maintain its 

complexity and capacity for self-organization with respect to species composition, physio-chemical characteristics, and functional processes (Fennessy and 

others 2004). These concepts provide a means of assessing differences between 

wetlands of a similar class with and without the effects of human alterations—a 

key objective in this HRV assessment.

The value attributed to wetland and riparian ecosystems is largely based on the 

various ecological, biogeochemical, and hydrologic functions they provide. 

These functions are numerous and varied, ranging from maintenance of species 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 39

diversity to biogeochemical cycling, and differ depending on the type of wetland 

and its landscape context. Although management and regulation generally strive, 

implicitly or explicitly, to preserve or restore ecological function, there is still 

considerable uncertainty on how to best identify and measure function. Numerous 

systems have been developed, such as the Hydrogeomorphic Method (HGM) 

(Brinson 1993a; Hauer and Smith 1998), that are applied by wetland managers.

Central to the idea of managing ecosystems within their historical range of variation is that, by maintaining or restoring the ecological integrity of ecosystems, 

managers are most likely to preserve or enhance important functions. Also, most 

stressors affecting ecosystems do not occur in isolation. Departures from the 

historical range of variation for many systems may be caused by the cumulative 

effects of several different factors, such as overgrazing, roads, and invasive species, and may occur abruptly at some threshold rather than gradually.

Some of the most significant anthropogenic changes to wetland and riparian 

ecosystems include hydrologic changes associated with flow regulation and the 

removal of beaver (Graf 1999; Nilsson and Berggren 2000; Wohl 2001), the 

introduction and spread of non-native species (Stein and Flack 1996; Stohlgren 

and others 1999), livestock grazing (Belsky and others 1999; Fleischner 1994), 

and cumulative effects on water and sediment dynamics associated with broadscale changes to watersheds, such as altered fire regimes resulting from land use 

change and fire suppression (Dwire and Kauffman 2003). Winters and others 

(2004) identified a total of 24 anthropogenic activities affecting aquatic, riparian, 

and wetland resources, organized into seven basic groups (Table 8).

Table 8—Anthropogenic activities arranged by primary use categories identified by Winters et al. (2004).

 Mineral Vegetation

Water use Transportation Recreation Biological extraction management Urbanization

Stream Roads Developed Invasive Hardrock Domestic Transmission

Diversions Recreation Species Mining Grazing Corridors

Reservoirs Trails Dispersed Beaver Placer Large Wild Urbanization

 Recreation Remova Mining Ungulates

Ditches Railroads Ski Area Pesticide Energy Commercial

Development Development

Transbasin Off Road Timber

Diversions Vehicle Use Harvest

Spring Natural Fire

Development

 Prescribed

 Fire

Tie Drives
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 40

The relative importance of these activities varies by wetland type and by National 

Forest or Grassland. For instance, placer and hard rock mining activities are relevant only in USFS units containing highly mineralized bedrock outcrops and are 

more likely to impact riparian systems than fens. A discussion of anthropogenic 

impacts to wetland and riparian areas is central to our HRV assessment and will, 

therefore, be presented in greater detail in subsequent chapters. The following is 

a broad overview of the major anthropogenic activities, following the structure 

laid out by Winters and others (2004).

Water developments

The development of water storage and diversion structures closely followed 

Euro-American settlement of the region. Although most initial developments 

were for agriculture and mining activities (Wohl 2001), there has been in recent 

years a trend toward increased municipal use (Wilkinson 1992). Most of the 

major rivers flowing through Region 2 are dammed, and numerous transbasin 

diversions have been constructed to augment natural stream flows to areas with 

large agricultural or municipal demand such as the South Platte basin. In addition to developments affecting large perennial rivers, many intermittent and 

ephemeral wetlands on the plains were altered by the construction of livestock 

watering impoundments (Smith and others 2002).

Groundwater extraction for irrigation purposes affects some wetland and riparian 

ecosystems (Cooper and others 2003c; Elmore and others 2003; Stromberg and 

others 1996). For example, there has been increasing concern about the effect 

of center-pivot irrigation in the Nebraska Sandhills, which uses groundwater 

extracted from the regional groundwater system (Bio/West 1986; Novacek 1989). 

However, since irrigated agriculture is an uncommon land use in most Region 2 

National Forests and Grasslands, the impacts of groundwater extraction on USFS 

lands is likely small and limited to localized areas in the immediate vicinity of 

pumping wells.

Water developments can have a wide variety of geomorphic, hydrologic, and 

ecological effects on wetland and riparian ecosystems. Physical changes include 

alteration of stream channel morphology (e.g., width/depth ratio and sinuosity), 

rates of sediment production or mobilization, and changes in the timing and 

magnitude of floods (Brandt 2000; Hirsch and others 1990; Kondolf 1997). Since 

flooding is one of the principal disturbances affecting vegetation in riparian 

areas and a primary control on the water table dynamics controlling vegetation in some other wetland types, such as marshes found in oxbows along river 

floodplains, these hydrologic changes can have profound effects on the patterns 

of plant establishment and mortality (Nilsson and Berggren 2000; Patten 1998). 

A variety of studies have demonstrated that altered flood regimes can reduce 

opportunities for the establishment of native species in the Genera Populus

and Salix, and favor the spread of non-native species such as Tamarix spp. 

(Cooper and others 1999; Shafroth and others 1998) and Elaeagnus angustifolia

(Russian-olive) (Katz 2001; Lesica and Miles 1999).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 41

Water developments can alter the amount and distribution of wetland and riparian ecosystems. For example, reservoirs formed behind large dams may destroy 

riparian communities, resulting in an overall decrease in riparian area. However, 

marsh communities may form along reservoir margins, resulting in a net increase 

in the areas occupied by marsh vegetation. Another example is the large number 

of livestock water impoundments constructed along intermittent streams in many 

National Grasslands. These can create marsh type habitats that would otherwise 

be rare or absent.

Herbivory

Wild ungulates—Prior to the arrival of Euro-Americans, most areas of Region 2 

supported native ungulate species including elk, deer, moose, bison, bighorn 

sheep, and pronghorn antelope. However, the importance of native grazers to the 

region’s ecology, particularly wetland and riparian areas, is unclear. Because of 

their great abundance, pre-Euro-American bison populations (Flores 1991) almost 

certainly had significant impacts on wetlands. For instance, bison were observed 

to use playas as wallows (Smith 2003) and, like domestic livestock, seasonally 

utilized river bottoms for forage and shelter during inclement weather. Though 

impossible to quantify the relative effects, these activities would likely have affected both the composition and structure of wetland and riparian communities.

Although there are no reliable estimates of pre-Euro-American period elk and deer 

populations, these species would have been expected to utilize, at least seasonally, 

riparian and wetland areas. Behavioral and dietary studies of contemporary elk 

herds, for example, have demonstrated that wetland and riparian species such as 

willows (Salix spp.) and sedges (Carex spp.) are important components of native 

ungulate diets (Hobbs and others 1981; Hunt 1979). Several studies have shown 

that elk herds are capable of significantly altering riparian communities, affecting 

the structure and regeneration of willows and cottonwood (Populusspp.) (Beschta 

2003; Brookshire and others 2002; Peinetti and others 2002b; Zeigenfuss 

and others 2002). However, these examples are not likely representative of preEuro-American settlement conditions, as contemporary elk population densities 

and foraging behaviors developed in the absence of significant predators. Other 

ungulates, notably moose, heavily utilize riparian areas (Oedekoven and Lindzey 

1987) and can have large impacts on vegetation, at least locally.

Though large wild ungulates can impact resources in similar ways as domestic 

livestock, wild ungulates tend to be more dispersed across the landscape, and therefore less likely to drive vegetation dynamics in wetland and riparian ecosystems. 

However, since populations of native ungulates are often intensely managed and 

most systems lack significant predator densities (excluding humans), the effects 

of contemporary native ungulate populations may not be representative of pre 

Euro-American settlement impacts, and their impacts on wetland and riparian 

systems outside the HRV.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 42

Livestock—Livestock grazing represents one of the most widespread historical 

and current land uses in Region 2. Domestic animals, including horses, sheep, 

and cattle, accompanied early European explorers and settlers. The horse, in 

particular, was rapidly and widely adopted by Native Americans, greatly altering 

their social systems and relationship with their environment (Flores 2001; West 

1998). With the exception of horses and the raising of sheep by the Navajo in the 

Southwest, Native Americans did not widely adopt domestic livestock. However, 

with increased Euro-American settlement during the Nineteenth Century, there 

was a large increase in the number of livestock moving through the region (White 

1994). The increased demand for commodities in the east and the expanding 

settlements in the west, coupled with increased efficiencies in transporting and 

marketing those commodities that was achieved with the spread of railroads, 

led to an explosion in livestock numbers, and within only a few years, many 

rangelands were seriously overstocked (Pieper 1994). The creation of public land 

management agencies like the USFS, increasingly led to greater management 

livestock populations, and for most USFS units, current livestock numbers are 

significantly lower than historical levels (Elmore and Kauffman 1994).

Livestock can impact wetlands and riparian areas in a variety of ways. Overgrazing can lead to significant changes in vegetation structure, negatively impacting wildlife and fish habitat (Fleischner 1994; Knapp and Mathews 1996; 

Schulz and Leininger 1991). Livestock grazing can reduce seed production and 

establishment by native vegetation (Brookshire and others 2002), affecting the 

successional dynamics of riparian ecosystems. Changes in species composition, 

increases in the incidence of exotic species, and changes in channel morphology 

have also been documented (Belsky and others 1999; Chaney and others 1990; 

Fleischner 1994). Moderate grazing can also have positive effects, particularly 

in ecosystems that have an evolutionary history of grazing. Grazing effects and 

potential departures from the HRV of wetlands and riparian areas are discussed 

in greater detail in subsequent chapters.

Invasive species

The introduction and spread of invasive and non-native species is one of the most 

significant changes to ecosystems across North America. Several different terms 

have been developed to describe non-native species. Weeds are defined as any 

plant that poses a major threat to agriculture and/or natural ecosystems; alien 

species are species introduced and occurring in locations beyond their known 

historical range; and invasive species are species that demonstrate rapid growth 

and spread and are capable of invading habitats and displacing native species. It 

is important to note that some species such as cattails (Typha spp.), though native 

to a particular area, may become invasive if environmental conditions change, 

for example, following disturbance. Likewise, some alien species, though well 

established in an area, may have few deleterious effects on native ecosystems.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 43

Alien and invasive species can have a variety of effects on native ecosystems. 

These include hydrologic and geomorphic changes, as has been documented 

with salt cedar (Tamarix ramosissima) (Busch and Smith 1995; Tickner and 

others 2001), reductions in native species cover and diversity due to competitive 

exclusion (Hager 2004), and reduced forage for wildlife and livestock. Invasion 

by exotic shrubs can alter vegetation structure, by decreasing, or, in the case of 

Tamarix spp. and willow flycatchers in the Southwest, increasing the suitability 

of habitat for birds (Graf and others 2002; Stevens and others 2001; Stoleson 

and Finch 2001). The introduction of non-native species can significantly alter 

natural disturbance regimes, with large implications not only for biodiversity but 

for a range of additional ecological characteristics like biogeochemical cycling 

rates (Mack and D’Antonio 1998; Mack and others 2001). Changes in nutrient 

availability associated with invasions of nitrogen-fixing plants like Elaeagnus 

angustifolia can promote the spread of additional exotics (Simons and Seastedt 

1999), causing positive feedback that leads to increased degradation of ecosystem integrity.

The distribution and abundance of exotic species varies considerably based on the 

life history and physiological tolerance of specific invasive species as well as on 

a variety of ecological characteristics that influence ecosystem vulnerability to 

invasion such as fertility and disturbance frequency. In general, many non-native 

species respond positively to disturbance, which is one of several reasons for 

the higher incidence of exotics associated with roads and other disturbed environments (Gelbard and Belnap 2003; Parendes and Jones 2000). Exotic species 

are often more common in areas with high diversity. For example, Stohlgren 

and others (1999) found higher numbers of exotic species in communities with 

higher numbers of native species such as wet meadows. Similar patterns of exotic 

species cover have been observed in riparian areas as well (Stohlgren and others 

1997; Stohlgren and others 1998).

Mineral extraction

Mineral extraction activities have been an important factor in the human and 

ecological history of Region 2. Initial Euro-American settlement patterns in the 

region were driven by the discovery of gold, which determined where many 

contemporary cities and towns were initiated and the patterns of public land 

ownership on many National Forests. Mining activities have had a variety of 

impacts on wetland and riparian ecosystems, and though the extent of mining 

activities has generally decreased relative to historical levels, mining is still 

relatively common in the region.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 44

Placer and hard rock mining have occurred in the region. The mining of placer 

deposits, which can be conducted via several techniques, including hydraulic 

mining and dredging, represents a direct disturbance to riparian areas and adjacent 

wetlands. Impacts included direct removal of vegetation and the alteration of 

the stream channel shape and hydraulics through placement of tailings deposits. 

When evaluating the effects of past placer mining, it is important to know the 

amount and distribution of mined areas and the specific methods used to recover 

mineral deposits (Winters and others 2004). An additional form of surface mining 

of importance to wetlands and riparian areas is aggregate mining for construction 

use. Alluvial gravel and sand deposits are mined directly and can cause a variety 

of hydrologic, geomorphic, and ecological effects. For instance, the mining of 

alluvium within active stream channels can cause channel incision, which may 

then propagate up- and downstream of the mine, reducing overall channel instability and lowering alluvial water tables (Kondolf 1997; Scott and others 1999). 

Gravel mining can result in either an increase or decrease in total wetland area 

resulting from the conversion of riparian forests to large open water gravel pits, 

which can support a narrow fringe of marsh communities along their boundaries.

Hard rock mining can occur in both subterranean and in open pits and was widespread in mineralized areas of Region 2. Though a celebrated part of the region’s 

history, in many instances, mining has left a toxic environmental legacy, and 

abandoned mines represent one of the greatest environmental problems plaguing 

the region. One of the most common problems is the degradation of water quality 

due to heavy metal contamination. The magnitude and nature of the problems 

and their effect on wetland and riparian resources can vary depending on the 

specific ores mined and the processing methods used. Additional mining impacts 

can include transport and deposition of tailings downstream of mine sites and 

the release of acidic and metal-laden runoff from tailing piles, shafts, and adits 

(Clements and Kiffney 1994; Howe and Noble 1985).

Development of energy resources, specifically oil, natural gas, coal, and, most 

recently, coal-bed methane, can also alter wetlands and riparian ecosystems. 

Although direct impacts can occur, most impacts are indirectly caused by the 

construction and maintenance of the extensive infrastructure required to access, 

store, and distribute the commodity. Impacts include chemical contamination, 

increased erosion and sedimentation, and hydrologic modification. The relative 

importance of each of these impacts likely varies depending on the specific commodity being extracted; the density of wells, pipelines, and access roads; and the 

specific climatic, physiographic, and geological setting.

Although the technology driving it is of recent origin, coal-bed methane development may have a greater effect on wetlands and riparian systems than 

traditional oil and gas extraction methods. A large amount of water is produced 

during coal-bed methane production, potentially altering the hydrologic regime 

of associated wetlands and streams. Changes in the magnitude and timing of 

streamflow can alter vegetation patterns by decreasing opportunities for native 

plant establishment or by facilitating the spread of exotic species and can result 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 45

in greater channel instability. Additionally, groundwater pumping may lower 

local water tables, negatively affecting dependent wetlands. Also, high levels 

of salts in pumped groundwater have the potential to alter the productivity of 

soils in affected areas.

Transportation

Roads influence many ecosystems in Region 2, including wetlands and riparian 

areas. Historically, roads and other linear transportation features, such as trails 

were constructed along valley bottoms, directly impacting riparian systems and 

associated wetlands. Streams were often channelized and wetlands were filled 

to remove obstacles that blocked road construction. The relatively high road 

densities found on many National Forests and Grasslands can have a variety of 

direct and indirect impacts on native vegetation. These can work independent of 

or synergistically with other agents of change, such as exotic species in affecting 

riparian and wetland ecosystems.

Transportation networks can have significant effects on local and watershed-scale 

hydrologic processes. Roads and trials and their associated engineering structures 

such as culverts, dips, and ditches, can significantly alter natural drainage patterns (Forman and Sperling 2002). Interception and infiltration rates generally 

decrease due to the removal of vegetation and soil compaction or paving, altering the hydrologic response of basins to both annual snowmelt runoff episodes 

and isolated convective storm events (Jones 2000; Jones and others 2000). The 

increased prominence of overland flow processes can cause more rapid and 

intense hydrograph response to precipitation events, resulting in greater stream 

energy and erosion potential.

Transportation networks can have a variety of additional effects on wetlands, 

including the introduction of pollutants from vehicle exhaust or fluid leaks 

(Havlick 2002) and the alteration of water chemistry (e.g., conductivity, cation 

concentrations, and pH) due to road dust, increased sediment deposition, and 

chemicals used in road maintenance, such as de-icing agents (Trombulak and 

Frissell 2000; Wilcox 1986a; Wilcox 1986b). A number of variables that can influence the effects of roads, including road density, road slope, and surface type 

and the number, size, and design of engineering structures. Since these can vary 

so greatly within and among National Forests, formulating general statements 

regarding their effect on wetlands and riparian systems is difficult.

Railroads are an additional component of the transportation infrastructure with the 

potential to affect wetlands or riparian areas (Winters and others 2004). Effects 

can include some of the same kinds previously described for roads, such as alteration of surface and subsurface water flow. In addition, water contamination can 

occur from leaking train cars, maintenance equipment, or creosote-impregnated 

railroad ties (Winters and others 2004). However, railroad densities are generally 

much lower than those of roads, so their direct effect on the HRV of wetlands 

and riparian areas at landscape and regional scales is likely small.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 46

Recreation

A variety of recreation activities impact riparian areas and wetlands. The magnitude and nature of impacts differ as a function of the kind of activity and the 

particular system in question. Although non-motorized, dispersed recreational 

activities such as fishing can have direct effects on vegetation through trampling, 

these effects tend to be localized in nature. In general, developed recreation has a 

greater potential for impacts. For example, the cutting of trees as part of ski area 

development can alter the amount and quality of surface and subsurface runoff, 

indirectly affecting dependent wetlands and riparian areas. In addition, surface 

water diversions for snowmaking at ski areas can reduce available water for riparian areas. The relative importance of such activities varies among USFS units, 

being of greatest importance in areas with large concentrations of developed ski 

areas (e.g., the White River National Forest) and of no consequence in others.

Additional impacts to wetlands from winter recreation can occur from activities such as cross-county skiing, snowshoeing, or snowmobiling. Compaction 

of accumulated snow alters the thermal properties of snow, potentially delaying 

spring melt and altering soil temperature profiles. These effects are again going 

to be localized in nature and of greatest importance to specific wetland types 

such as fens.

Fire

Fire is a natural component of nearly all Region 2 ecosystems, including wetlands and riparian areas. However, the specific effects of fire, both natural and 

anthropogenic in origin, differ greatly depending on a wide range of factors. These 

include variables related to the fire, such as fire intensity and duration, which 

directly impact vegetation through plant mortality (DeBano and others 1998). At 

landscape scales, the relationship of wetland and riparian communities to fire is 

driven largely by the disturbance regimes of surrounding cover types. However, 

the significant differences in vegetation composition, hydrogeomorphology, 

microclimate, and fuels characteristics between wetlands and riparian areas and 

surrounding uplands may contribute to different fire regimes and behavior. The 

frequency of fire may be lower and fire severity may be more moderate than in 

adjacent uplands, but there are documented instances where fires appear to burn 

wetland and riparian areas with comparable frequency (Dwire and Kauffman 

2003).

Because of their small area and the presence, at least seasonally, of saturated 

or inundated soils, fires rarely start in wetlands. Consequently, the fire regimes 

characterizing the surrounding landscape, including those that deviate from their 

HRV due to anthropogenic influences such as fire suppression, strongly influence 

fire regimes in wetlands. Land use and management may affect fire regimes and 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 47

properties through additional mechanisms such as livestock grazing, logging, 

hydrologic alterations such as damming or flow regulation, or the introduction 

of invasive species that change the composition, structure, and distribution of 

fuel loads (Dwire and Kauffman 2003). An example of the latter is the introduction and spread of Tamarisk along many riparian areas in the region, which has 

resulted in significant sifts in fire behavior and fire regimes (Busch 1995; Busch 

and Smith 1995). The cumulative impacts from anthropogenic alterations are 

most likely to influence fire attributes during periods of drought or extreme fire 

weather (Dwire and Kauffman 2003).

Functional differences in attributes such as mean water table depth among wetland types also influences the response of wetlands to fire. For instance, direct 

fire effects are relatively minor in subalpine fens, due in part to their high, stable 

water tables. In contrast, marshes on the Great Plains, which in some years may 

become completely dry, may experience more frequent and severe fires. Many 

wetland and riparian species, including members of the Salicaceae family such 

as Salix spp. Populus spp., are capable of re-sprouting or vegetatively spreading, allowing recovery following both flood and fire disturbances (Mitsch and 

Gosselink 2007). The adaptations to fluvial disturbances may facilitate survival 

and reestablishment following fires, contributing to the rapid recovery of many 

wetland or riparian ecosystems (Dwire and Kauffman 2003).

Vegetation management

Direct vegetation management such as timber harvest occurring in landscapes 

surrounding wetlands and riparian areas may influence their structure and function. For example, timber harvest may alter the quantity and quality of surface 

and subsurface water reaching riparian areas (Stottlemyer and Troendle 1999; 

Stottlemyer and Troendle 2001; Troendle and King 1985). Tree harvest may reduce 

the amount of rain or snow intercepted and subsequently lost by evaporation or 

sublimation and may affect runoff by influencing evapotranspiration (ET) rates 

and promoting increased snowpack accumulation in openings, with affects on 

water yields lasting, at declining levels, for up to 80 years (Troendle and King 

1985). In addition to the effects of commercial logging, many ecosystems have 

been affected by historical utilization of forests associated with Euro-American 

settlement and development. For example, large volumes of timber were harvested for use as railroad ties, charcoal, and mining supports in many parts of 

the Rocky Mountains (Wohl 2001). In some areas, severe impacts to riparian 

areas resulted from the use of tie drives to transport lumber (Wohl 2001; Young 

and others 1994).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 48

Beaver

Few mammals, with the obvious exception of humans, have as large of an ecological impact as the beaver (Castor canadensis). Through their feeding and dam 

building activities, beaver strongly modify local hydrologic regimes and influence a wide range of biotic and biogeochemical processes (Burns and Mcdonnell 

1998; Johnston and Naiman 1990a,b; Naiman and others 1994). In many areas 

of Region 2, the arrival of beaver trappers marked the first significant interaction of native peoples and landscapes with Europeans (Sandoz 1978). However, 

because they primarily affect riparian ecosystems, a more detailed discussion of 

their influence is presented in a subsequent chapter.

Cumulative watershed impacts

Individual anthropogenic impacts may act synergistically on watershed processes 

(Reid 1993; Winter 1988). For instance, the cumulative effects of tree harvest, 

road construction, and fire in a given watershed may result in greater changes 

to water and sediment yields than expected by simply summing the contributions of each individual stressor. Consequently, discussion of cumulative effects, 

presented in subsequent chapters, is warranted for this assessment
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4. Historical Range of Variation for 

Region 2 Riparian Ecosystems

Definitions and Concepts of Riparian _____________

The term riparian has been variously defined and applied. The term originates 

from the Latin word riparius, which means “belonging to a bank or shore” and 

indicates a place on the landscape adjacent to a stream, but is also used to refer to 

areas that are periodically flooded. Some definitions of riparian, such as the one 

used by the U.S. Fish and Wildlife Service, emphasize geographic proximity to 

water bodies, while others are based on either biological or hydrologic criteria. 

An overview of riparian definitions is presented in Chapter 1. The definition 

we apply in this assessment closely follows that used by Naiman and Decamps 

(1997): the portion of the stream channel occurring between the low and high 

water marks and adjacent terrestrial areas extending from the high water mark 

toward the uplands where vegetation may be influenced by elevated water tables 

or flooding. A key element of this definition is the existence of a physical connection, at least intermittently, between the stream and riparian area. In the case 

of ephemeral streams, this connection may be limited to the physical effects of 

infrequent flood events. In contrast, along perennial systems, streams may exert 

a more constant and dominant control on ecological function through flooding 

as well as by influencing water table dynamics.

Structural and functional characteristics vary widely among different riparian ecosystem types, making generalizations difficult, but there are several characteristics 

common to all ecosystems. Most important is the presence, at least periodically, 

of flowing water. Flood events are key drivers of geomorphic, biogeochemical, 

and biological characteristics of riparian areas (Bowden 1987; Knighton 1998; 

Pinay and others 2002; Wohl 2000), and they act to differentiate riparian from 

other ecosystem types supporting mesic or hydric plant species. Riparian areas 

also commonly have elevated water tables relative to adjacent uplands and support distinct vegetation associations (Carsey and others 2003), although this 

is not always true in ephemeral or intermittent systems (Shaw 2006). In many 

arid and semi-arid areas, riparian areas provide the only arboreal habitats in the 

landscape, thus providing essential habitat for birds and other dependent wildlife 

species (Finch and Ruggiero 1993).

Geomorphic Setting and Principal 

Ecological Drivers _____________________________

In contrast to fens, where geomorphic changes are extremely slow, streams and 

their associated riparian zones are highly dynamic. Region 2 riparian areas occur in a variety of geomorphic settings, along high-gradient, bedrock-dominated 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 50

streams to low-gradient, alluvial rivers on the plains. Principal drivers of geomorphic processes at broad temporal and spatial scales include geology, climate, 

and tectonics. At the time scales most relevant to this assessment, these variables 

are effectively independent, driving such dependent processes like vegetation 

composition and structure, channel, morphology, and hydrologic regime 

(Figure 9, Table 9). While at local or reach scales, riparian vegetation is often 

determined by factors such as cross-sectional morphology, flood frequency, 

moisture gradients, and biotic factors such as beaver, these processes reflect the 

influence of macro-scale processes (Baker 1989).

Figure 9—Schematic model illustrating the relationships among geomorphic processes, terrestrial plant succession, and aquatic ecosystems in riparian zones. 

Directions of arrows indicate predominant influences of geomorphic and biological 

components (rectangles) (Modified from Gregory and others 1991).
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Table 9—Scale dependency of variables affecting geomorphic and hydrologic characteristics of riverine 

systems (Knighton 1998).

 Medium time scale Short time scale Instantaneous time scale

Variable 103-104 years 101-102 years <10-1 years

Geology Independent Independent Independent

Climate Independent Independent Independent

Regional relief Independent Independent Independent

Vegetation properties Dependent Dependent Independent

Mean water and Dependent Dependent Independent

sediment discharge

Channel morphology Dependent Dependent Independent

Instantaneous flow Indeterminate/ Indeterminate/

characteristics Irrelevant Irrelevant Dependent

Regional climate characteristics of particular importance for riparian ecosystems 

include the amount, form, and seasonality of precipitation, temperature and 

wind, as well as the timing of snowmelt. These factors influence stream flow 

regimes, including flow variability, base, mean, and peak flow volumes, and 

fluvial geomorphic processes such as sediment erosion and deposition (Wohl 

2000). Climate also directly affects ecosystems through its influence on energy 

and water exchanges between riparian areas and adjacent aquatic and upland 

areas (Naiman and Decamps 1997), influencing basic ecological attributes such 

as primary productivity, plant establishment and mortality, and competition 

(Mitsch and Gosselink 2007).

Basin physiography, including watershed size, relief, and elevation, are also of 

broad importance to riparian areas. For example, orographic effects in mountainous 

terrain strongly influence precipitation patterns (Kittel and others 2002), directly 

affecting stream flow regimes. Watershed size influences a variety of processes 

that are important to riparian areas such as flow volume, flood frequency and 

magnitude, and measures of flood variability. Stream gradient influences the 

relationship between mean and peak flood levels and stream energy, variables of 

particular importance to reach-scale fluvial geomorphic processes and riparian 

vegetation dynamics (Knighton 1998; Wohl 2000).

Other broad-scale factors influencing riparian ecosystems include lithology and 

glacial history (Wohl 2000). Erosion and sediment transport characteristics for 

rivers in the same climate regime may differ depending on whether the dominant 

rock types are igneous or sedimentary. The spatial patterns of flow volume along 

rivers flowing through limestone, for example, can be highly variable due to the 

effect of dissolution features on surface and groundwater flow (Driscoll and others 2000). While no USFS unit in Region 2 was directly affected by continental 

glaciation, periodic expansion of glaciers originating in high mountain cirques 

have shaped many alpine and subalpine landscapes in Colorado and Wyoming, 

creating broad, U-shaped valleys instead of the V-shaped valleys characteristic 

of unglaciated areas (Windell and others 1986).
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Complex interactions over a range of spatial and temporal scales occur among 

hydrologic and geomorphic processes, terrestrial plant succession, and aquatic 

ecosystems (Gregory and others 1991), making the task of isolating the influence 

of individual processes difficult. Many studies of riparian ecology emphasize 

geomorphic processes as key drivers of ecological processes. These create a 

mosaic of landforms such as channels, floodplains, point bars, and in-channel 

islands, which function differently regarding hydrologic regime and disturbance, 

such as flooding (Gregory and others 1991). These functional processes in turn 

drive the spatial pattern and successional development of riparian vegetation 

(Hughes 1997).

Hydrologic, ecological, and disturbance processes

The relative importance of different water sources varies for riparian areas. 

Surface water can originate as overland flow, shallow subsurface storm flow, 

direct precipitation, and recharge from local alluvial aquifers. Water can leave 

riparian systems through direct evaporation, transpiration, overbank flooding, 

and percolation into alluvial or bedrock aquifers (Dunne and Leopold 1978; 

USDA NRCS 1998) (Figure 10). The importance of each water budget component varies among riparian areas. Whether precipitation falls mainly as rain or 

snow influences stream flow regimes and dependent biota. High flows following 

rainstorms may occur hours or even minutes after precipitation events, depending on soil permeability and other watershed characteristics. In contrast, where 

snow is the dominant form of precipitation, melting may occur over a period of 

days or weeks and slowly build the peak flood discharge (Poff and others 1997).

Figure 10—Principal hydrologic pathways in riparian areas 

(FISRWG 1998).
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Vegetation in Region 2 riparian ecosystems is distinct from uplands because 

of the greater availability of water in an otherwise water-limited landscape 

(Goodwin and others 1997). Typically, species composition changes with increasing distance from and elevation above the stream channel due to differences 

in water table depth and flood frequency, although the specific pattern varies 

widely among rivers and between individual stream reaches. Key variables 

driving vegetation dynamics in riparian sites are base and mean flow levels; the 

magnitude, seasonality, and frequency of peak flows; and the relative influence 

of groundwater on water tables.

Hydrologic attributes vary widely among rivers in Region 2 (Figure 11). The 

flow regime of rivers can be examined in terms of magnitude, frequency, duration, timing, and rate of change (Poff and others 1997) (Table 10). Streams in 

mountainous areas with snowmelt-dominated precipitation regimes typically have 

pronounced peak flows during the spring and early summer, corresponding 

with the melt-off of snow from higher basins (for an example, see the hydrograph 

of the Big Thompson River in Figure 11). Although these peak flows can vary 

widely in magnitude depending on factors such as snow-pack depth, temperature, and precipitation (e.g., rain on snow) events, they are relatively predictable 

as far as timing is concerned. In contrast, intermittent and ephemeral streams 

with small basins and precipitation regimes dominated by convective rain fall 

events can be highly variable in magnitude, timing, rate of change, duration, and 

frequency (Friedman and Lee 2002; Shaw 2006).

Figure 11—Typical discharge hydrographs for a range of Region 2 rivers. Rivers originating 

in the high mountains such as the Animas and Shoshone have large, snowmelt-dominated 

peak discharges. In contrast, flood flows along Great Plains rivers such as the Cimarron and 

Cheyenne are driven principally by rain events and are frequently less predictable and more 

flashy in nature. The Niobrara River and other rivers in the Sandhills region of Nebraska 

have continuous groundwater inputs and relatively steady stream discharge year round.
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Table 10—General characteristics of flow regimes in rivers (Poff et al. 1997).

Flow

characteristic Description

Magnitude The amount of water moving past a fixed location per unit time; 

can refer either to absolute or to relative discharge (e.g., the 

amount of water that inundates a floodplain).

Frequency How often a flow above a given magnitude recurs over some 

specified time interval. Frequency of occurrence is inversely 

related to flow magnitude.

Duration The period of time associated with a specific flow condition; can 

be defined relative to a particular flow event (e.g., a floodplain 

may be inundated for a specific number of days by a ten-year 

flood), or it can be a defined as a composite expressed over a 

specified time period (e.g., the number of days in a year when 

flow exceeds some value).

Timing The regularity with which flows of a defined magnitude occur. 

For example, annual peak flows may occur with low seasonal 

predictability or with high seasonal predictability 

Rate of change How quickly flow changes from one magnitude to another; flashy 

streams have rapid rates of change, whereas stable streams 

have slow rates of change.

Streams associated with eolian features (e.g., the Nebraska Sandhills) represent 

one end of flooding and stage variability gradients. Because of the high permeability and hydraulic conductivity of the sands composing landforms, almost no 

surface runoff is formed during precipitation events. Surface water in streams, 

lakes, and wetlands originates as groundwater discharge, and stage and discharge 

levels remain relatively stable seasonally and annually. An example is the Niobrara 

River in the Nebraska Sandhills (Figure 11); water tables are stable and flood 

frequency is low enough to allow for the development of fens adjacent to streams.

Peak flow characteristics are also an important factor influencing riparian vegetation. High magnitude floods are important agents of geomorphic change and 

represent a major disturbance agent in riparian ecosystems. As with other hydrologic measures, patterns of peak flow vary among Region 2 rivers (Figure 12). 

Landscape changes affecting stream power or material flows inevitably affect 

riparian area characteristics (Naiman and Decamps 1997; Naiman and others 1993).

Riparian ecosystems are generally maintained by an active natural disturbance 

regime operating over multiple spatial and temporal scales. Natural physical disturbance agents include floods, fire, landslides, avalanches, and channel migration. 

These perturbations and the ecological responses to the changed conditions are 

important because many anthropogenic impacts are manifested through changes 

to disturbance regimes (Brinson 1981; Patten 1998; Shafroth and others 2002). 
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Figure 12—Plots demonstrating different inter-annual patterns of peak annual 

discharge for three streams in the region. Although there is considerable 

variability in the magnitude of peak flows, peak annual discharge on the 

Niobrara River (upper panel) is greater than ~51 m3/s for all years of record, 

illustrating the influence of groundwater inputs from the adjacent Sandhills 

in maintaining minimum flows. In contrast, Spearfish Creek (middle panel), 

which lacks similar groundwater inputs and instead loses water as it flows 

through limestone-dominated Spearfish Canyon, exhibits a wider range of 

variability in peak annual flows. The La Plata River exhibits characteristics 

intermediate between the two, exhibiting greater inter-annual variability than 

Spearfish Creek, but fewer extreme events.
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Turner and others (1993) considered four major factors in characterizing the 

landscape dynamics of disturbances: (1) disturbance frequency (or its inverse, 

the interval between successive disturbances), (2) rate of recovery from disturbance, (3) the size or spatial extent of the disturbance events, and (4) the size 

or spatial extent of the landscape (Turner and others 1993). Characteristics for 

each of these factors vary with disturbance type and geographically throughout 

Region 2, making generalizations difficult.

Floods are the most important disturbance type in many riparian ecosystems. 

Because floods have such a broad range in magnitude, it is useful to reserve the 

term “flood disturbance” for those flows that directly or indirectly cause plant 

mortality. The effect of floods on riparian vegetation includes direct plant mortality, as well as indirect effects on vegetation through altered environmental 

conditions and resource availability. Ice flows can also be a factor influencing 

patterns of mortality among riparian species (Auble and Scott 1998).

Through their foraging, beaver can profoundly change the age and size structure 

of woody riparian communities (Baker and others 2005a; Breck and others 2003; 

Johnston and Naiman 1990b). Their dam building activities strongly modify 

local hydrology, thereby influencing a wide range of biotic and abiotic processes. 

Beaver dams can radically change hydrologic regimes, expanding wetland area 

and altering sediment regimes (Butler and Malanson 1995; Meentemeyer and 

Butler 1999; Wohl 2000). Widely recognized as a keystone species, beaver are 

an important factor influencing local and regional biodiversity patterns and geomorphic change (Dickens 2003; Johnston and Naiman 1987; Naiman and others 

1994; Wright and others 2003; Wright 2002).

Specific ecological effects from beaver dams can vary depending on the size 

and ecological context of the river and the age and function of dams. The lifespan of dams and associated ponds can vary widely, from less than one year to 

decades or longer depending on characteristics such as stream mean and peak 

discharge, gradient, and power (Collen and Gibson 2001; Gurnell 1998). Beaver 

dam failure can also modify stream channels and open up large areas for new 

willow establishment, and dams can cause channel evulsions, allowing for plant 

establishment in the abandoned channel (Cooper and others 2006).

The effects of beaver on riparian ecosystems vary depending on the spatial scale 

considered. The landscape pattern of beaver-altered habitat changes along stream 

channels in response to pond age, successional status, and the local environment 

(Naiman 1988; Ray and others 2001; Wright and others 2003). There is also an 

important temporal component to landscape-level patterns of beaver occupation 

influenced by factors such as beaver population abundance and dam persistence.
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Although widely studied in upland forests, relatively few studies have investigated the role of fire in western riparian ecosystems. Research suggests that in 

montane and subalpine settings, fire frequency in riparian areas is often lower 

than in adjacent upland cover types, although differences may be minor in drier 

habitats like ponderosa pine forests and grasslands (Dwire and Kauffman 2003). 

Because fire regimes vary so widely among upland cover types, generalizations 

are difficult to make for riparian areas. However, the broad gradients in fire 

frequency that are documented for uplands—high-elevation ecosystems rarely 

experiencing fire and low-elevation shrublands and grasslands experiencing 

relatively frequent fires—likely apply to riparian systems (Pyne and others 1996).

Effects of fire on individual riparian species can include injury or mortality 

(Brown and others 2000). Some species are more resistant than others to the 

effects of fire because of morphological or physiological characteristics such as 

thick bark. For example, mature Populus deltoides have thick bark with relatively 

low specific gravity and thermal conductivity. As a result, it takes a fairly long 

exposure to high temperatures to transmit enough heat to cambial layers to cause 

mortality (Brown and others 2000). In contrast, non-woody species, thin-barked 

woody plants, and seedlings and saplings of woody species are more vulnerable 

to damage from fires (Whelan 1995).

Fires may promote the establishment of certain species by clearing vegetation and 

litter, providing sites for seedling germination (Whelan 1995). Openings in closed 

forest canopies also allow for shade-intolerant species such as aspen, willows, and 

cottonwoods to thrive. Fire may also favor species capable of resprouting such 

as Populus angustifolia (Gom and Rood 1999). Many other species, including 

several common riparian shrubs such as Alnus spp. (alder), Betula spp., Ribes 

spp., Rosa spp., Salix spp., and Symphoricarpos spp. can sprout from stumps, 

root crowns, and belowground stems following fire (Dwire and Kauffman 2003).

Riparian Classification and Gradients_____________

The spatial connectivity of riparian areas with streams influences the flow of 

energy, propagation of disturbance (e.g., flooding), and a variety of ecological 

processes such as plant and animal dispersal. Along streams, material and energy 

flow longitudinally along channels, laterally between the channel and floodplains, 

and vertically with the soil and groundwater environment (Naiman and Decamps 

1997; Stanford and Ward 1988; Vannote and others 1980). These, together with 

the temporal dimension, have been described as the four dimensions of rivers 

(Ward 1989) and form a basis for understanding different functional processes 

in riparian ecosystems.
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Riparian areas are highly connected upstream and downstream, and diverse 

physical and ecological processes prevent the formulation of a single, all-purpose 

classification. Numerous classification schemes have been developed for rivers. 

Some emphasize hydrologic or geomorphic characteristics (Montgomery 1999; 

Newson and others 1998; Rosgen 1994), while others emphasize biotic elements 

such as vegetation (Carsey and others 2003; Girard and others 1997b; Jones and 

Walford 1995; Walford and others 1997). Others, are more integrative, combining 

biological and physical elements (Brinson and others 1995; Euliss and others 

2004; Zogg and Barnes 1995).

One basic distinction among rivers is the duration of flow. Perennial systems 

have flow throughout the year. Flow in intermittent and ephemeral systems is 

limited temporally to specific periods when groundwater or precipitation inputs 

are available. Divisions between intermittent and ephemeral streams are somewhat 

arbitrary and are generally made in reference to the length of time surface flows 

are present and the frequency of flow events. Another distinction is whether the 

flow of groundwater is toward the river—gaining—or whether the direction of 

groundwater flow is from the river to the riparian zone and uplands—losing. 

Notably, individual rivers can have both gaining and losing reaches, affecting 

plant growth (Harner and Stanford 2003).

Another distinction commonly made among streams is that of stream order. 

Stream order designations (e.g., first, second, and so on) refer to the relative 

position of a stream segment in a drainage basin network. First order streams 

are the smallest, unbranched, perennial tributaries; the junction of two first-order 

streams produces a second-order stream; the junction of two second-order streams 

produces a third-order stream, etc. (Knighton 1998). A variety of changes in 

structure and function occur along gradients from low-order headwater systems 

to larger, high-order systems (Brinson 1993b; Naiman and others 1987). For 

example, sources of floodwater delivery change dominance from groundwater 

discharge and overland runoff in low-order streams to dominance by overbank 

flooding in high-order streams (Brinson 1993b). Geomorphic characteristics 

such as sediment delivery and dominant sediment size also vary systematically 

along streams (Church 2002).

The size of riparian areas differs along broad geographic and elevation gradients. For example, although more numerous, small headwater streams are often 

almost completely embedded in the surrounding forest. The riparian zone is 

larger in mid-sized streams and generally contains distinct patches of vegetation 

determined by long-term (>50 years) channel dynamics and the annual discharge 

regime (Naiman and Decamps 1997). The riparian areas and floodplains adjacent 

to large streams can be physically complex and characterized by long periods 

of seasonal flooding, lateral channel migration, the presence oxbow lakes in old 

river channels, and a diverse mosaic of wetland communities.
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Using the classification scheme developed by Cowardin and others (1979) and 

used by the National Wetlands Inventory (NWI) mapping program, riparian areas 

may be placed in either the palustrine or riverine system. Within each system 

except palustrine, subsystems are identified, with four for riverine wetlands: 

tidal, lower perennial, upper perennial, and intermittent (Tiner 1999). Only the 

last three are relevant to Region 2. Within each subsystem, a variety of classes 

can be designated based on substrate conditions or vegetation (Cowardin and 

others 1979).

The Hydrogeomorphic (HGM) approach, developed for wetland functional assessments, emphasizes hydrologic and geomorphic processes as they influence 

wetland and riparian ecosystem structure and function (Brinson 1993a). Riparian 

ecosystems are in the Riverine HGM class (Brinson and others 1995; Hauer and 

others 2002a). Riverine wetlands can be further divided into subclasses that more 

precisely reflect stream flow and fluvial processes (Tiner 1999) (Figure 13).

Figure 13—DCA ordination of Region 2 riparian vegetation data illustrating main HGM subclasses as 

identified by Cooper (1998). See Chapter 2 for description of methods used in ordination and data set 

description.
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In a classification of Colorado wetlands using HGM as a base, Cooper (1998) 

described a total of five Riverine subclasses. Wetlands in the Riverine subclass 1 

(R1) typically occur along steep-gradient, low-order streams and springs on 

coarse-textured substrates (Cooper 1998a). Thirteen R1 plant associations were 

defined by Carsey and others (2003) for this subclass mostly occurring in the 

subalpine zone. Riverine subclass 2 (R2) communities occur on coarse and finetextured substrates along moderate-gradient streams at intermediate elevations 

and often support beaver pond complexes. In their analysis, Carsey and others 

(2003) identified 46 plant associations including coniferous and deciduous 

forests, shrublands, and herbaceous types. Riverine Subclass 3 includes shrub 

and tree-dominated riparian areas along middle-elevation reaches of loworder and mid-order streams. Subclass R4 occupies lower-elevation canyons 

in the foothills and plateaus along larger rivers or small intermittent streams 

(Carsey and others 2003). Seventy-six plant associations were identified in 

these two combined subclasses. Subclass 5 (R5) wetlands are found on lowelevation floodplains of mid- to high-order streams. Substrates are typically 

fine-textured and streamflow is usually perennial but occasionally intermittent. Associations in this subclass are most common below 7000 ft (2130 m) 

but may occur up to 9,800 ft (2,990 m). Thirty-four plant associations were 

identified in the R5 subclass, including shrublands, grasslands, or deciduous 

woodlands (Carsey and others 2003).

Riparian vegetation in the region

Zonation within riparian areas—Because of the linear nature of riparian environments, gradients in key environmental variables and resulting vegetation 

patterns vary both laterally and longitudinally, although the scale differs depending 

on the direction examined. Variation across mountain stream valleys is typically 

seen on a scale of meters, while longitudinal variation may be more appropriately 

measured in kilometers (Bendix 1994). Key gradients driving vegetation patterns 

laterally across streams are water table depth, elevation, soil characteristics, and 

flood disturbance. These often create distinct vegetation zones across riparian 

areas. Along meandering rivers with broad floodplains, episodic flood events 

and channel migration cause pulses of establishment of woody species such as 

willow and cottonwood, creating patches of different age.

Water source affects both the width of the riparian zone and its vegetation composition. For example, water tables along intermittent and ephemeral streams 

often drop well below stream channels, limiting plant species to deep rooted 

phreatophytes or upland species that are tolerant of low soil moisture (Goodwin and others 1997). In contrast, many headwater streams receive perennial 

groundwater inputs from side slopes, helping to maintain high water tables long 

after seasonal peak flows have diminished. Whether a stream is gaining or losing 

also affects the sensitivity of riparian communities to changes in surface water 

regimes associated with river regulation.
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Longitudinal changes in riparian communities—Longitudinal changes in 

riparian vegetation can differ among streams depending on stream length and 

elevation gradient. A typical progression along a Region 2 river might proceed 

as follows. In high-elevation headwater areas, vegetation is often dominated by 

herbaceous plants, such as Senecio triangularis, Mertensia ciliata, and Cardamine 

cordifolia. Low stature shrubs, such as Salix planifolia, may also be common, 

particularly where shallow groundwater and snowmelt are the principal water 

sources. Moving downstream, the stature, composition, and diversity of shrub 

dominants typically increases. Salix spp. are dominant in broad, open basins 

and low-gradient streams, and along higher-gradient streams, Alnus incana 

ssp. tenuifolia and Betula fontinalis are abundant. Along larger, mid-elevation 

montane streams, and lower in the foothill valleys, riparian forests dominated 

by Populus angustifolia and Picea pungens are common along with the various 

shrub and herbaceous species. Streams in the lowlands are often dominated by 

large cottonwoods such as Populus deltoides in the Great Plains and P. fremontii 

west of the Continental Divide. Certain species tend to be diagnostic of particular 

riparian types (Figure 13, Table 11).

Region 2 supports a wide range of specific riparian vegetation types. Using GAP 

data, the Nature Conservancy developed a vegetation classification for all of 

Region 2 based on the primary classification unit ecological systems, defined 

as assemblages of biological communities found in sites with similar physical 

environments, disturbance regimes, underlying environmental features, or environmental gradients (Comer and others 2003b). In their analysis, Comer and 

others (2003b) describe three riparian ecological systems, comprising approximately 2.48 percent of the Region 2 area (Table 12).

Western Great Plains riparian/western Great Plains floodplain—The 

western Great Plains riparian/floodplain system includes riparian areas associated 

with medium and small rivers and streams throughout the western Great Plains. 

It extends through the central shortgrass prairie and northern Great Plains 

steppe into Wyoming, and it occurs on alluvial soils in highly variable landscape 

settings, from deep cut ravines to wide, braided streambeds (Comer and others 

2003b). Relative to larger rivers, these systems have relatively flashy hydrologic 

regimes. Common species include Populus deltoides, Salix exigua, Artemisia

cana ssp. cana, Pascopyrumsmithii, Sporobolus cryptandrus, and Schizachyrium

scoparium. Non-native species including Tamarix spp. and Elaeagnus angustifolia

may dominate in heavily degraded areas, (Comer and others 2003b).

Rocky Mountain subalpine-montane riparian shrubland—The Rocky 

Mountain subalpine-montane riparian shrubland system occurs throughout 

the Rocky Mountain cordillera from New Mexico to Montana and includes 

montane to subalpine riparian shrublands in narrow to wide, low-gradient valley 

bottoms and floodplains with sinuous stream channels (Comer and others 2003b). 
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Table 11—HGM indicator species identified by Cooper (1998).

Scientific name Common name HGM subclass

Mertensia ciliata tall fringed bluebells R1

Mimulus guttatus seep monkeyflower R1

Oxypolis fendleri Fendler’s cowbane R1

Calamagrostis canadensis bluejoint reedgrass R1

Cardamine cordifolia heartleaf bittercress R1

Glyceria striata fowl mannagrass R1

Salix boothii Booth willow R2

Salix geyeriana Geyer willow R2

Salix monticola Rocky Mountain willow R2

Geum macrophyllum var. perincisum largeleaf avens R2

Heracleum maximum common cowparsnip R2

Cornus sericea ssp. sericea red-osier dogwood R3

Picea pungens blue spruce R3

Poa pratensis Kentucky bluegrass R3

Populus angustifolia narrowleaf cottonwood R3

Rosa woodsii Woods’ rose R3

Rudbeckia laciniata var. ampla cutleaf coneflower R3

Salix ligulifolia strapleaf willow R3

Alnus incana ssp. tenuifolia thinleaf alder R3

Equisetum arvense field horsetail R3

Acer negundo var. interius boxelder R4

Salix irrorata dewystem willow R4

Agrostis gigantea redtop R4

Bromus inermis smooth brome R4

Carex pellita woolly sedge R4

Corylus cornuta beaked hazelnut R4

Hordeum jubatum ssp. jubatum foxtail barley R4

Juncus torreyi Torrey’s rush R4

Phragmites australis common reed R5

Populus deltoides eastern cottonwood R5

Rhus trilobata var. trilobata skunkbush sumac R5

Salix amygdaloides peachleaf willow R5

Salix exigua narrowleaf willow R5

Salix fragilis crack willow R5

Spartina gracilis alkali cordgrass R5

Spartina pectinata prairie cordgrass R5

Tamarix ramosissima saltcedar R5

Vitis riparia riverbank grape R5

Carex emoryi Emory’s sedge R5

Chrysothamnus linifolius spearleaf rabbitbrush R5

Elaeagnus angustifolia Russian olive R5

Lycopus americanus American water horehound R5
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Table 12—Approximate area of different Region 2 ecological systems as determined by 

Comer et al. (2003). Riparian systems are indicated in bold.

Ecological Systems Hectares Percent area

Inter-Mountain Basins Playa 12,749 0.0

North American Arid West Emergent Marsh 128,301 0.1

North Central Interior Floodplain/Wooded Draw 771,987 0.7

Rocky Mountain Subalpine Mesic Meadow 884,960 0.8

Western Great Plains Closed Depression 817,203 0.7

Western Great Plains Riparian/ 1,488,930 1.3

Western Great Plains Floodplain

Western Great Plains Saline Depression 2,828 0.0

Rocky Mountain Subalpine Mesic-Spruce-Fir 48,164 0.0

Forest and Woodland

Rocky Mountain Lower Montane Foothill 848,753 0.7

Riparian Woodland and Shrubland

Rocky Mountain Subalpine-Montane 475,655 0.4

Riparian Shrubland

Region 2 Total 113,543,893

These communities are typically found at higher elevations, from 5577 to 

11,400 ft (1700 to 3475 m). Communities often form a mosaic of shrub- and 

herb-dominated vegetation types with species such as Alnus incana, Betula nana, 

B. occidentalis, Cornus sericea, and a number of different willows, including 

Salix spp. (Comer and others 2003b).

Rocky Mountain lower montane foothills riparian woodland—The Rocky 

Mountain lower montane foothills riparian woodland includes seasonally flooded 

conifer and aspen woodlands found at montane to subalpine elevations of the 

Rocky Mountain cordillera. These are found on floodplains or terraces between 

4900 and 10,825 ft (1500 and 3300 m) above sea level (Comer and others 2003b). 

Dominant species include Abies lasiocarpa, Picea engelmannii, Pseudotsuga

menziesii, Picea pungens, Populustremuloides, and Juniperusscopulorum. Also 

commonly present are Alnusincana, Pinus contorta, Populus angustifolia, Acer

negundo, and Juniperus osteosperma.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 64

HRV of Riparian Ecosystems in Region 2 __________

Overview of anthropogenic impacts

Streams and rivers are among the most systematically altered ecosystems in North 

America. Anthropogenic impacts are diverse and include both direct alterations 

occurring within ecosystem boundaries and indirect alterations. The latter include 

alterations to contributing watersheds that affect the quantity, quality, or timing 

of sediment, water, or other critical inputs such as plant propagules. Because of 

the large geographic scope of this assessment and the high ecological diversity 

of riparian types in Region 2, it is not possible to provide a single estimate of the 

HRV and possible deviations since the arrival of Euro-Americans that would be 

applicable to all wetland and riparian ecosystems. Our approach is to provide a 

review of the main anthropogenic impacts that have affected riparian areas in 

the region with the understanding that the relative importance of any particular 

impact on the HRV of a given area will differ (Table 13).

Flow regulation and diversion—Since the mid-Nineteenth Century, water in 

Region 2 has been used to support municipalities, agriculture, power production, mining, and recreation. The infrastructure and consumptive uses of water 

increased dramatically during the mid-Twentieth Century with a large increase in 

the number of dams constructed on rivers in the region (Graf 1999; Wohl 2000) 

(Figure 14). Reservoir storage on the North Platte River, for example, experienced 

more than an eight-fold increase between 1900 and the present.

Dams have been constructed for a variety of purposes. In the NID dataset, the 

use ascribed to 34.5 percent of dams in Region 2 is fire/farm pond. These are 

generally small dams and are most concentrated in the plains. The next most 

common use listed is flood control (20.4 percent) followed by water supply 

(16 percent) and irrigation (15.6 percent). Hydroelectric is listed as the primary 

purpose for only 0.7 percent of dams; however, this portion includes some of 

the largest dams in the region.

Dams are distributed throughout the region, from small earthen dams on headwater 

streams to large dams on big rivers managed for electricity production and water 

storage. In terms of overall numbers, dams are most densely concentrated in areas 

of agricultural production such as eastern and central South Dakota, Kansas, and 

Nebraska. Dams are also concentrated along population centers, for example, 

the Colorado Front Range. Because the majority of rivers in the region originate 

as snowmelt in the mountains, many large dams occur in mountainous areas.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 65

Table 13—Summary of main impacts to rivers in the Front Range of Colorado and their approximate primary period of 

impact (Wohl 2005).

Land use Period Influence on rivers

Beaver trapping 1815–1840 Removal of beaver dams increased flow velocity, stream bed and 

bank erosion. Decreased sediment storage, channel stability, and 

diversity. 

Placer mining 1859–1940s Direct effects: Disruption of stream bed and bank structure, 

increased sediment and channel mobility, altered flow regime 

(if diversion occurs), and introduced toxic contaminants such 

as mercury. Indirect effects: increased regional population, 

the amount of timber harvested, the number of transportation 

corridors, and the amount of sediment and contaminants entering 

channels. 

Railroad tie drives 1860s–1890s Effects included modification of the channels prior to tie drives 

(e.g., the removal of obstructions and naturally occurring wood 

and the blocking-off of overbank areas); scouring effects from 

pulses of water and wood. 

Diversions and dams 1859 to present Altered magnitude, duration, and frequency of flows, as well as 

sediment transport, disturbance regimes, water chemistry, and 

water temperatures. 

Timber harvest 1859–1940s Effects included destabilizing hill slopes and increasing water and 

sediment yields to rivers. 

Transportation corridors, 1860 to present Effects include increasing sediment to rivers from destabilized hill

such as roads and railroads slopes, erosion of unpaved roads, and application of traction sand 

and gravel during the winter. Transportation corridors also reduce 

the width of the floodplain and riparian corridors.

Lode mining 1859–1980s Effects include increased sediment yield from hill slopes and 

introduction of toxic contaminants to rivers. 

Urbanization 1859 to present Effects include an initial increase of water and sediment yield 

to rivers, followed by an increase primarily in water yield that 

introduces contaminants into rivers and constrains channel and 

floodplain space and mobility. 

River recreation 1909 to present Fishing creates pressure on native species and promotes the 

introduction of other species. Whitewater rafting locally creates 

trampled streambanks with compaction, decreased infiltration, 

increased runoff and erosion, and damage to riparian vegetation. 

Grazing 1860 to present Effects include removing riparian vegetation and compacting 

streambanks, bank erosion, wider and shallower channel crosssections, finer stream bed substrates, increased nutrient input 

to rivers, warmer water temperatures, and reduced aquatic and 

riparian habitat.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 66

Figure 14—Dam construction in the states that are part of Region 2; data from National Inventory of Dams (U.S. Army Corps of Engineers 1996).

In addition to large dams, numerous trans-basin diversions occur throughout 

the region, affecting both receiving and donor rivers (Henszey 1993; Johnson 

and others 1999; Pringle 2000; Woods and Cooper 2005). These convey water 

over drainage divides using tunnels or ditches intercepting tributaries (Chimner 

and Cooper 2003; Woods and Cooper 2005). General effects can include altered 

hydrologic regimes, disturbance processes, chemical cycling, and community 

composition (Adair and others 2004; Nilsson and Berggren 2000; Wohl 2000). 

Trans-basin diversions are most numerous in Colorado, with nearly 40 structures found in the South Platte basin (Wohl 2001). Hydrologic and ecological 

effects vary among streams. For example, high-gradient, step-pool streams are 

less likely to see major changes in response to increased discharge than lowergradient streams (Wohl 2000).

Hydrologic effects of dams and diversions vary but typically include alteration 

of the magnitude of monthly water conditions, reduction in the magnitude and 

duration of annual extreme discharge events, the changes in the timing of annual 

extreme discharge events, and the frequency and duration of high/low flow pulses 

and hydrograph changes (Richter and others 1997). Changes in magnitude can 

cause increased scour of vegetation, promote the invasion of exotic species, and 

reduce water and nutrients to floodplain vegetation (Nilsson and Svedmark 2002; 

Poff and Hart 2002). Water releases from reservoirs may cause channel incision, 

isolating adjacent floodplains or tributary outlets (Kondolf 1997; Petts 1984).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 67

On large rivers throughout the West, the decline of native cottonwood riparian 

forests has been a large concern (Cooper and others 2003b; Patten 1998). Because cottonwoods depend on shallow floodplain groundwater, they are particularly vulnerable to modifications of river flow. Hydrologic modifications from 

river regulation can reduce cottonwood growth rates and cause canopy dieback 

and tree death, as well as inhibit new seedling establishment and recruitment 

(Mahoney and Rood 1998; Williams and Cooper 2005). Changes in the timing 

of peak flows can also affect Populus spp. and Salix spp., which disperse seed 

in the early summer and produce non-persistent seeds with narrow windows of 

viability (Karrenberg and others 2002). Management of dams for hydropower 

typically involves dramatic daily flow variation (Poff and Hart 2002; Poff and 

others 1997).

Large changes in disturbance frequency or level of landscape connectivity 

can disrupt plant dispersal, resulting in reduced plant diversity and compromised ecological integrity (Andersson and others 2000; Merritt and Wohl 

2002; Merritt and Wohl 2006). For example, Merritt and Wohl (2006) found 

significant differences in community composition upstream and downstream 

from reservoirs in Colorado, suggesting that 50 to 100 years of fragmentation 

has caused community-wide effects along Rocky Mountain streams, which the 

authors partially attributed to dispersal limitation due to dams and reservoirs. 

Landscape-level impacts are particularly important if one considers the extent of 

river regulation. For example, less than 50 high-quality, undammed rivers longer 

than 200 km remain in the continental United States (Benke 1990).

Beaver trapping—The first and one of the largest impacts to riparian areas by 

Euro-Americans in Region 2 was the trapping and near removal of beaver from 

regional streams. Beaver were widespread throughout most of North America, 

numbering between 60 and 400 million before European settlement. Although 

natural predators such as wolves, coyotes, and bear opportunistically fed on 

beaver, until Europeans arrived in North America, beaver populations were 

controlled principally by disease and food availability (Sandoz 1978). Beaver 

played a keystone role in riparian ecosystem functioning, profoundly altering 

the vegetation and hydrologic regimes of their environment through their feeding and dam building activates. Beaver dams and their associated ponds elevate 

water tables, trap sediment, and alter biogeochemistry (Naiman and others 1988).

Most beaver populations were decimated by fur trappers during the 1700s and 

1800s (Sandoz 1978; Wishart 1974). By 1840, beaver populations were depleted 

in many areas of the Rocky Mountains (Travis and others 2002). Although beaver reintroduction and a decline in demand for pelts has helped populations to 

recover throughout much of their former range, beaver have failed to recover in 

some riparian environments that have become heavily browsed by livestock or 

ungulates since European settlement (Baker and others 2005a).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 68

In some areas, the removal of beaver increased stream channel down-cutting, 

reducing the complexity of wetland habitats (Wohl 2001). In Rocky Mountain 

National Park, beaver absence caused in part by habitat degradation from excessive elk populations has resulted in the conversion of willow riparian shrub 

communities to dry meadows (Baker and others 2005a; Peinetti and others 

2002a). Considering their importance as geomorphic and ecological drivers, the 

decimation of beaver populations represents one of the most significant deviations from the HRV for riparian ecosystems.

Agricultural impacts—Agricultural impacts include all activities directly 

related to the production of food, feed, or fiber, excluding the raising of domesticated animals, which is discussed subsequently. Agricultural impacts can 

include direct effects on riparian vegetation and functions such as vegetation 

clearing or channel realignment, soil compaction by heavy machinery. Indirect 

effects can include fragmentation and reduction of wildlife habitat, reduced 

plant diversity and invasion of exotic species, and increased water temperature 

(USDA NRCS 1998).

The extent, nature, and magnitude of impacts to riparian ecosystems from historical and contemporary agricultural activities in Region 2 vary widely. Excluding 

ranching, most mountainous areas have never supported agriculture. In contrast, 

agriculture on the Great Plains and Intermountain valleys is common. EuroAmerican settlement and agriculture in the region began in the mid-Nineteenth 

Century, largely to support population growth in mining areas, and it increased 

with the completion of the railroads. Proximity to water was a priority, so riparian 

areas tended to be among the first that were cleared and settled. The dust bowl 

drought of the 1930s, in particular, saw the failure of many farms, especially 

in areas west of the 100th meridian. Many of these lands ultimately entered the 

public domain and are managed as National Grasslands. Although, the Great 

Plains experienced substantial declines in population between 1950 and 2000, 

the area in crops declined by less than 1 percent overall (Brown and others 2005).

Livestock grazing—Livestock grazing can have significant impacts on 

riparian ecosystems, as illustrated by a variety of studies (Belsky and others 

1999; Chaney and others 1990; Fleischner 1994). Effects include removing riparian vegetation and compacting streambanks, bank erosion, wider and shallower 

channel cross-sections, finer stream bed substrates, increased nutrient input to 

rivers, warmer water temperatures, and impaired aquatic and riparian habitat. 

Fleischner (1994) identified three principal impacts of livestock grazing: (1) 

changes of species composition, including decreases in density and biomass of 

individual species, reduction of species richness, and changing community organization; (2) disruption of ecosystem function, including altered nutrient cycling 

and successional patterns; and (3) alteration of ecosystem structure, including 

changing vegetation stratification, soil erosion, and decreasing availability of water 

to biotic communities. Livestock can also impact riparian areas by transporting 

seeds and vegetative propagules of invasive species (Belsky and others 1999).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 69

Historically, many areas in the region were overgrazed, although livestock 

numbers are now significantly lower than at their historical peak (Elmore and 

Kauffman 1994). The legacy of overgrazing varies over the region; some areas 

have recovered while others still bear scars, particularly in areas where significant geomorphic change such as channel downcutting has occurred. The level 

of impact can vary widely depending on the sensitivity of the system, and the 

intensity and seasonality of grazing (Chaney and others 1990).

In general, impacts to riparian areas depend on the species of livestock and the 

characteristics of the habitat present (Gillen and others 1985; Marlow and Pogacnik 1986). Cattle, in particular, are typically regarded as having the greatest 

potential impact on riparian systems. They tend to prefer low-gradient riparian 

systems because of available water and abundant palatable forage (Belsky and 

others 1999). Seasonality of use also affects cattle use and impacts. For example, 

late summer pastures tend to have more concentrated cattle use in riparian areas 

(Parsons and others 2003). In contrast, sheep typically prefer uplands, utilizing 

riparian areas more to acquire water than for forage (Glimp and Swanson 1994); 

however, in sufficient numbers, they too can cause significant physical site impacts.

Native ungulates can also have significant effects on wetlands and riparian areas. 

Impacts to the physical environment include soil compaction and bank instability, direct effects to vegetation like herbivory and trampling, and indirect effects 

like nutrient enrichment via urine or fecal deposits (Hobbs 1996; Opperman 

and Merenlender 2000). Moose use riparian systems almost exclusively in the 

summer and can have significant effects on Salix spp. (Oedekoven and Lindzey 

1987; Singer and others 1994). Historically, elk would have been less common in open riparian areas due to predation avoidance behavior; however, the 

extirpation of predators such as wolves has resulted in large increases in elk 

populations in several areas of the Rocky Mountains (Coughenour and Singer 

1996; Peinetti and others 2002b). In addition to their greater numbers, elk spend 

a large amount of time in riparian areas, causing significant impacts to riparian 

vegetation (Beschta 2003; Fortin and others 2005; Ripple and Beschta 2004; 

Wolff and Van Horn 2003).

Bison were widespread throughout the Great Plains and some intermountain basins prior to the arrival of Euro-American in the region (Knapp and others 1999). 

Their effects on riparian areas are generally poorly studied but are likely similar 

to those of cattle, including soil compaction and bank instability, herbivory, and 

nutrient enrichment via urine or fecal deposits (Hobbs 1996). However, bison 

feeding behavior differs from that of cattle, both in terms of preferred species 

and habitats. Like cattle, bison feed primarily on grasses and sedges, and avoid 

herbaceous dicotyledons and woody species, which usually constitute less than 

10 percent of their diet (Knapp and others 1999), although they generally spend 

less time in riparian areas. Their impacts were likely more localized than those 

of cattle and centered around major stream crossings (Fritz and Dodds 1999). 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 70

The elimination of bison has been cited, along with other environmental changes 

such as the cessation of prairie fires and river regulation, as a possible factor influencing the expansion of woodlands along the Platte River system in Nebraska 

over the last century (Johnson 1994).

Timber harvest—The effects of timber harvest on riparian areas has a variety 

of direct and indirect effects on biophysical processes and habitat characteristics. 

Impacts can include changes in the timing or magnitude of runoff events, reduced 

stability of stream banks, increased sediment supply to channels, and changes 

in in-channel sediment storage and structure. Timber harvest can also directly 

impact riparian areas by causing plant mortality. Logging can affect stream flow 

by altering the watershed processes controlling water movement from hillsides 

to stream channels. Other activities associated with timber harvest, including 

skid trails, landings, and roads, can also impact riparian and wetland resources.

Changes in basin vegetation cover due to timber harvest can alter surface runoff from basins through effects on evapotranspiration rates and snowpack accumulation patterns. For example, canopy removal in a subalpine watershed in 

Colorado increased precipitation reaching the forest floor by approximately 40 

percent and, increased peak snowpack water equivalent by more than 35 percent 

(Stottlemyer and Troendle 1999; Stottlemyer and Troendle 2001). Clearcutting 

and partial thinning both result in increased total annual and peak streamflow in 

logged watersheds (Troendle and King 1987).

Of particular historical importance to many riparian systems were tie-drives. 

Beginning in the 1860s, demand for timber increased dramatically, driven in part 

by the expansion of the railroads (Wohl 2005). Rivers were used to transport 

huge numbers of railroad ties downstream from harvest areas to collection points. 

Wohl (2005) reported that more than 200,000 ties went down Colorado’s Poudre 

River in Colorado annually from 1868 to 1870. To facilitate conveyance of the 

ties, stream channels were often strongly modified, for example, by artificially 

straightening meanders with cutoffs. Additional impacts resulted from the scouring effect of moving large log masses downstream. Impacts to some streams are 

still discernible over a century after the drives occurred and can include reduced 

levels of mature riparian vegetation and wider, shallower channels with less pool 

volume (Neal and Zubick 2003; Wohl 2005; Young and others 1994).

Mining impacts—Mineral extraction was one of the primary historical drivers 

of Euro-American settlement in many portions of Region 2 (West 1998). Mining is still widespread in the region, although the number of active operations 

is a fraction of historical levels (Table 14). Impacts from mining’s legacy are 

among the most pressing environmental concerns in the region today (Limerick 

and others 2005; Wohl 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 71

Placer mining was a significant historical disturbance to many aquatic and riparian 

habitats in the region, accomplished with hand operation of gold pans, shovels, 

and sluices and later by larger scale hydraulic mining and dredging (Wohl 2005). 

Dredging entails the excavation of the streambed and floodplain, which destroys 

vegetation and undermines channel stability. The spoils are typically spread or 

piled on the valley margin and are often difficult to revegetate. In hydraulic 

mining, jets of water are directed onto ore-bearing alluvial deposits. Because 

of environmental concerns, hydraulic mining is generally prohibited through 

legislation, but early gold miners widely employed the technique.

Gravel and sand mining has increased in some areas of Region 2 in response to 

demand from development. Extraction of aggregate materials from rivers and 

streams alters channel morphology, streambank stability, channel patterns, and 

riparian vegetation (Kondolf 1994; Kondolf and Swanson 1993). Effects can 

include increased sedimentation, introduction of contaminants, widening of 

channels, direct loss of vegetation, and modification of streams and wetlands due 

to tailings (Meador and Layher 1998; Scott and others 1999). Aggregate operations are generally confined to sites along larger alluvial rivers where significant 

deposits are available.

Hard rock mining influences many riparian areas through disruption of the stream 

bed and bank structure, increased sediment and channel mobility, altered flow 

regimes, and the introduction of contaminants such as lead and mercury (NRC 

1999; Wohl 2001). Hard rock mining can significantly impair water quality and 

effect aquatic, riparian, and wetland biota. Specific impacts differ depending on 

the type of ore, and the extraction and processing methods used. Because mines 

typically extend below the regional water table, groundwater may flow into the 

mine pit, ultimately exiting contaminated with metals through drainage adits 

(Limerick and others 2005). Overburden, waste rock dumps, tailing impoundments, leach pads, process ponds, and mills can serve as sources of contaminants 

like heavy metals that negatively affect the quality of surface and groundwater.

Table 14—Mining claim totals by type and status in Colorado (Hyndman and Campbell 1999).

Type of Claim Lode Placer Mill Tunnel All claims

Number of open mining claims 7,491 2,058 222 17 9,788

Number of closed mining claims 216,206 18,638 3,502 121 238,467

Grand Totals 223,697 20,696 3,724 138 248,255
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 72

Energy development—Energy development, including coal, oil, and gas 

development, is an additional anthropogenic factor that locally impacts riparian areas in the region. Direct effects from extraction activities are generally 

minor, although indirect effects from the associated infrastructure such as 

roads may be more common. A notable exception is coal-bed methane (CBM), 

which typically produces large volumes of water. CBM discharge waters often 

cause an increase of salinity and sodicity within exposed soils, altering soil 

physical and chemical properties and affecting the density and diversity of 

vegetation (Stearns and others 2005). CBM operations are concentrated in 

certain portions of the region such as the Powder River Basin (USDI 2000), 

and, thus, are not a factor for much of Region 2 riparian ecosystems.

Transportation—Transportation infrastructure, including roads, trails, and 

railroads, have affected riparian areas by altering local and watershed-scale 

hydrologic processes. Structures such as bridges and railroad prisms may confine river channels, in turn causing increased flow velocity, which scours and 

coarsens bed sediments (Wohl 2005). Engineering structures such as culverts and 

ditches can alter natural drainage patterns, reduce interception and infiltration 

rates due to the removal of vegetation and soil compaction, and alter the hydrologic response of basins to both annual snowmelt runoff episodes and isolated 

convective storm events (Forman and Sperling 2002; Jones 2000). In addition, 

increased overland flow typically results in a more rapid and extreme hydrologic 

response to precipitation events.

Additional effects can include the introduction of pollutants and the alteration of 

water chemistry from road dust, vehicle exhaust, and chemicals such as deicing 

agents used in road maintenance (Trombulak and Frissell 2000; Wilcox 1986a; 

Wilcox 1986b). Road density; slope and surface type; and the number, size, and 

design of engineering structures are all factors that influence the specific impact of roads. Transportation infrastructure can also affect riparian ecosystems 

by promoting weed invasions. Construction or maintenance activities create 

disturbances conducive to the establishment of many exotics and provide an effective conduit for the dispersal of weed propagules (Gelbard and Belnap 2003; 

Parendes and Jones 2000).

Urbanization—Impairment of riparian communities commonly accompanies 

urban and exurban development (Hansen and others 2005). For example, Auble 

and others (1997) observed declining Populus deltoides establishment on the 

floodplain and terrace of Boulder Creek in Boulder, Colorado from 1937 to 1992, 

attributing the decline in part to the effects of stream diversions and channel 

stabilization. These changes led to decreased channel movement, reductions in 

peak flow volume and flood frequency, reducing opportunities for cottonwood 

establishment and favoring exotic species like Salix fragilis and Elaeagnus angustifolia (Auble and others 1997).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 73

Atmospheric deposition of airborne pollutants such as nitrogen may impact 

riparian ecosystems in the region (Fenn and others 2003a). Typically, hotspots 

of elevated nitrogen deposition occur downwind of large metropolitan centers, 

although elevated levels can also be observed in the proximity of large agricultural 

operations (Fenn and others 2003b), and most areas of Region 2 are exposed to 

some level of atmospheric nitrogen deposition.

Invasive species—The spread of non-native species is changing the biotic 

composition of many Region 2 riparian ecosystems. Woody species, including 

Tamarix spp. and Elaeagnus angustifolia, are particularly widespread along large 

rivers at low elevations in the region, as are herbaceous species such as Phalaris 

arundinacea (reed-canary grass) (Katz and Shafroth 2003; Pearce and Smith 

2003). Stands of exotic plants can be dense, effectively shading out seedlings of 

native competitors like cottonwoods and preventing their establishment (Cooper 

and others 1999). Tamarix spp. can alter geomorphic processes, cause channel 

narrowing by promoting island formation and growth, affect water tables, and 

alter fire regimes (Di Tomaso 1998). As a nitrogen fixer, Elaeagnus angustifolia 

can influence biogeochemical process, possibly facilitating invasion by other 

invasive species (Simons and Seastedt 1999).

In sampling the woody riparian species present at 475 randomly selected stream 

gauging stations in 17 western states, Friedman and others (2005) found that 

Tamarix ramosissima and Elaeagnus angustifolia are widespread across the 

western United States. They found that the two species are now the third and 

fourth most frequently occurring woody riparian plants in the region; and based 

on mean normalized cover, are the second and fifth most dominant woody riparian species. The frequency of occurrence of T. ramosissima may be constrained 

by sensitivity to frost, while the frequency of E. angustifolia decreases with 

increasing minimum temperatures (Friedman and others 2005).

Cumulative and threshold effects—To understand possible deviations from the 

HRV of riparian ecosystems, evaluation of individual stressors alone is insufficient. Increasingly, cumulative impacts (defined as the incremental effect of an 

impact added to other past, present, and reasonably foreseeable future impacts) 

have become a concern for wetlands and watershed managers (Gosselink and 

Lee 1989; Johnston 1994; Preston and Bedford 1988). Cumulative effects can 

be direct or indirect, i.e., they can occur at a location distant from the impacts. 

For example, hydrologic modifications from dams can cause impacts to riparian 

systems hundreds of miles downstream from the location of a particular dam.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 74

Effects can accumulate in a variety of ways. Perturbations that are tightly 

crowded spatially or temporally can overwhelm a system’s ability to recover 

from disturbances (Johnston 1994). In addition, disturbances can interact synergistically to produce effects qualitatively and quantitatively different from 

the individual disturbances. For example, dams, diversions, flow regulation, 

stream channelization, wetlands drainage, and groundwater extraction can 

interact, creating conditions that facilitate the proliferation of exotic species 

in riparian habitats (Busch and Smith 1995; Cooper and others 1999; Pringle 

2000; Shafroth and others 2002).

Changes in the landscape configuration of wetlands and riparian areas may also 

affect ecosystems at the site and reach level. For example, native cottonwood 

forests along the upper Colorado River systems are increasingly fragmented 

due, in part, to the increasing dominance of Tamarix spp. and the conversion of 

floodplains to land uses like agriculture (Northcott and others 2006). The resulting isolation of remaining stands may constrain future cottonwood establishment 

in otherwise suitable areas, as seed dispersal rain levels may be inadequate for 

regeneration.

Cumulative effects may occur gradually or appear once an ecological threshold is 

reached (i.e., a point at which there is an abrupt change in an ecosystem quality, 

property, or phenomenon, or where small changes in an environmental driver 

produce large responses in the ecosystem) (Groffman and others 2006). For example, many riparian areas in Rocky Mountain National Park have historically 

supported high beaver densities (Packard 1947). When beavers cut willows for 

dam construction, they closely crop plants to the ground surface. Following abandonment, heavily used sites are generally unsuitable for new beaver occupation 

until resprouting willows have had an opportunity to grow to sufficient size again 

to provide enough browse and building materials. However, intense elk browsing can maintain willows in a short stature, reducing the suitability of habitat 

for beaver (Baker and others 2005a; Mitchell and others 2000). Over time, this 

has contributed to a decline in beaver populations, altering the basic geomorphic 

and ecological processes of sites and constraining new willow establishment 

(Cooper and others 2006; Gage and Cooper 2005; Peinetti and others 2002b).

HRV assessment for high-elevation springs and creeks

Composition and stand structure—Riparian ecosystems associated with small, 

low-order headwater streams are typically dominated by low-stature, shrubdominated or herbaceous communities. Vegetation associations vary across the 

region but are often dominated by herbaceous dicotyledonous species, including 

Caltha leptosepala, Mimulus guttatus, Saxifraga odontoloma, Mertensia ciliata, 

Cardaminecordifolia and Seneciotriangularis, or monocots such as Calamagrostis

canadensis, Glyceria striata, and Carex utriculata (Carsey and others 2003; 

Walford and others 1997). Woody species may be present in some areas and 

include the willows Salix brachycarpa, S. wolfii, and S. planifolia, and Betula 

glandulosa. When present, woody species rarely exceed 3.3 ft (1 m) in height.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 75

High-elevation springs and creeks are little altered compared with riparian areas 

at lower elevations. As a consequence, the species composition and structure 

of these communities appears to be within the HRV. Other than the presence of 

exotic plants such as dandelion (Taraxacum officinale), high-elevation springs 

and brooks generally lack invasive plants. Because there is a simple vertical 

structure of vegetation in pristine communities, no significant changes from the 

HRV are likely.

Hydrologic regimes—As a class, brooks and streams at higher elevations may 

represent the most hydrologically intact and unmodified riparian ecosystem 

types in the region. Springs and brooks are fed largely by snowmelt or groundwater discharge from shallow aquifers—water sources that in many cases are 

too dispersed, small, and high-elevation to be efficiently exploited by humans 

in most cases. However, there are notable exceptions, such as Spring Creek on 

the Grand Mesa National Forest in Colorado, where surface water is diverted for 

irrigating high-elevation pastures. Other examples of high-elevation diversions 

of very small streams also occur, often for water supply storage in reservoirs. 

However, these impacts are localized and most hydrologic variables of interest, 

such as flood magnitude, duration, frequency, and timing, are likely within the 

HRV for the vast majority of brooks and small streams.

Woody plant establishment—There is scant information available about the 

population structure and establishment dynamics of riparian communities along 

brooks and springs. Many of the woody species that occur in these communities 

are capable of vegetative reproduction (Cottrell 1993, 1995). Because of the 

relatively low incidence of physical disturbance compared to lower elevations 

and lower energy fluvial settings, sexual reproduction is probably episodic. Salix

spp., while low in stature, can be of significant age, reproducing vegetatively for 

centuries (Cooper and others 2006). Although there are few data from which to 

make an assessment, it appears that establishment dynamics are likely within the 

HRV, except where high native or domestic ungulate use has severely impacted 

the structure of willow stands.

Fire regimes—Fire regimes in many portions of the region are outside of the 

HRV, particularly at lower elevations (Dillon and others 2005; Grissino-Mayer 

and others 2004; Meyer and others 2003; Veblen and Donnegan 2005). However, 

under natural fire regimes, fire is a rare occurrence in alpine and subalpine areas 

(Sherriff and others 2001; Sibold and others 2006). As a consequence, the fire 

disturbance regime appears to be within the HRV for high-elevation brooks and 

springs.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 76

Beaver abundance—Beavers typically inhabit streams large enough to support 

their dams and with a sufficient base of suitable plants available for food and dam 

construction. Total biomass of winter food cache plants, specifically woody species like aspen, willow, and cottonwood, may be limiting in many high-elevation 

sites (Allen 1983). Beavers do inhabit ponds in subalpine fens but typically at 

low densities. This suggests that anthropogenic impacts to beaver populations 

and associated ecological effects were likely smaller than those seen along larger 

streams and at lower elevations.

HRV assessment for high- to middle-elevation riparian 

shrublands

Composition and stand structure—Shrub-dominated riparian ecosystems are 

one of the most distinctive features of middle- to high-elevation valleys throughout the region (Comer and others 2003). Although a variety of plant associations 

have been described based upon floristics (Carsey and others 2003; Walford and 

others 1997), ecological processes are generally similar across these types. At the 

highest elevations, low stature willow stands are common due to the very heavy 

snow load and are typically dominated by Salix wolfii and S. planifolia, with 

herbaceous associates, including Calamagrostis canadensis and Carex aquatilis

(Carsey and others 2003). At lower elevations, community stature typically increases and a different suite of woody species become dominant, including Salix

monticola, S. boothii, S. drummondiana, S. geyeriana, and S. bebbiana. Common herbaceous associates are Calamagrostis canadensis, Heracleum maximum

and Geum macrophyllum, and in former beaver ponds, Carex aquatilis and C. 

utriculata are common (Carsey and others 2003).

A variety of anthropogenic impacts have placed many middle-elevation riparian shrublands out of the HRV for vegetation composition. At lower elevations, 

exotic species may be common due to amenable environmental conditions for 

their spread and greater levels of anthropogenic disturbance. Notable exotics 

are Breea arvense and a variety of introduced pasture grasses, such as Agrostis

gigantea, A. stolonifera, Dactylis glomerata, and Bromopsis inermis. Although 

no invasive woody species occur, non-native species herbs are common in the 

understory of many riparian communities.

Significant departures from historical conditions have been caused primarily by 

livestock and high native ungulate population herbivory and trampling. Both 

can affect vegetation structure by direct browsing of mature plants, reduced 

recruitment of seedlings due to trampling effects, and complex interactions with 

beaver, which can result in the conversion of tall (>21 ft) morphotype willows 

to short (approximately 3.3 ft) morphotype plants that are insufficient for dam 

creation (Baker and others 2005a,b; Brookshire and others 2002; Peinetti 2000; 

Peinetti and others 2001; Schulz and Leininger 1990). In addition to ungulate 

impacts, cover of willows has decreased along many rivers due to direct clearing (Figure 15).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 77

Figure 15— (A) 2005 illustrating changes in willow riparian communities from land use, (B) Tenderfoot Creek, Blackhills, 

SD, about 7 miles north of Custer, SD., in 1890.

Hydrologic regimes—Hydrologic processes in many high- and mid-elevation 

riparian areas have been altered by anthropogenic activities and now likely lie 

outside of the HRV. The primary cause is flow regulation from dams and diversions (Baxter 1977; Nilsson and Svedmark 2002; Pringle 2000; Wohl 2005). 

A large number of dams occur in mountainous parts of Region 2, although the 

effects of these hydrologic changes on riparian vegetation are poorly known 

when compared with riparian vegetation along large, low-elevation rivers. Specific hydrologic effects of these dams vary depending on size and management 

but can include reduced frequency and magnitude of peak flows increased base 

flows, and a reduction of flow variability. Reduced beaver populations relative 

to historical levels have also likely affected the hydrologic functioning of many 

mid- to high-elevation riparian areas.

Woody plant establishment—The spatial and temporal dynamics of willow 

and other riparian shrub dynamics are complex. Establishment events can occur regularly along some streams and some landforms, for example, along the 

margins of meander bends or episodically associated with low frequency event 

like high magnitude floods, oxbow formation, or beaver dam failure (Cooper 

and others 2006; Dickens 2003; Gage and Cooper 2005) (Table 15). Because 

establishment is intimately tied to basic hydrologic processes, changes in hydrologic regimes and beaver populations seen in many portions of the region have 

likely altered woody plant establishment dynamics, placing them outside of the 

HRV. More research, however, is needed to understand whether such departures 

are widespread.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 78

Table 15—Main patterns of willow establishment along the Big Thompson River in Rocky Mountain National Park 

in relation geomorphic and hydrologic characteristics (Cooper et al. 2006).

Geomorphic 

process Landform Effective flood regime Willow establishment patterns

Meandering	 Point bar	 Moderate to high flows	 •	 Continuous establishment

(2–5-year return interval flow)	 •	 Continuous establishment at

intermediate elevations above

channel 

Beaver pond	 Abandoned	 Flood event (>5-year	 •	 Periodic establishment prior

drainage beaver pond return interval flow) to dam breach

•	 Brief episodes of widespread

establishment following 

abandonment 

•	 Establishment at high elevations

relative to active channel 

Channel	 Abandoned	 Moderate to high flows	 •	 Continuous establishment over

avulsion channel (2–5-year return interval flow) limited period of time (as oxbow fills) 

•	 Establishment at low to moderate

elevations relative to active channel

Salix spp., and other members of the family Salicaceae such as Populus spp., 

disperse seeds for a brief period in early summer, usually timed to follow peak 

discharge along snowmelt dominated streams (Gage and Cooper 2005; Karrenberg 

and others 2002; Moss 1938). These seeds exhibit high initial germinability; 

however, their viability rapidly decreases over time and if seeds do not reach 

suitable bare, moist substrates for establishment soon after dispersal, the seed 

desiccates. Where the timing and magnitude of floods has been altered due to 

river regulation, willow establishment may be reduced or eliminated.

Excessively high populations of native ungulates have altered establishment 

dynamics in some areas of the Rocky Mountain region. For example, elk browsing has been identified as a primary cause of willow decline in montane willow 

carrs in Rocky Mountain National Park (Dickens 2003; Peinetti and others 2001; 

Peinetti and others 2002a). High densities of elk may reduce opportunities for 

establishment in several ways. Where elk numbers are high, they may trample 

or browse nearly all seedlings. In addition, willow aments form in the spring 

on last year’s stem tissue and browsing that removes young stems eliminates 

tissue that could bear aments, thereby reducing or completely eliminating seed 

production (Kay and Chadde 1992). Willow seed rain levels drop off steeply 

with increasing distance from seed-producing plants and heavily browsed areas 

have been shown to have seed rain levels two orders of magnitude lower than 

lightly browsed stands (Gage and Cooper 2005).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 79

Fire regimes—Although fire regimes in riparian areas typically differ in basic 

characteristics from adjacent uplands, ignitions in riparian areas are rare, so 

fires must spread from the surrounding landscape (Dwire and Kauffman 2003). 

Changes in fire regimes in upland forests documented in some portions of the 

region may have indirectly affected the frequency of fires in riparian areas, pushing them outside of the HRV. However, what little evidence there is suggests that 

fire’s role in mid- to high-elevation shrub carrs is minor, so such changes may 

not have much consequence for riparian vegetation communities. Wolf (2004) 

found that the 1988 fires in Yellowstone National Park triggered the largest willow recruitment event during 1989 in recent decades, likely due to upland erosion 

and sediment deposition in riparian areas.

Beaver abundance—Beaver are keystone species in middle- to high-elevation 

shrublands, strongly influencing shrubland structure and function (Naiman and 

others 1988). Beavers utilize willows for food and dam building and can significantly affect the structure of riparian communities, reducing Salix spp. stature and 

favoring non-preferred woody species like Betula occidentalis. However, their 

greatest effect on riparian communities is a result of their modification of hydrologic regimes. Beaver dams can inundate large areas of floodplains, significantly 

affecting biogeochemical processes, sediment dynamics, and groundwater flow 

patterns. For example, Westbrook and others (2006) found that a beaver dam 

complex on the Colorado River influenced the surface inundation, groundwater 

levels, and groundwater flow patterns over a greater area and for a longer duration than a flood with a 10-year recurrence-interval when the dam was absent.

Historically, beaver would likely have occurred in nearly all major drainages 

supporting riparian shrublands. Most of these watersheds were affected by beaver trapping in the Nineteenth Century, and while populations have recovered in 

some areas, overall numbers are still well below historical levels (Wohl 2001). 

Thus, it is likely that measures related to beaver, such as the density of beaver 

and beaver ponds, is likely outside of the HRV for many watersheds in the region.

HRV assessment for high-gradient riparian forests at 

medium elevations

Composition and stand structure—Forested riparian communities dominate 

streams in subalpine and montane areas of the region, particularly along highergradient reaches with narrow channels (Peet 1981). Although specific vegetation 

types vary with elevation and latitude, riparian areas throughout the region share 

similar structural and functional characteristics.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 80

In subalpine areas, dominant overstory species include conifers that are typical of 

the surrounding upland forests, including Abies lasiocarpa and Picea engelmannii, and deciduous species such as Populus tremuloides (Peet 1981). Although 

overstory dominants may not significantly differ from surrounding forests, these 

riparian forests typically support hydrophytic or mesophytic shrubs not found 

outside of riparian zones, including species of Salix and Alnus, as well as herbaceous species such as Mertensia ciliata (Carsey and others 2003). In lower 

elevation montane areas, overstory dominants include Populus angustifolia and, 

in the southern part of the region, Picea pungens.

In general, mountains have fewer non-native plants than lowland areas, perhaps 

due to the greater prevalence of National Parks and Forests, which have had less 

intensive land uses compared with agricultural and urbanized landscapes (Parks 

and others 2005). However, exotic species can be locally abundant, particularly 

where sites have been disturbed. The most common exotics are introduced pasture 

grasses such as Phleum pratense (Carsey and others 2003) and widely distributed 

herbaceous dicots such as Taraxacum officinale and Verbascum thapsus (Carsey 

and others 2003; Parks and others 2005). These species may contribute minor 

cover and, unlike lower-elevation riparian ecosystems, no exotic tree or shrub 

species occur in montane or subalpine riparian forests. As a consequence, most 

riparian forests are likely within the HRV for most measures of species composition, although the lack of vegetation composition data for pre-Euro-American 

periods reduces our confidence in this assessment.

We found little evidence to suggest that the age and size structure of communities is outside of the HRV for most Region 2 riparian forests, although the lack 

of research conducted along montane and subalpine streams prevents us from 

making definitive conclusions. Certainly, structure in areas that are subject to 

more intensive anthropogenic impacts, such as logging or tie-drives, may have 

pushed conditions outside of the HRV for specific areas (Wohl 2005). Likewise, 

where streams have been highly modified by flow regulation, altered woody plant 

establishment and recruitment may have resulted in stand and age structures 

outside of the HRV. Areas subjected to intense grazing by livestock brought to 

the Rocky Mountains to support the burgeoning mining population (Veblen and 

Donnegan 2005) may also exhibit compositional or structural differences from 

historical conditions. In general, contemporary stocking rates are well below 

those seen at the peak of grazing in the early-Twentieth Century, so many areas 

that were degraded in the past may have recovered.

For forest communities in narrow valley settings, age and size structure may be 

driven in large part by disturbances that originate outside of the riparian zone 

such as fire and flooding. Thus, where altered landscape disturbance regimes 

have pushed upland forests outside of the HRV, the structure of riparian communities may also be affected. This is more likely to be the case in lower elevation areas that historically had more frequent fires than in the subalpine, where 

pre-Euro-American fire return intervals were low (Aplet and others 1988; Dillon 

and others 2005; Sherriff and others 2001; Sibold and others 2006; Veblen and 

Donnegan 2005).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 81

Hydrologic regimes—Montane and subalpine forests experience seasonal 

floods driven by spring snowmelt, although peak flows may also occur following convective summer thunderstorms. Such storms are particularly prevalent in 

the southwest portion of Region 2 (Baker 1990; Fleener 1997). Many mountain 

streams, particularly at high elevations, are gaining with respect to groundwater. 

Consequently, vegetation may be only weakly dependent on surface water levels 

in streams. The importance of flood regimes on vegetation dynamics may be 

linked more to flood influences on processes of plant establishment. Here, high 

magnitude, low-frequency events are of primary importance, as only these have 

sufficient power to create disturbance patches important for seed germination of 

species like willows (Karrenberg and others 2002).

Although most research and discussion on river regulation has occurred on large 

rivers, many small to moderate sized streams have also been modified (Pringle 

2000). The hydrologic regimes of many streams have been pushed outside of 

the HRV (Richter and others 1996). Effects include reductions in the magnitude 

of peak flows, reduced mean flows, altered flood timing, and changes in water 

thermal characteristics (Baxter 1977). Specific effects differ as a function of a 

variety of factors, such as the size and location of the dam or diversion, the hydrologic characteristics of the impacted stream, and management characteristics 

including the seasonality of water releases and diversions.

Woody plant establishment—The greatest impact of flow regulation on montane 

and subalpine riparian forests is likely through effects on woody plant establishment. Although little research has been conducted on woody plant dynamics in 

montane riparian communities, the existing studies suggest that establishment 

is episodic. In southwest Colorado, Baker (1990) found that stands of Populus 

angustifolia originated in ten to thirteen discrete periods between the midNineteenth and late-Twentieth Centuries (Baker 1990).

Years with high seedling establishment may be associated with broad scale climate factors. For example, Baker (1990) identified winter conditions in the North 

Pacific and a persistent late summer Arizona monsoon as important variables in 

the establishment of Populus angustifolia in southwest Colorado. His models 

suggested that good seedling years occurred more frequently (approximately 

every 3.4 years) than stand-origin years (approximately 10-15 years) and that 

good seedling years and stand-origin years were more common from 1848 to 

1985 than from 1556 to 1848 due to more frequent cool, wet years since 1848 

(Baker 1990).

A variety of studies have examined the flow requirements for the establishment 

of cottonwood species, including Populus angustifolia. Along smaller montane 

streams and relatively large semi-arid rivers, only floods of sufficient magnitude 

can create nursery sites for seed germination and they must be timed before seedling release (Mahoney and Rood 1998; Samuelson and Rood 2004). In addition, 

it is critical for both Salix spp. and Populus spp. establishment that sufficient 

soil water is available through the summer to prevent drought-induced mortality 

(Gage and Cooper 2004a; Kranjcec and others 1998; Woods and Cooper 2005).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 82

Fire regimes—Fire regimes in many portions of the region are outside of the 

HRV, particularly at lower elevations with relatively active historical fire regimes 

(Dillon and others 2005; Grissino-Mayer and others 2004; Meyer and others 

2003; Veblen and Donnegan 2005). Fire regimes in riparian areas, particularly 

narrow, forested ones, are driven by the fire regimes in surrounding landscapes 

(Dwire and Kauffman 2003). Where significant departures from the HRV of 

upland fire regimes have occurred, they have likely also affected riparian areas.

Beaver abundance—As detailed elsewhere in this assessment, changes in beaver abundance are one of the primary anthropogenic impacts forcing riparian 

ecosystems outside of the HRV. However, beavers typically favor streams with 

gentle gradients and in wide valley settings (Allen 1983); thus, it is likely that 

historical levels of beaver were low to begin with in the high gradient riparian 

forests at middle and high elevations. As a result, current beaver densities may 

be within the HRV for most streams, although departures from the HRV have 

likely occurred in some individual streams.

HRV assessment for low-gradient riparian forests and 

shrublands and middle and low elevations

Composition and stand structure—Because of the varied hydrologic regimes 

and geomorphic settings characterizing low-elevation riparian ecosystems in 

the region, there is considerable diversity among riparian forest and shrub communities. Native overstory dominants vary, but typically include species such 

as cottonwoods (Populus deltoides in the Great Plains, P. fremontii west of the 

continental divide, and P. balsamifera ssp. trichocarpa in the northern portion 

of the region) and willows (Salix exigua, S. goodingii). A variety of other native 

tree and shrub species in a number of genera may be present including Fraxinus, 

Quercus, Ulmus, Chrysothamnus, Symphoricarpos, Amelanchier, Rhus, Cornus, 

Prunus, Shepherdia, Rosa, Sarcobatus, and Purshia spp. (Carsey and others 2003; 

Comer and others 2003b). Herbaceous components vary as well depending on 

particular vegetation association.

Compositionally, many of the riparian areas in this group are outside of the 

HRV due to the presence of non-native species. For example, Tamarix spp. and 

Elaeagnus angustifolia are widespread throughout the region, occurring along 

streams of all sizes and flow regimes (Di Tomaso 1998; Friedman and others 

2005; Lesica and Miles 2001; Pearce and Smith 2003; Stein and Flack 1996). 

These two species represent the third and fourth most frequently occurring woody 

riparian plants along low-elevation streams in the West (Friedman and others 

2005). Exotic herbaceous species are also common and range from non-native 

pasture grasses that were purposely introduced such as Poa pratensis and Bromopsis inermis, to noxious weed species such as Bromus tectorum, Euphorbia

spp., and Breea arvense (DiTomaso and Healy 2003; Whitson and Burrill 2000). 

As an example of the near ubiquity of exotic species along low-elevation rivers, 

exotic species comprised one-third or greater of all species observed in plots 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 83

along both a regulated and unregulated river in northwest Colorado (Uowolo 

and others 2005). The rapid spread of exotic shrubs along rivers, upwards of 

approximately 20 km/year (Graf 1978), has been attributed to several factors. 

For example, Tamarix produces abundant, viable, aerially dispersed seed, and 

vast numbers of seedlings may occur in suitable habitat; these occupy similar 

habitats as cottonwood and willow.

In general, exotic species richness is greater in more species-rich communities 

(Stohlgren and others 1999; Stohlgren and others 1998). For example, 63 percent 

of the variance in exotic species richness was explained by a positive relationship 

with native species richness along the Green and Yampa Rivers in Northwest 

Colorado (Uowolo and others 2005). However, results of that study did not 

indicate that river regulation increased the probability of invasion in contrast to 

studies conducted in other regions (Decamps and others 1995). Overall patterns 

of species richness differ among age classes within and between regulated and 

unregulated rivers. Along both kinds of rivers, species richness was greatest in 

the youngest age classes and declined as stand age increased (Uowolo and others 2005). Thus, major changes in age structure of riparian communities may 

indirectly affect diversity at landscape scales.

In addition to non-native species, changes in hydrologic regimes have allowed 

the invasion of flood-intolerant shrub species into riparian communities. For 

example, Merritt and Cooper (2000) described channel narrowing processes 

along the regulated Green River in northwest Colorado (Figure 16) and noted 

the increasing presence of flood-intolerant species such as rabbitbrush (Chrysothamnus nauseosus) and greasewood (Sarcobatus vermiculatus). These shrubs 

are highly sensitive to soil saturation (Groeneveld and Crowley 1988) and are 

only able to survive because of the elimination of high peak flows (Merritt and 

Cooper 2000; Williams and Cooper 2005).

Figure 16—Images of Lodore Canyon in 1870 (A) and present (B) illustrating vegetation changes driven by Tamarix 

invasion.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 84

The vertical structure of many low-elevation riparian communities in Region 2 

is also outside of the HRV due to changes in canopy height, architecture and 

spatial arrangement. The invasion and increasing dominance of exotic shrubs 

like Tamarix spp. and Elaeagnus angustifolia have altered the structure of many 

riparian communities, resulting in closed canopy stands that are more dense than 

native stands thereby reducing light to the forest floor and suppressing seedlings 

of full-sun plants like Populus spp. (Cooper and others 1999). In addition, the 

widespread failure of native species to regenerate due to factors such as river 

regulation (Rood and others 1995) has altered the structure of riparian communities 

(Cooper and others 1999; Shafroth and others 2002; Williams and Cooper 2005).

Where extreme reductions in peak flow, river stage, floodplain water table levels, 

and soil water recharge occur, severe water stress may develop in riparian trees, 

causing the dieback of leaves, branches, and roots in the upper soil profile as 

well as plant mortality (Amlin and Rood 2003; Williams and Cooper 2005). The 

sacrifice of branches appears to be an adaptive response to drought, improving 

the water relations of remaining branches as leaf area adjusts to available water 

(Horton and others 2001; Rood and others 2000, 2003a). Xylem cavitation and 

increased stomatal closure followed by leaf senescence and branch death has been 

reported for Populus deltoides and P. fremontii in response to drought; however, 

P. trichocarpa, P. angustifolia, and P. balsamifera experience the patchy summer 

branch senescence typical of P. deltoides and P. fremontii less frequently (Rood 

and others 2000).

Changes in root distribution and abundance have also been documented along 

regulated rivers. For example, Williams and Cooper (2005) reported that Populus fremontii root density was highest at depths of 2.5 to 4.9 ft (0.75 to 1.50 m) 

along the unregulated Yampa River versus 5.7 to 7.4 ft (1.75 to 2.25 m) along 

the regulated Green River. They also found that the density of roots in the upper 

7.4 ft of soil was more than double at the unregulated than the regulated site. In 

the unregulated site, the soil depths with the highest root density corresponded 

to the average peak water table elevation for the previous 10 years and where 

loam-textured soil layers with high water-holding capacity occurred, while the 

vertical distribution of roots at the regulated site was relatively uniform with 

depth (Williams and Cooper 2005).

In addition to reach-scale changes in structure due to factors such as grazing, 

Tamarix spp. invasion, and poor Populus spp. regeneration, the landscape structure, including the size and distribution of any low-elevation riparian systems 

has changed relative to historical conditions. Many riparian forests are highly 

fragmented due to the effects of land use changes (Table 16). In many areas, the 

average patch area of riparian forests has declined along larger rivers.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 85

Hydrologic regimes—The hydrologic regimes of low-elevation riparian ecosystems have been significantly altered by flow regulation, and many hydrologic 

variables are now outside of the HRV (Patten 1998; Rood and Heinze-Milne 

1989; Williams and Cooper 2005). Nearly all major rivers in the region have been 

dammed, altering the natural flow regimes (Graf 1999; Patten 1998; Poff and 

others 1997). Intermittent and ephemeral drainages have also been impacted by 

dams, although not as systematically as perennial rivers. Many of these streams 

have had dugouts or livestock impoundments created in-channel to intercept 

flows (Smith and others 2002).

Multiple examples of hydrologic effects of flow regulation on riparian communities can be found across the region. For example, reduced magnitude and 

frequency of floods and an increase in base flow levels has been documented 

along the regulated Green River, reducing overall flow variance and leaving 

soils on former floodplains in a state of perpetual drought. These changes have 

facilitated the invasion of drought-tolerant exotics such as Tamarix spp. and 

flood intolerant species such as rabbitbrush, and they have hastened the decline 

of mature Populus spp. trees on floodplains and allowed for the development of 

marsh communities on bars (Merritt and Cooper 2000). Along the Platte River, 

reductions in peak flow (Figure 17) have initiated channel adjustments, including 

narrowing of the river, which is correlated with the expansion of riparian forests 

relative to historical conditions (Johnson 1994).

Groundwater pumping has impacted cottonwood forests in some areas. Groundwater pumping can reduce water table depths, resulting in severe drought stress, 

branch cavitation and dieback, and, in some instances, tree death. Along the South 

Platte River in Colorado, Populus deltoides trees responded to abrupt alluvial 

water table decline with decreased shoot water potential followed by leaf mortality (Cooper and others 2003c). Xylem pressure potentials in stems and rates of 

leaf loss were significantly related to the magnitude of water table declines, and 

impacted trees experienced crown dieback caused by xylem cavitation (Cooper 

and others 2003c)

Table 16—Proportions of the floodplains in each of seven land cover categories–Cottonwood Forest (C), Agriculture (Ag), Industry (I), Roads (Rd), Urban (U), Open Water (W), and Other (O)–delineated by Northcott et al. (2006) 

for the Upper Colorado River Basin.

River Proportion of total floodplain Proportion of

regulation area in land cover category (%) floodplain area

Mainstem river regime C Ag I Rd U W O developed (%)

White 0 58 32 1 0 0 0 7 34

San Miguel 0 60 22 6 1 5 1 5 34

Yampa 0 19 52 0 0 0 0 29 52

Lower Dolores 1 38 4 0 0 0 0 48 28

Lower Colorado 1 61 30 3 0 0 0 7 33

Upper Colorado 1 47 31 14 2 3 0 3 50

North Fork of the Gunnison 1 39 51 8 1 1 0 2 60

Lower Green 1 38 27 0 0 0 31 3 60

Upper Dolores 2 40 5 0 0 0 0 55 5

Upper Green 2 70 4 0 0 0 19 7 23
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 86

Vegetation establishment and successional dynamics—Several conditions 

must be met for cottonwood and willow establishment, including: the presence 

of suitable substrates for seed germination, adequate levels of seed rain, and sufficient soil moisture during the late summer months when seedlings are vulnerable to water stress (Amlin and Rood 2002; Cooper and others 1999; Sacchi and 

Price 1992). Although cottonwood seedlings can appear over a wide elevational 

gradient following a flood event, seedlings may require three to four years to 

grow roots to the late summer water table, during which they are susceptible to 

mortality from desiccation (Cooper and others 1999). In addition, floods in the 

years following initial seedling establishment can influence patterns of seedling 

recruitment into successive age and size classes. For example, very high peak 

flows in the years immediately following seed germination may wash away 

seedlings in low topographic positions.

The timing of dispersal of cottonwood and willow seeds is also critical. Seed 

dispersal typically coincides with the timing of peak snowmelt-driven floods, 

which, for many rivers, occurs in late May to early June. Although the viability 

of seed is usually very high (Cooper and others 1999; Gage and Cooper 2004a), 

viability rapidly drops following dispersal; thus, these species form no soil seed 

bank. Where flow regulation has altered the timing of floods, establishment may 

be constrained due to a lack of synchrony between dispersal and the availability 

of suitable substrates for seed germination.

Figure 17—Instantaneous peak flow at North Platte River at North Platte, NE (USGS gauge 

#06693000). Note the sharp decrease in flow variability and peak flow magnitude following 1940.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 87

Some life history characteristics of exotics may influence their spread along 

many rivers. For example, Tamarix spp. produces seed for a longer proportion 

of the summer than cottonwoods and willows (Cooper and others 1999). These 

characteristics along with the species’ greater drought and salt tolerance relative 

to native woody plants, have allowed Tamarix spp. to thrive, particularly along 

rivers where climate change and flow regulation have altered flood regimes.

Establishment dynamics of woody riparian species are complex and vary naturally 

among species and rivers. Establishment events can be either gradual or episodic, 

resulting in the varied age and size structures often seen in riparian communities. For example, cottonwood establishment can occur associated with the 

formation of point bars along meandering channels in an incremental process 

dominated by moderately high flows occurring every few years (Shafroth and 

others 1998) (Figure 18). Where establishment occurs in such a fashion, stands 

may develop with relatively varied age and size structure along a given stream 

reach (Johnson and others 1976; Merritt and Cooper 2000). Along meandering 

channels, the rate of meandering directly influences the relative abundance and 

spatial distribution of the bare sites required for Populus spp. and Salix spp. seed 

germination (Bradley and Smith 1986; Johnson 1994). In contrast, establishment 

along ephemeral streams may more closely match the history of extreme floods 

(Friedman and Lee 2002). Because of the dependency of many riparian species 

on floods for establishment, establishment dynamics on regulated rivers are 

typically outside of the HRV (Cooper and others 1999; Katz 2001; Nilsson and 

Svedmark 2002; Stromberg 1998).

Vegetation dynamics along low-gradient riparian ecosystems at low and intermediate elevations can be complex, with a variety of physical and biological factors 

influencing community development and successional patterns. Fluvial processes 

are central in creating the bare alluvium required for cottonwoods and willows 

to establish from seed (Karrenberg and others 2002); although once established, 

the specific successional trajectory of stand can vary.

Figure 18—Schematic illustrating processes of 

lateral channel migration and meander cutoff 

(modified from Richter and Richter 2000). Establishment of cottonwoods and willows may occur 

more regularly as a consequence of channel 

migration and point bar formation (green areas), 

as new sites for establishment are created by 

the erosion of stream banks on the outsides of 

meander bends and the deposition of transported 

sediments on point bars. Rare events, such as 

channel abandonment or beaver pond failure, 

may open large areas of suitable substrate for 

establishment, allowing for the formation of large 

cohorts of seedlings (Cooper and others 2006; 

Westbrook and others 2006).
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Figure 19—Flow chart illustrating a conceptual model of riparian patch succession and woody plant establishment 

along the Yampa River (re-drawn from Richter and Richter 2000).

Richter and Richter (2000) developed a model to describe the development and 

successional dynamics of Populus angustifolia communities along the Yampa 

River in Colorado (Figure 19). In addition, their model includes a variety of 

shrub and herbaceous communities, which may co-occur as distinct patches 

on the landscape. Hydrologic regime, including both flooding and water table 

dynamics, is the key determinant of the particular patch types and their relative 

abundances on the landscape. Although this model was developed for the Yampa 

River, the model’s overall structure is more widely applied. The specific patch 

types and transitions vary depending on factors such as river size, flood regime, 

regional floristic composition, and elevation. Along intermittent and ephemeral 

streams, fluvial marsh patches are not present in the model, while xeric communities may be important components.
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Because floods are the main disturbance agent creating conditions for new 

Populus spp. establishment, the frequency and magnitude of floods are central to 

the successional dynamics of lowland riparian areas. Along unconfined alluvial 

rivers, geomorphic processes such as channel meandering are also important; as 

channels migrate across floodplains, cottonwood patches may become isolated 

from the physical effects of flooding and stands may persist throughout succession, ultimately forming late-successional cottonwood gallery forests (Johnson 

and others 1976; Stromberg 1997). As a result, stands with a range of ages may 

be present at any given time across a gradient from channel margin to high 

floodplain surfaces (Lyttle and Merritt 2004).

At broad scales, hydrologic and geomorphic processes are influenced by watershed characteristics such as watershed size, vegetation cover, and land use. 

These characteristics explain, in part, differences in vegetation dynamics among 

rivers. Three ephemeral stream types, based on basin and flow characteristics, 

were identified in Arizona (Shaw 2006). Type I streams drain small watersheds, 

are highly vulnerable to flood disturbance, and have xeric alluvial groundwater 

regimes. Streams of this type are typically occupied by upland plant associations. 

Type II streams have larger watersheds, more moderate shear stresses, more 

persistent alluvial groundwater, and support a mix of upland and mesophytic 

riparian species such as cottonwoods. Type III reaches drain large watersheds 

and have hydrologic regimes driven mainly by upstream hydro-climatic conditions. These streams have the lowest bankfull shear stresses and perennial water 

tables that span expansive floodplains and support riparian tree and shrub communities (Shaw 2006).

Channel narrowing has occurred along many regulated rivers, altering geomorphic 

and hydrologic characteristics and vegetation. As vegetation becomes established 

in the active channel, plants may facilitate sediment accretion by increasing 

hydraulic roughness and increasing sediment deposition rates. Such changes 

may accelerate the establishment of exotic species, as has been suggested for the 

lower Green River (Allred and Schmidt 1999; Graf 1978; Merritt and Cooper 

2000). Although Populus spp. have been declining along many rivers that are 

experiencing channel narrowing, narrowing has been linked to the expansion of 

Populus deltoides forests along the Platte River (Johnson 1994).

In addition to flow regulation, cottonwood regeneration rates are typically reduced 

in areas subject to intense human activities such as land clearing associated with 

agriculture. For example, Northcott and others (2006) found that the widespread 

anthropogenic development of riverine floodplains in the upper Colorado River 

basin may be as significant of a factor in the low levels of Populus spp. regeneration observed throughout the region. The authors found that the likelihood of 

young Populus fremontii cohorts along floodplains was reduced by 65 percent 

by development (Northcott and others 2006), and they suggested that riparian 

forests may also be in jeopardy along unregulated rivers.
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Fire regimes—Altered fire regimes in upland areas have likely pushed riparian 

fire regimes outside of the HRV for some streams, although little research has 

been directed to the question. Structural and composition changes in riparian 

vegetation (e.g., Tamarix spp. invasion) have also likely altered the dynamics 

of fire in many low-elevation riparian areas (Busch and Smith 1993; Di Tomaso 

1998). However, any changes are likely highly coupled to changes in upland fire 

regimes and are expected to vary widely across the region.

Beaver abundance—The widespread trapping of beaver impacted low-elevation 

streams as well as higher mountain ecosystems. Slow recovery of populations 

in many portions of the region suggests that population densities may still be 

outside of the HRV in some areas. However, the effects of such changes on 

riparian area function is poorly understood along larger rivers where beavers 

do not build dams but rather den in river banks (Breck and others 2001, 2003).

Management Opportunities and Constraints _______

Management opportunities

As knowledge of the impacts and threats to riparian resources has increased, so 

too have opportunities for improved management. However, more area-specific 

research is needed to improve our understanding of the abundance, distribution, 

and condition of riparian areas at the scale of individual National Forests. By 

developing a more thorough accounting of the kinds of historical and contemporary impacts affecting a given stream, specific management prescriptions can 

be developed.

Because so many riparian areas are functionally impaired, many opportunities 

exist for restoring and enhancing riparian areas in the region. The scientific 

foundation of riparian restoration has improved over the years, and guidance can 

be found for many of the specific riparian types in Region 2 (Allen and others 

1997; Hughes and others 2005; Hughes and Rood 2003; Rood and others 2003b; 

Wohl 2005). However, there are many constraints that limit management and 

restoration options. The prior appropriation doctrine governing water allocation 

in states like Colorado has not traditionally recognized in-stream flows for ecological purposes. Societal demands for water associated with regional population 

growth are likely to increase as are periods of increased scarcity resulting from 

climate change (Baron and others 2002; Patten 1998; Pringle 2000).

A broad goal should be to identify systems where ecological integrity is relatively intact and develop means to ensure the protection of these ecosystems as 

centers of natural biodiversity, management models, and benchmarks for future 

ecological research (Frissell and Bayles 1996). Broad-scale inventories of riparian area condition can identify sites where dams or diversions are present but are 

unneeded or inefficient. There are numerous small earthen dams in high basins 
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throughout the region (Pringle 2000), many of which occur on Federal lands. 

Built early in the period of Euro-American settlement, many of these have functionally been replaced by larger engineering structures downstream and may be 

ideal candidates for removal.

More quantitative analyses of flow regimes on regulated and unregulated rivers 

are needed. Richter and others (1996) provided a useful framework for such 

analyses, using specific indicators of hydrologic alteration based upon available hydrologic data from existing measurement points such as stream gauges 

or model-generated data in which 32 parameters, organized into 5 groups, are 

used to statistically characterize hydrologic variation within each year. These 

metrics can be used to evaluate the effects of hydrologic perturbations associated 

with activities such as dams and stream diversions (Richter and others 1996). 

Application of an objective, quantitative approach to characterizing anthropogenic impacts to riverine systems would provide an effective tool for identifying 

impaired systems and prioritizing restoration activities.

Restoration

Due to the high connectivity of streams and riparian ecosystems, impacts from 

dams can impact riparian areas distant from the source of the impact. Riparian 

restoration, even of small sites, requires a watershed approach (Wohl and others 

2005). Restoration conducted in a watershed context demands critical examination of geomorphic and hydrologic processes and vegetation at multiple scales. 

Harris and Olson (1997) describe four key spatial scales relevant to riparian restoration: (1) the entire stream and its associated floodplain; (2) the stream reach; 

(3) the plant community; and (4) the specific site (physical setting). Different 

actions can be taken at each of these scales to improve riparian area condition 

and function. For example, at the broadest scales, management of flows along 

regulated rivers can be improved to better mimic historical flow characteristics. 

At intermediate and fine spatial scales, a variety of approaches can be taken to 

identify good sites for restoration. O’Neill and others (1997) used a GIS and 

color-infrared aerial photographs to evaluate sites for restoration potential on the 

Arkansas River in Colorado. The authors found that at the reach scale, specific 

stream power was a useful index for discriminating between high-energy reaches 

with low restoration potential and low-energy sites with high restoration potential 

(O’Neill and others 1997).

GIS analyses in conjunction with field data collection have been used elsewhere 

to evaluate the degree of impact to riparian ecosystems and to identify restoration 

candidates (Russell and others 1997). For example, Northcott and others (2006) 

used GIS and aerial photo analyses to evaluate the regeneration of Populus fremontii forests along the upper Colorado River basin in relation to land use and 

river regulation. Such approaches may be useful in other watersheds.
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Because of the many physical and biological variables important to riparian area 

structure and function, data must be collected and incorporated into restoration 

plans. Particularly important elements are identification of the geomorphic and 

hydrologic characteristics of impacted streams and, where available, reference 

streams. Failure to properly characterize hydrologic regimes is perhaps the single 

greatest cause of failure in river restoration (Kondolf 1998; Kondolf and Micheli 

1995; Wohl and others 2005).

Flow regulation represents the single greatest anthropogenic impact to riparian ecosystems; thus, where possible, decommissioning unnecessary dams and 

diversions should be pursued. However, because of the increasing demands on 

water resources forecasted for the region, wide-scale decommissioning of dams is 

unlikely. Improving management of flows for riparian communities may represent 

a more achievable approach to restoring desired ecological characteristics back 

into the HRV. The use of controlled releases from dams to mimic historical floods 

has been conducted on several large rivers, with the principal goal of improving the viability aquatic and riparian ecosystems, particularly endangered fish.

An example is the controlled release from Flaming Gorge reservoir in 1999. 

High snowmelt runoff down the Green River filled Flaming Gorge Reservoir and 

necessitated a controlled release, which had a maximum discharge (10,948 ft3/s) 

that closely approximated the pre-dam mean annual peak flow and was the second 

highest flow since dam completion. Controlled releases have also been conducted 

on the Colorado River below Glen Canyon dam with some success (Patten and 

others 2001; Schmidt and others 2001). For example, a seven-day experimental 

controlled flood was conducted at Glen Canyon Dam in 1996; peak discharge 

was limited to 44,991 ft3/s, lower than pre-dam spring floods. Rood and others 

(2003) describe a restoration program along the Truckee River—aimed at halting 

the collapse of the Federally endangered cui-ui sucker (Chasmistes cujus), an 

endemic fish species—in which flows were increased during the spring to permit 

fish spawning. The higher flows were successful at increasing fish reproduction 

and resulted in extensive Populus fremontii and Salix exigua seedling recruitment. 

These results suggest that instream flow regulation can promote ecosystem restoration along other dammed rivers worldwide (Rood and others 2003b). Where 

water table decline rates can be made to mimic natural recession, establishment of 

native woody plant seedlings may be achieved (Auble and others 1994; Mahoney 

and Rood 1998; Rood and others 1998; Shafroth and others 1998).

Other factors that need to be addressed in restoration design are herbivory from 

livestock and native ungulates. Either can represent a significant limitation to 

restoration success unless addressed, for example, through the construction of 

exclosures (Brookshire and others 2002; Elmore and Kauffman 1994; Gage 

and Cooper 2004b; Opperman and Merenlender 2000). Population control of 

unnaturally high ungulate populations or modification of behavior through the 

reintroduction of predators may be a successful element of riparian restoration 

in some areas (Beschta 2003; Ripple and Beschta 2003).
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Future research—Although some topics in riparian ecology and associated 

disciplines like fluvial geomorphology and hydrology have been extensively 

studied, there are many topics for which little information has been collected. 

For example, while requirements for low-elevation Populus spp. establishment 

are well known, little research has been directed towards establishment dynamics of high-elevation willow carrs. Although studied for some species (Amlin 

and Rood 2001, 2002), more information regarding tolerable rates of watertable decline and flood depths is needed for many species. Where economic 

and political constraints allow, controlled releases from dams should be more 

extensively studied. Also unknown for many riparian community types is the 

minimum flow needed to promote establishment of native woody species. Such 

information would be useful for managers addressing riparian ecosystems in 

developing flow management.

We also need a better understanding of the effects of diversions on small order 

streams. Do plants respond differently to flow diversion along gaining versus 

losing streams? How do communities respond to small versus large impoundments? Although it has been the focus of some studies (Dominick and O’Neill 

1998; Henszey 1993), relatively little work has been done on the effects of flow 

augmentation. Hyporheic flows have been shown to affect cottonwood growth 

(Harner and Stanford 2003), but little is known about their effects on other native or exotic woody species.

Many studies have examined the effects of native ungulates on willows (Brookshire 

and others 2002; Manoukian and Marlow 2002; Maschinski 2001; Singer and 

others 1994; Zeigenfuss and others 2002), although most have been conducted 

on low- and mid-elevation community types. Less is known about effects on 

willows at higher elevations. Also poorly understood are the effects of moose 

on woody shrubs. Much of the existing literature regarding native ungulate effects involve elk (Baker and others 2005a; Baker and others 2005b; Peinetti and 

others 2001; Ripple and Beschta 2004; Singer and others 1998).

Management constraints

The prevalence of hydrologic changes to streams is probably the single greatest 

factor impacting riparian areas in the region, and represents the greatest constraint on managing rivers for ecological function. Despite greater recognition of 

the ecological impacts of dams and diversions by scientists, forecasts of future 

population growth and water demands make wide-spread dam removal unlikely 

(Baron and others 2002; Naiman and Turner 2000; Tockner and Stanford 2002).
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Climate change is an additional factor likely to impact riparian areas in the future, 

although the problem is largely beyond the scope of land managers. There is 

considerable uncertainty about its effects in the region. Increased regional temperatures and precipitation have been predicted under some scenarios, although 

the confidence in these predictions is low (U.S. EPA 1998a, 1998b, 1998c; Wagner 

2003). This underlying uncertainty regarding possible changes makes developing 

of mitigation plans difficult. One thing managers can do is to institute baseline 

data collection and monitoring efforts.
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5. Historical Range of Variation for 

Region 2 Marshes

Definitions and Concepts of Marshes _____________

The term marsh has been variously used by different authors. Keddy (2000) defined marshes as “wetland communities dominated by herbaceous plants rooted 

in hydric soils, but not peat”. In contrast, Mitsch and Gosselink (2000) applied a 

more general definition: “frequently or continually inundated wetlands characterized by emergent herbaceous vegetation adapted to saturated soil conditions.” 

Weller (1994) described marshes as “any low area that will hold water over soil, 

even temporarily, forming a suitable basin for the invasion of water-tolerant, 

rooted, soft-stemmed plants (hydrophytes) such as grasses, sedges, cattails, and 

bulrushes.” Each of these definitions emphasizes vegetation, soil characteristics, 

and hydrology, although in different terms. Marshes can generally be recognized 

by (1) the dominance of emergent, non-woody vegetation such as grasses, sedges, 

and rushes; (2) the presence of a shallow-water hydrologic regime; and (3) minor 

or non-existent accumulations of peat (Mitsch and Gosselink 2007).

Marshes in Region 2 are diverse functionally and compositionally and range 

from cattail-dominated wetlands on the plains to sedge-dominated ecosystems 

along the margins of montane lake basins. Additional examples include playas 

(depressional wetlands hydrologically supported principally by precipitation 

and surface runoff) and emergent plant communities found in abandoned meander bends or oxbows of large alluvial rivers. Though occurring along a wide 

elevational and latitudinal gradient and supporting diverse plant communities, 

marshes share some fundamental hydrologic attributes; these drive key ecological 

processes and effectively differentiate these systems from other wetland types 

such as fens or wet meadows.

Geomorphic Setting and Principal 

Ecological Drivers _____________________________

Marshes are broadly distributed throughout Region 2, occurring in all National 

Forest units. However, there are significant differences in the abundance, distribution, and functional characteristics of marshes within the region. Several 

ecological drivers are particularly important in creating these patterns. At broad 

regional and landscape scales, climate and basin physiography are most important, 

while at more localized scales, the morphology of individual basins supporting marshes is key, with each driver primarily influencing vegetation patterns 

and ecological function through their effects on hydrology (Kantrud and others 

1989a) (Figure 20).
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Differences in climate occurring along the region’s elevational and geographic 

gradients have a strong influence on the relative abundance of marsh ecosystems, 

their functional characteristics, and their characteristic plant communities. The 

specific climatic variables most important to marsh formation, persistence, and 

function are temperature and precipitation. These variables in turn determine 

whether conditions of moisture surplus or moisture deficit occur—a basic climatic control of marsh distribution and hydrologic function (Kantrud and others 

1989a,b; Winter and Rosenberry 1998).

Many marshes are hydrologically supported primarily by precipitation and surface water runoff (Mitsch and Gosselink 2007; Weller 1994; Winter and others 

2001) and are therefore particularly responsive to climatic fluctuations (Winter 

2000) (Figure 21). Orographic effects in mountainous areas strongly influence 

precipitation patterns (Kittel and others 2002), thereby influencing landscape level 

patterns of marsh abundance, structure, and function in mountainous portions 

of Region 2. Orographic effects are not a factor in areas with small elevation 

gradients such as the Great Plains, which generally receive lower total precipitation relative to higher-elevation areas at the same latitude. In addition, temporal 

patterns of precipitation are often more variable. As a consequence, water levels 

in marsh types, such as playas, that rely nearly exclusively on direct precipitation and runoff from surrounding areas are particularly variable, a central factor 

shaping vegetation patterns in these systems (Smith 2003; Weathers 2000).

Figure 20—Schematic illustration of broad-scale drivers of marsh formation, ecological function, 

and vegetation patterns (modified from Kantrud 1989).
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Figure 21—Dominant water sources supporting a prairie pothole marsh. Note that the importance of each hydrologic input or 

output varies from wetland to wetland (Kantrud et al. 1989a).

Basin physiography is an additional broad-scale driver of marsh abundance, 

structure, and function (Kantrud and others 1989a) (Figure 20). Variables of particular importance are the size of contributing watersheds, patterns of topographic 

relief, and lithology. Because of the region’s diverse climates and geological 

composition, the specific geomorphic settings in which marshes form vary. All 

marshes are found associated with topographic depressions, although the size, 

morphology, and origin of landforms differ widely. Marshes occur in small, barely 

noticeable depressions as well as in large lake basins. Since marsh communities 

commonly fringe open-water lacustrine systems, they are also found associated 

with the numerous man-made ponds and lakes in the region, ranging from large 

storage reservoirs to small livestock watering impoundments (Evans and Kerbs 

1977; Rumble 1989).
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Landforms supporting marshes are formed by a variety of mechanisms, including fluvial, eolian, and glacial erosion processes (Bolen and others 1989; Mitsch 

and Gosselink 2007; Weller 1994). Given the region’s semi-arid climate, overall 

wetland area is generally greater at higher elevations due to the cooler and wetter 

microclimates. However, the relative abundance and distribution of marshes is also 

strongly shaped by the physiographic and geologic characteristics of particular 

landscapes. For example, marshes occur at very high densities in glaciated portions of the Great Plains. The so-called “prairie pothole” region has a very high 

density of marshes, as does the playa lakes region (Kantrud and others 1989a; 

Smith 2003). In contrast, marshes are generally rare in many mountainous areas, 

where steep topography provides few sites suitable for marsh formation.

Past glaciation is an important factor influencing broad-scale patterns of marsh 

occurrence in many northern regions, but with the exception of the eastern portion 

of South Dakota, none of the states encompassing Region 2 was directly affected 

by continental glaciation. However, periodic expansion of glaciers originating in 

high mountain cirques have helped shape many alpine and subalpine landscapes 

in Colorado and Wyoming, often creating a high density of depressional basins 

filled with lakes or kettle ponds. Depending on basin physiography, they often 

support emergent marsh vegetation (Winter and Woo 1990).

Other geological factors that influence marsh distribution and function include 

the mineralogy and grain structure of geological strata underlying the wetlands. 

Lithology can influence basic hydrologic processes such as infiltration, groundwater recharge, and discharge, in part shaping individual wetland hydrologic 

regimes. Lithology can also strongly affect water chemistry parameters, including 

nutrient concentrations and salinity levels (Winter 2001). Lithology, along with 

climate and tectonics, also affect the physiographic characteristics of landscapes, 

thereby influencing marsh distribution and functional characteristics.

Also worth noting are the extensive dune fields located throughout the region, 

especially the Nebraska Sandhills, as well as smaller features found in Wyoming, 

Kansas, and Colorado. Because of sand’s high permeability and hydraulic conductivity, almost no surface runoff is formed during precipitation events. As a 

result, wetlands only occur where groundwater systems approach the surface, 

typically between dunes (Drda 1998; Novacek 1989). In some areas, these interdunal lakes and wetlands can be quite abundant, often supporting a high density 

of marshes. Studies have demonstrated that these eolian features alternate between 

periods of relative stability and increased dune activity resulting from prolonged 

drought conditions (Clarke and Rendell 2003; Forman and Maat 1990; Muhs 

and Holliday 1995).

Although generally discussed in the context of riparian areas, beaver are an additional factor that can affect marsh development. The construction of ponds can 

lead to the formation of marsh communities; depending on the stability of dams, 

these can persist on landscapes for decades (Naiman and others 1988). Although 

exhibiting different functional attributes than marshes formed in depressions, 
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the reduction in stream energy and impoundment of water can create conditions 

suitable for the development of emergent species. The lifespan of dams and associated ponds can vary widely, from less than one year to decades or longer 

depending on characteristics such as mean and peak discharge, stream gradient, 

and stream power (Collen and Gibson 2001; Gurnell 1998), which affects vegetation response and subsequent successional development.

Hydrologic, ecological, and disturbance processes

In both pristine and anthropogenically altered or created systems, the hydrologic 

regime is a primary factor that influences wetland structure and function, affecting 

everything from the physiological performance of individual plants (Wilson and 

Keddy 1985) to rates of productivity, nutrient cycling, and species composition 

(Neill 1993, 1995; van der Valk 1994; van der Valk and others 1994). A key attribute that distinguishes marshes is the presence of standing water, at least on 

an intermittent or seasonal basis, although fens may also support shallow surface water features (e.g., flarks or pools). Many marshes also naturally exhibit 

relatively high intra- and inter-annual variability in the depth of inundation and 

water table depths (Winter and others 2001). Characterizing the amplitude and 

temporal frequency of such fluctuations is, therefore, essential to understanding 

marsh function and response to anthropogenic activities.

The seasonality, depth, and duration of inundation or saturation (i.e., hydroperiod) 

all influence marsh function and vegetation dynamics (Weller 1994) and vary 

widely among different wetland types and within an individual class of wetlands 

such as marshes (Winter and Woo 1990). Variability in these key hydrologic 

attributes is important in structuring species composition, both spatially, within 

marsh complexes, and temporally, from wet to dry years (Seabloom and others 

1998; Smith and Haukos 2002; van der Valk and others 1994). Consequently, 

most wetland classification schemes incorporate some degree of hydrological 

characterization (Brinson 1993a; Cowardin and others 1979; Zoltai and Vitt 

1995). In contrast to fens, where mean water depth is a significant predictor of 

vegetation patterns, measures of hydrologic variability are often better predictors 

of vegetation patterns in marshes (Asada 2002).

As with wetlands in general, marsh formation is ultimately dependent upon the 

presence of sufficiently wet conditions to drive the development of hydric soils 

and vegetation. A wetland’s water balance, i.e., the difference between water 

inflows and outflows and storage, drives the local hydrology, which helps shape 

vegetation (Mitsch and Gosselink 2007; Winter and Woo 1990). Hydrologic inputs 

include direct precipitation, surface water inflows, and groundwater discharge. 

Outflows can occur as surface water, direct evaporation, plant transpiration, or 

percolation into groundwater aquifers. The relative importance of each of these 

processes varies widely among marshes as functions of climate, lithology, and 

physiographic setting and is largely responsible for the different hydrological 

regimes occurring in Region 2 marshes.
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The source of water to wetlands—precipitation, overland flow, and/or groundwater—varies among marshes in the region. Although direct precipitation inputs 

are certainly important components of marsh water budgets, unlike specific wetland types such as ombrotrophic bogs or vernal pools, many marshes also rely 

upon significant surface or groundwater inflows (Mitsch and Gosselink 2007; 

Weller 1994). Exceptions are playas and similar marsh types, which receive 

water almost exclusively as surface runoff or direct precipitation inputs (Bolen 

and others 1989; Smith 2003).

Research in a variety of settings has demonstrated that many wetlands, including marshes, are integral parts of groundwater flow systems (van der Valk 2005; 

Winter and others 2001). In regional groundwater flow systems, recharge typically 

occurs at high topographic areas, while discharge occurs at low areas. However, 

local flow systems associated with wetlands are often superimposed on these 

regional systems, resulting in complex interactions between groundwater and 

surface water, regardless of regional topographic position (Winter and Woo 1990). 

Consequently, generalizations are difficult to make regarding hydrologic function.

Considering the importance of hydrologic regime to patterns of vegetation composition and dynamics, assessments of the current or past conditions need to be 

made in light of hydrology. Unfortunately, hydrologic data are rare for marshes, 

although a handful of long-term studies are available (Labaugh and others 1996; 

Winter and Rosenberry 1995). For historical reconstructions, the paucity of 

hydrologic data necessitates the use of proxy information such as paleoclimatic 

reconstructions and knowledge gleaned from contemporary studies of human 

impacts to marshes.

Although hydrology is the single most important factor influencing marsh 

vegetation dynamics, other disturbance agents can also be important. Fire is a 

nearly ubiquitous disturbance type in North American terrestrial ecosystems; and 

over sufficiently long time scales, nearly all terrestrial ecosystem types in the 

region, except alpine tundra, experience it (Pyne and others 1996). Although fire 

is regarded as a key disturbance type influencing vegetation patterns in upland 

ecosystems in Region 2 Forests and Grasslands (Dillon and others 2005; Goldblum and Veblen 1992; Romme 1982; Sherriff and others 2001), the historical 

importance of fire in structuring wetlands, and marsh ecosystems in particular, 

is less evident.

The effect of fire on marshes has received some study, but mostly from 

outside of the region. For example, many marshes in the Gulf Coast region 

are intensively managed through a combination of fall/winter burning and 

construction of impoundments to improve wintering waterfowl habitat and 

create emergent wetlands (Gabrey and others 1999). The use of fire in marsh 

management is also practiced in the Great Plains. For example, fire has been 

used as a tool to control cattails with the objective of improving habitat for 

migratory birds (Higgins and others 1989; Kostecke and others 2004). 
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Studies document the use of fire in marshes to open up dense stands of vegetation, to control plant succession, and to promote production and palatability 

of the aquatic plants providing forage for waterfowl (Ford and Grace 1998).

Fire can affect marshes in variety of ways. One obvious effect is through direct 

plant mortality, although the susceptibility of different species varies. In general, fire reduces aboveground biomass in marshes; for example, in a study of 

Louisiana costal marshes, mean total biomass was over 1.5 times greater in the 

plots that remained unburned than in those that were burned (Taylor and others 

1994). Different marsh vegetation types may differ in their response to fire. For 

example, Smith and Kadlec (1985) found in a study of Utah marshes that vegetation regrowth was rapid following fire in communities dominated by Typha 

spp., Scirpus lacustris, and S. maritimus, but was very limited in more saline 

sites dominated by Distichlis spicata (Smith and Kadlec 1985). Fire effects may 

differ depending on the seasonality of burning. For example, Mallik and Wein 

(1986) found that fire reduced Typha glauca cover, stem density, and plant height 

regardless of the seasonality of burning, but reductions were significantly greater 

following summer burns (Mallik and Wein 1986).

In addition to direct effects such as plant mortality, fire can indirectly affect 

vegetation through a variety of mechanisms. For example, in marsh communities 

along Great Salt Lake in Utah, fire was observed to promote increased herbivore 

use of the habitat, presumably due to increases in protein content of the vegetation following fire (Smith and Kadlec 1985). However, such an effect has not 

always been reported; Taylor and others (1994) found no short-term change in 

herbivory impacts due to burning in a Louisiana coastal marsh. Additional effects 

are changes in water and soil chemistry such as short-term increases in nutrient 

availability (Boerner 1982).

Because marshes are geographically discreet features and comprise such a small 

portion of landscapes, the likelihood of natural ignition in marshes is extremely 

small. Therefore, characteristics of the fire regime, fire frequency in particular, 

are derived from the surrounding forest, shrub, or grassland cover. These can 

vary widely across the region, making generalizations difficult. In addition, 

complicating the assessment of historical conditions is the uncertain and variable 

role of Native Americans in shaping fire regimes (Vale 2002). They likely had 

relatively significant influence on grassland fire regimes. For example, Higgins 

(1986) concluded that fires set by Native Americans occurred primarily during 

two periods: March through May, with a peak in April, and July through early 

November, with a peak in October. Natural ignition from lightning also commonly occurred, generally during summer and early fall when thunderstorms are 

most common (Higgins and others 1989). Higgins also suggested that nearly all 

fires during the fall-winter period that have been reported in journals and letters 

from the region since about 1750 were probably not caused by lightning strikes, 

but rather were ignited by humans (Bragg 1995; Higgins 1984).
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Native American influences on forest fire regimes across Region 2 are relatively 

poorly understood and probably varied across the region. While burning by native 

peoples was likely a factor of local importance, particularly at lower elevations, 

it is likely that their impact to larger landscapes was relatively modest (Baker 

and Ehle 2001). Fire return intervals in the region’s forests vary widely among 

different cover types and specific regions and have varied temporally in response 

to both natural climatic variability and anthropogenic influences. In general, 

fire return intervals in lower-elevation cover types (e.g., ponderosa pine) would 

have been much shorter than in higher elevations cover types (e.g., subalpine 

spruce-fir forests). However, the abundance of marshes is generally very low in 

the lower forested zones of the region.

Considering the active fire regimes characterizing much of the region’s forests 

and grasslands, it is reasonable to conclude that fire affects Region 2 marshes on 

at least an occasional basis. Although not likely to affect the overall occurrence 

of marshes on the landscape, fires likely had impacts on species composition. 

For example, cattail-dominated marshes in North Dakota were apparently less 

common in the early Twentieth Century; they were typically restricted to the 

margins of large freshwater lakes (Kantrud 1992). The more restricted distribution of cattail in the past may have been due to the influence of prairie fires and 

large ungulate grazing.

A variety of animals may use marshes directly or indirectly, thereby affecting vegetation patterns. Herbivory is one of the most obvious direct impacts, 

although its importance can vary widely depending on the nature of the wetland and of the fauna in question. Historically, large mammals such as bison 

would have at least occasionally utilized marshes for water and possibly 

forage. However, there is insufficient evidence to make broad generalizations 

about their ecological effects on marshes. It is likely that effects resembled, 

in part, those observed from livestock use of wetlands and could include 

plant mortality from direct herbivory and trampling, soil compaction, and 

reduction in overall plant cover.

An additional animal of importance to marshes is the muskrat, which is 

widely distributed throughout Region 2, found along the edges of alpine 

tarns to riparian wetlands on the plains and in semi-desert valleys (CDOW 

2005). Muskrats are primarily herbivores, feeding on a range of aquatic and 

emergent species. Several studies, particularly in the prairie pothole region, 

indicate that muskrats can have significant effects on marsh vegetation structure. Muskrat “eat-outs,” impacting nearly all emergent plants in a marsh, 

are common and along with climatically induced hydrologic change, can 

drive cyclical vegetation patterns observed in many pothole wetlands (van 

der Valk and Davis 1978).
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A wide variety of birds utilize marshes, although how they affect vegetation 

patterns, as opposed to how vegetation influences habitat suitability, is largely 

unexamined. Some birds such as geese can influence marsh vegetation by direct 

herbivory (Belanger and Bedard 1994; Smith and Odum 1981; van den Wyngaert 

and others 2003). In addition, waterfowl may influence marsh ecosystems indirectly by serving as vectors for seeds (Charalambidou and Santamaria 2005) 

and by influencing nutrient dynamics (Post and others 1998).

Marsh Classification and Gradients_______________

Typically, marshes are described in terms of both vegetation and hydrology, with 

an emphasis placed on the presence of emergent herbaceous plant species and at 

least occasional inundation. Several different wetland types meet this definition, 

going by various names regionally. Specific types of non-tidal marshes include 

prairie potholes, playas, lacustrine fringe wetlands, and kettle ponds.

True prairie potholes do not occur in USFS Region 2, but because of their great 

importance to migratory birds, they are among the most studied marsh types in 

North America and deserve discussion in this assessment. The term “prairie pothole” is generally reserved for glacially-formed shallow depressional wetlands 

distributed throughout the upper Midwest and Canadian plains. Prairie potholes 

can be inundated on a temporary, seasonal, semi-permanent, or permanent basis, 

with the specific hydrologic regime being dependent on factors such as basin 

size, hydrologic connectivity to adjacent wetlands and groundwater systems, and 

climate patterns (Arndt and Richardson 1988; Galatowitsch and van der Valk 

1996; Kantrud and others 1989b).

 Generally speaking, playas are a special type of marsh found in arid or semiarid regions that have distinct wet and dry seasons. In North America, the term 

playa is typically reserved for small marshes formed in basins similar to prairie 

potholes but of different geologic origin and function. Playas are found in greatest 

abundance in the southern Great Plains, specifically the “Playa Lakes region” of 

Texas, New Mexico, Kansas, and southeast Colorado (Bolen and others 1989; 

Nelson and others 1984). There is still some controversy regarding their origin, 

with evaporative and dissolution processes, wind erosion, and bison wallowing 

all cited as possible causes of playa formation (Hovorka 1995; Rosen 1994). Most 

exhibit a distinct circular or oval shape when viewed from overhead, a factor 

driving the intense speculation among geologists regarding their origins (Smith 

2003). Playas are supported hydrologically almost exclusively by precipitation, 

with flood waters retained due to the presence of impermeable clay layers in the 

bottom of basins (Smith 2003). Although they are often thought to be restricted 

to the playa lakes region centered in Texas, similar systems can be found in 

eastern Wyoming, the rainwater basin of Nebraska, and unglaciated portions of 

the South Dakota mixed grass prairie (Smith 2003).
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Marshes also commonly occur along the margins of ponds, lakes and reservoirs. 

Such wetland communities have been characterized as “lacustrine fringe,” a 

generic term describing hydrogeomorphic setting (Brinson 1993a). Lacustrine 

fringe wetlands are broadly distributed along both natural and artificial water 

bodies throughout Region 2, occurring on the plains as well as in the mountains. 

Their presence or absence, as well as their relative width along shorelines can 

vary with basin size and shape, water management, or factors such as exposure 

to wave action. For example, communities are generally lacking along lakes or 

reservoirs with steep banks or with highly variable water levels as is often seen 

with managed reservoirs. Due to winds or currents, surface water movement is 

generally horizontal, but vertical fluctuations resulting from seasonal water level 

fluctuations can also occur, particularly in heavily managed systems. Lacustrine 

fringe wetlands always occur adjacent to open water aquatic systems, differentiating them from marshes formed in many smaller basins.

The term “kettle pond” refers to depressional ponds or wetlands found in glaciated landscapes. Generally, geologists reserve the term for features formed 

by the melting of large ice blocks lodged in deposits of glacial till or outwash, 

although the term is often generically applied to any small depression resulting 

from glacial action (Elias 1995; Zogg and Barnes 1995). Vegetation and hydrology in potholes can vary, with marsh communities often grading into either 

floating aquatic communities or peat-forming systems. Although relatively rare 

landscape features, wetlands that are formed in potholes often support marsh 

species and communities.

Marshes in Region 2 would be placed in either the palustrine or lacustrine systems 

according to the classification scheme employed by the U.S. Fish and Wildlife 

Service’s National Wetlands Inventory (NWI) program (Cowardin and others 

1979). Most marshes would be placed in the emergent (EM) class. Specific 

subclass and modifiers (e.g., salinity and water regime) can vary from wetland 

to wetland.

An additional framework developed to classify wetlands is the Hydrogeomorphic approach (HGM) developed by Brinson (1993a). From this original 

framework, a variety of regional classifications have been developed (Cooper 

1998a; Hauer and others 2002a, 2002b; Stutheit and others 2004). Cooper (1998) 

investigated the relationship among geomorphology, wetland vegetation, and 

wetland functions in Colorado and produced an HGM classification with four 

hydrogeomorphic classes: riverine, slope, depression, and mineral soil flats. 

Within these, he identified several subclasses that are comprised of wetlands 

sharing general characteristics and similar functions. Landscape position and 

water source supporting the wetland are the critical factors that distinguish the 

four classes. In the Cooper classification, all marshes fall into the Depressional 

class, with differences among marsh types driving placement of wetlands into 

one of five subclasses (Table 17).
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Carsey and others (2003) combined several of Cooper’s original subclasses as 

part of their comprehensive, statewide classification in Colorado. Wetlands placed 

in the Depressional subclass 1 (D1) are found in mid to high-elevation basins 

with peat soils along with communities along the fringes of lakes. Most of the 

basin peatland types would be considered fens, not marshes, but some of the 

lacustrine fringe wetlands on mineral substrates would be considered marshes. 

Carsey and others (2003) identified two seasonally flooded herbaceous wetland 

types in this subclass: Carex utriculata and Carex aquatilis-Carex utriculata, 

the former being most common.

Depressional wetlands in subclasses 2 and 3 (D2/D3) occur at lower elevations 

and have permanently or semi-permanently flooded hydrologic regimes. Examples include reservoir and pond margins as well as smaller basin marshes. 

Vegetation is typically dominated by species such as cattails, bulrushes, sedges, 

grasses, and rushes. Carsey and others (2003) identified 14 plant associations in 

this subclass. Wetlands in Depressional Subclasses 4 and 5 (D4/5) are found in 

low-elevation basins that are temporarily or intermittently flooded. Marsh types 

in these subclasses include playa lakes. Vegetation cover is often poorly developed, and the depression bottom may be barren (Carsey and others 2003). Also 

included are abandoned beaver ponds and small irrigation ponds. Twelve plant 

associations were identified by Carsey and others (2003) in these subclasses, all 

dominated by forbs or graminoids.

Water chemistry can vary widely among marshes, with the concentration of salts 

being a particularly important variable influencing marsh vegetation. Depending 

on the concentration of ions, water in marshes can be classified as fresh, brackish, 

or saline. “Brackish” has generally been used to refer to waters of intermediate 

salinity (Stewart and Kantrud 1972) and the term “saline” can be used to indicate 

the presence of any one of a number of cations (calcium, magnesium, sodium, 

and potassium) or anions (carbonate, sulfate, and chloride) (Cowardin and others 1979). Salinities are usually expressed in units of specific conductance or as 

percent salt (Ungar 1974a), although various descriptive modifiers have been 

developed as part of different classifications.

Table 17—Hydrogeomorphic (HGM) classes identified by Cooper (1998) as part of the statewide wetlands 

classification of Colorado (Carsey and others 2003).

HGM Subclass Description Characteristic species

Depressional 1 Mid- to high-elevation basins with peat soils and lake Carex utriculata

fringes with or without peat soils. 

Depressional 2 Permanently or semi-permanently flooded low-elevation Typha spp., Scirpus spp.

basins, including reservoir and pond margin wetlands as 

well as marshes. 

Depressional 3 Seasonally flooded low-elevation basins that are dry for Eleocharis palustris

long periods. 

Depressional 4 Temporarily flooded low-elevation basins flooded for Polygonum lapathifolium

short periods in the spring and early summer. 

Depressional 5 Intermittently flooded low-elevation basins that are Xanthium strumarium

not flooded annually or are largely barren of vegetation. 
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Marsh Vegetation in the Region __________________

When analyzed over a range of time periods, many marshes exhibit wide water 

level fluctuations around a long-term mean. Evidence suggests that such oscillatory water level fluctuations have occurred for thousands of years and are key 

drivers of changes in wetland vegetation (van der Valk 2005). Alternating wet 

and dry periods help create two kinds of vegetation change: fluctuations and 

successions. The former refers to changes in the relative abundance of species 

between the wet and dry phases of the cycles and occurs whenever the range of 

water levels during a cycle is small, as in seasonal wetlands. In contrast, successions are large shifts in species composition occurring where the range of water 

levels is large, as in semi-permanent wetlands (van der Valk 2005).

High water levels during the wet phase of successions often eliminate most 

emergent species, while low water levels during the dry phase allows for the 

re-establishment of species from seed (Galatowitsch and van der Valk 1996; 

Seabloom and Van Der Valk 2003). The role of seed banks is particularly important 

in reestablishment of plants in most marsh types and has received considerable 

attention in the literature (Baldwin and others 1996; Galatowitsch and van der 

Valk 1996; van der Valk and Davis 1976, 1978; Wilson and others 1993).

A commonly observed feature of many marshes is the presence of distinct patterns of zonation, with different communities arranged in concentric areas around 

open water or the deepest portion of the marsh. Such zonation patterns are a 

common feature in a variety of marsh types including playas, prairie potholes, 

and lacustrine fringe (Hoagland and Collins 1997; Kantrud and others 1989a; 

Spence 1982). Although there are certainly differences in composition among 

individual marshes, groups of species tend to occur in roughly the same relative 

position along flooding gradients. Zonation patterns are often quite distinct and 

can be caused by a variety of factors, but hydrology, specifically the depth and 

length of inundation, appears to be the most important factor (Seabloom and van 

der Valk 2003; van der Valk 1994). These variables can affect standing vegetation by impacting physiological performance of plants and are important factors 

shaping seed germination and seedling recruitment patterns (Seabloom and Van 

Der Valk 2003; Wilson and Keddy 1985; Wilson and others 1993). Although the 

concentric rings of vegetation in marshes appear simple, research suggests that 

the species composition of these zones is the outcome of complex interactions 

between abiotic conditions and differential rates of propagule dispersal, seed 

germination, and seedling recruitment, all of which can be affected by antecedent conditions.

For example, Seabloom and others (2001) found that the importance of historical 

recruitment events declines relative to nature of the current conditions after one 

to two years after a major hydrologic change is imposed (Seabloom and others 

2001). Welling and others (1988) examined establishment of common emergent 

marsh species along a height gradient and found that differences in environmental conditions seem to have had less impact than the distribution of seeds 
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on the distribution of seedlings. The authors suggested that these differences in 

distributions along the height gradient between seedlings and adult plants were 

due in part to post-recruitment processes (Welling and others 1988).

Region 2 supports a wide range of marsh communities, not surprising considering 

the size and diversity of the region. Several different vegetation classification 

schemes have been undertaken, differing in methods and geographic scope. For 

example, the Nature Conservancy (TNC) developed a vegetation classification 

for all of Region 2, using “ecological systems” as the primary classification unit 

(Comer and others 2003b). Ecological systems data are coarser in scale than either alliance or plant associations, which represent the finest classification units 

defined by the National Vegetation Classification system (Comer and others 

2003b). In their analysis of Ecological Systems in Region 2, Comer and others 

(2003b) described three marsh system types, collectively comprising less than 1 

percent of the region’s land area (Table 18). The authors used TNC ecoregions 

to more explicitly tie their classification geographically to Region 2, but some 

systems, notably the North American Arid West Emergent Marsh, were very 

similar across multiple Ecological Divisions (Comer and others 2003b).

Carsey and others (2003) describe 21 different marsh vegetation associations 

(Table 19). These include monotypic communities dominated by hydrophytes 

such as cattails (Typha spp.) and bulrushes (Schoenoplectus spp.) as well as 

communities dominated by facultative species such as buffalo grass (Buchloe

dactyloides) and western wheatgrass (Pascopyrum smithii). These latter communities are typical of playas, which are flooded with the lowest frequency of 

any marsh type in the region. Non-native species such as barnyard grass (Echinochloa crus-galli) and reed canary grass (Phalaris arundinacea) are dominants 

in two of the vegetation types. Indicator species for HGM subclasses were also 

reported (Carsey and others 2003).

Table 18—Approximate area of different Region 2 ecological systems relevant to marshes as reported by Comer 

and others (2003). Marsh systems are indicated in bold.

Ecological systems Hectares Percent area

Inter-Mountain Basins Playa 12,749 0.0

North American Arid West Emergent Marsh 128,301 0.1

North Central Interior Floodplain/Wooded Draw 771,987 0.7

Rocky Mountain Subalpine Mesic Meadow 884,960 0.8

Western Great Plains Closed Depression 817,203 0.7

Western Great Plains Riparian/Western Great Plains Floodplain 1,488,930 1.3

Western Great Plains Saline Depression 2828 0.0

Rocky Mountain Subalpine-Montane Riparian Woodland and Shrubland 71 0.0

Rocky Mountain Subalpine Mesic-Spruce-Fir Forest and Woodland 48,164 0.0

Rocky Mountain Lower Montane Foothill Riparian Woodland and Shrubland 848,753 0.7

Rocky Mountain Subalpine-Montane Riparian Shrubland 475,655 0.4

Region 2 total 113,543,893
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Table 19—Vegetation types identified by Carsey and others (2003) and organized by HGM subclass.

Depressional subclass 1

Scientific description Common description

Carex aquatilis – Carex utriculata herbaceous vegetation water sedge – beaked sedge herbaceous vegetation 

Carex utriculata herbaceous vegetation beaked sedge herbaceous vegetation 

Depressional subclasses 2 and 3

Bidens cernua herbaceous vegetation nodding beggartick herbaceous vegetation 

Carex nebrascensis herbaceous vegetation Nebraska sedge herbaceous vegetation 

Carex pellita (lanuginosa) herbaceous vegetation woolly sedge herbaceous vegetation 

Eleocharis palustris herbaceous vegetation marsh spikerush herbaceous vegetation 

Glyceria grandis herbaceous vegetation American mannagrass herbaceous vegetation 

Hordeum (Critesion) jubatum herbaceous vegetation foxtail barley herbaceous vegetation 

Phalaris arundinacea western herbaceous vegetation reed canary grass western herbaceous vegetation 

Schoenoplectus acutus -Schoenoplectus hardstem bulrush -softstem bulrush herbaceous vegetation

tabernaemontani herbaceous vegetation 

Schoenoplectus pungens herbaceous vegetation threesquare bulrush herbaceous vegetation 

Scirpus pallidus herbaceous vegetation cloaked bulrush herbaceous vegetation 

Typha angustifolia – Typha latifolia herbaceous vegetation cattail herbaceous vegetation 

Depressional subclasses 3 and 4

Alopecurus aequalis herbaceous vegetation shortawn foxtail herbaceous vegetation 

Echinochloa crus-galli herbaceous vegetation barnyardgrass herbaceous vegetation 

Eleocharis acicularis herbaceous vegetation needle spikerush herbaceous vegetation 

Hordeum (Critesion) jubatum herbaceous vegetation foxtail barley herbaceous vegetation 

Pascopyrum smithii –(Buchloe dactyloides) -Ambrosia western wheatgrass -(buffalo grass) -plains

linearis -Ratibida tagetes herbaceous vegetation ambrosia -coneflower herbaceous vegetation 

Polygonum arenastrum herbaceous vegetation ovalleaf knotweed herbaceous vegetation 

Polygonum lapathifolium herbaceous vegetation smartweed herbaceous vegetation 

Veronica catenata – Juncus bufonius herbaceous vegetation speedwell – toad rush herbaceous vegetation 

Xanthium strumarium herbaceous vegetation rough cockleburr herbaceous vegetation

HRV of Marshes in Region 2_____________________

As discussed in Chapter 3, there are relatively few historical or archeological 

resources available from which to make confident assessments about the abundance or condition of wetlands prior to large-scale Euro-American settlement. 

For most of Region 2, including the central and northern Great Plains and major mountain ranges of present-day Colorado and Wyoming, initial European 

exploration was by French and French-Canadian trappers seeking beaver and 

other pelts (Goetzmann and Williams 1992). Unfortunately, they generally left 

no records of their travels useful for our purposes.
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Likewise, there are few resources from which to evaluate the influence of native 

peoples on marshes of the region. Although only a handful of wetland sites have 

yielded significant archeological evidence of Native American use (Reider 1990), 

it is likely that they utilized wetlands on occasion for water. It is also likely that 

some direct utilization of marsh plants for food may have occurred. For example, 

the young shoots, young and mature rhizomes, staminate flowers, and pollen of 

cattails are all edible and may have been utilized for food. However, it is unlikely 

that subsistence-based utilization of Typha spp. or other marsh species would 

have been intense or prolonged enough to result in lasting effects to marshes.

Overview of anthropogenic impacts

Since no data are available detailing historical wetland conditions or abundance, 

our assessment is, by necessity, derived largely from an analysis of contemporary 

studies of anthropogenic impacts to marshes. Although many were not conducted 

in Region 2 proper, they offer valuable insights into how humans can alter marshes, 

either indirectly or directly. However, because of the widely varying nature of 

marshes in the region and the differing historical and ecological context of each 

National Forest or Grassland, our generalizations need to be viewed with caution.

Direct hydrologic alteration, typically draining for agricultural purposes, is likely 

the most significant anthropogenic impact to marshes in the region. Agricultural 

impacts were, by far, the greatest contributor to wetland loss according from the 

late Eighteenth to the late Twentieth Century (Dahl 1990). Indirect hydrologic 

alterations are also common, whereby the quality, quantity, or timing of surface 

and groundwater flow into marshes is changed as a result of indirect effects in 

surrounding areas. Examples include reductions or increases in surface flow 

from road networks or lowered water tables resulting from groundwater pumping 

(Richter and others 1996). An additional example is the effect of storm water and 

irrigation outflows. Because marshes are formed in basins, they are often used 

as detention ponds. Such incidental inflows from adjacent irrigated areas were 

identified as a central factor changing species composition in playas (Hoagland 

and Collins 1997).

One significant difference between contemporary and historical landscapes 

is the presence of extensive road networks. These can significantly alter local 

and watershed-scale hydrologic processes, thereby affecting marsh function. 

By affecting natural drainage patterns, roads and their associated engineering 

structures such as culverts and ditches can alter natural drainage patterns and 

reduce interception and infiltration rates by removing vegetation and causing 

soil compaction (Forman and Sperling 2002; Jones and others 2000).
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In addition to hydrologic impacts, roads can be a source of pollutants to marshes 

and can alter water chemistry parameters (e.g., cation concentrations, conductivity, 

and pH) through road dust, increased sediment deposition, and the direct application of chemicals (e.g., deicing agents used in road maintenance) (Trombulak and 

Frissell 2000; Wilcox 1986a, 1986b). Changes to marsh water chemistry can also 

occur in agricultural or suburban settings, where increased nutrient concentration 

in runoff can cause significant changes to vegetation (Aldous and others 2005; 

Drexler and Bedford 2002; van der Hoek and others 2004).

Domestic livestock can have significant effects on wetland flora through direct 

effects like herbivory and trampling and through indirect effects like nutrient 

enrichment via urine or fecal deposits (Gunnell and Smith 1972). Livestock can 

also alter soil physical properties due to hoof action and can serve as a vector for 

the spread of non-native propagules (Belsky and others 1999; Keeley and others 

2003). It is important to note that many portions of the region historically supported bison, whose general effects on vegetation, in some ways, resemble those 

of cattle (Plumb and Dodd 1993). However, considering the differences in the 

numbers and distribution of livestock versus historical bison populations, such 

comparisons may not be valid. In general, marshes appear to be more resilient 

relative to arid uplands to livestock grazing (Fleischner 1994), although there is 

scarce evidence one way or another.

Changes in the fire regimes of many of the region’s upland ecosystems relative to 

historical conditions are well documented (Brown and Sieg 1999; Donnegan and 

others 2001; Meyer and others 2003; Tinker and others 2003). As a consequence, 

changes in the frequency of fire in marshes have likely resulted. In addition to the 

direct effects of fire, fire in adjacent uplands may result in increased water and 

sediment yield and changes in water chemistry, although the magnitude of these 

changes relative to pre-fire conditions should decrease over time as the density 

and cover of upland vegetation increases (Troendle and King 1985). In some 

areas, native peoples may have influenced marsh vegetation through burning. 

For example, Davis and others (2002) concluded that prior to the historic period, 

burning was frequent enough to exclude woody plants (Celtis, Cephalanthus, 

Populus, Fraxinus, and Salix spp.) from and suppress the abundance of Scirpus

spp. (Davis and others 2002).

Changes in the abundance and distribution of marshes

Marshes are found throughout North America, covering an estimated 10,000,000 

ha (24,710,500 acres) in the coterminous United States (Mitsch and Gosselink 

2007). Reliable statistics pertaining to marshes in Region 2 are lacking, but in 

their analysis of GAP vegetation data, Comer and others (2003) present a value 

of 960,000 ha (2,372,211 acres) for the three main Ecological System types 

they identify as characteristic of marshes. Although much of the region has been 

mapped as part of the NWI project, although maps for many areas have yet to 

be digitized and data are in many instance decades old, limiting the accuracy 
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and feasibility of potential analyses. In the contemporary Region 2 landscapes, 

wetland acreage is generally greater in montane areas relative to the areas of the 

Great Plains. This also applies to marshes, particularly in regard to USFS lands 

in Region 2, as most National Grasslands tend to have fewer water resources 

relative to surrounding lands that were better able to sustain homesteaders and 

thus remained in private hands.

The natural distribution of different marsh types varies widely. For example, 

playas occur exclusively on the Great Plains. They are most abundant in the 

Playa Lakes Region in the southern Great Plains, where nearly 20,000 playa 

basins have been estimated to occur (Wood 2002). The Playa Lakes Region 

is centered in Texas, but extends into southeastern Colorado and southwestern Kansas, areas that support moderately high numbers of playas (Table 20). 

Playas, or functionally similar systems, also occur elsewhere in Region 2. For 

example, nearly 14,000 ha of playas are thought to exist in the Rainwater Basin 

in Nebraska (Smith 2003), and approximately 450 playas have been documented 

in Wyoming’s Powder River Basin (Brough 1996). Playa numbers elsewhere in 

Wyoming or in northeastern Colorado are unknown (Smith 2003). Other marsh 

types are more widely distributed. For example, lacustrine fringe communities 

can occur along lake and reservoir margins from the plains to the alpine, although 

plant community composition changes over the elevation gradient.

Pre-settlement distribution of marshes would have been governed largely by 

the previously discussed drivers of climate, hydrology, and geology. The latter 

can be regarded as largely invariable over the time period covered by this assessment (approximately 1600 to 1850 for pre-settlement and 1850 to present 

for post-settlement), although climate has fluctuated widely in this time period, 

undoubtedly affecting marsh hydrology.

Table 20—Number and area of playa basins in Colorado 

and Kansas (Guthry and others 1981, as presented in Smith 2003).

State/county Number of playas Area (ha)

Colorado

Baca 198 675

Kansas

Grant 232 752

Haskell 701 2755

Meade 712 3645

Morton 58 430

Seward 294 1734

Stanton 676 1900

Stevens 133 746

Total 2806 11,962
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Because of the importance of marshes in the Great Plains to migratory birds, they 

have generally received proportionally more study than marshes in mountainous 

regions. In addition, many marsh types like prairie potholes are easily photointerpreted because of their distinctive morphology and the lack of forest cover 

(Tiner 1999). Several additional factors also contribute to our poorer understanding 

of the distribution of montane marshes. One issue is that the classification used 

in NWI mapping efforts doesn’t directly facilitate separation of marshes from 

other wetland types such as wet meadows and fens, all of which are generally 

classified as palustrine wetlands. However some estimates are available for more 

limited geographic areas.

The abundance of marshes on the contemporary landscape is undoubtedly lower 

than prior to Euro-American settlement of the region, although precise estimates 

of wetland loss are not available. Studies conducted at broad scales suggest 

that impacts have been significant. The total wetland area in the conterminous 

United States in the 1780s has been estimated at 221 million acres (Dahl 1990); 

by the 1980s, only 103 million acres were thought to remain (Dahl and Johnson 

1991). Over the past 200 years, 22 states have lost more than 50 percent of their 

wetland habitat; estimates of wetland losses in Region 2 states are generally 

lower but are still significant. However, it is unknown what proportion of lost 

wetlands were marshes.

Although the overall trend in wetland area nationally and in Region 2 is certainly 

negative, causes and patterns of wetland loss vary geographically and among 

wetland types. Wetlands in the more arable portions of the Great Plains are 

among the most heavily impacted, largely due to agricultural draining (Brinson 

and Malvarez 2002; Mccauley and Jenkins 2005). Water control associated with 

urbanization has also had significant impacts, though effects are concentrated 

where current or past population growth and economic development have been 

greatest.

Complicating our assessment of historical changes in marsh abundance and distribution is the widespread occurrence of man-made water bodies and wetlands 

across the contemporary landscapes. These wetlands have often been purposefully created for reasons such as regulatory compliance (e.g., mitigation activities 

required under Section 404 of the Clean Water Act) or wildlife habitat improvement, although many have been created incidentally as part of water storage 

projects for agricultural and municipal use, flood control, or power generation 

use. Examples range from the large reservoirs built on the region’s big rivers to 

the numerous livestock watering impoundments scattered throughout the high 

plains. Many of these were constructed in the 1930s and 1940s with the aid and 

encouragement of Government cost-sharing programs. Although not built or 

managed for ecological functions, many of these features support emergent vegetation and provide habitat value for wildlife (Rumble 1989; Uresk and Severson 

1988). Other sources of man-made water bodies on the contemporary landscape 
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include coal, bentonite, and gravel surface mine impoundments and water treatment ponds (Barrett 1999; Mckinstry and Anderson 2002; Rumble and others 

1985). Some of these impoundments do support small marsh communities along 

their margins, although often their steep banks limit the occurrence of emergent 

vegetation communities (Rumble and others 1985).

As part of its responsibilities under the Dam Safety Acts of 1972 and 1986, the 

U.S. Army Corps of Engineers produced the National Inventory of Dams (NID), 

which lists all artificial barriers that impound or divert water and that are 25 ft 

or more in height or have a maximum water storage capacity of fifty acre-feet or 

more (National Atlas 2002). Graf (1999) analyzed the NID and concluded that 

there are approximately 75,000 artificial dams across the United States which 

impound an amount of water approximately equivalent to one year’s runoff 

from the continent (Graf 1999). In our analyses of the NID data from Region 2, 

we found a total of 882 features, the majority of which occurred in Colorado. 

The primary purpose listed for most of these dams was either irrigation or flood 

control, with a more limited number of features classified for fish and wildlife, 

hydroelectric, or other uses (National Atlas 2002).

Unfortunately, neither the NID data nor other widely available digital data sets 

such as the hydrography layer available from the National Atlas or the Census 

Bureau’s TIGER data, provides comprehensive information on small water bodies. While the NID data provide a fairly complete inventory of large features, 

they miss virtually all features smaller than 105 m2, which includes most small 

water bodies such as livestock watering ponds, thus grossly underestimating the 

total area of created ponds and wetlands in the region (Smith and others 2002). 

Although their distribution is poorly understood, collectively these features have 

great importance, resulting in regionally elevated evaporation rates, decreased 

downstream flow, and altered groundwater recharge and sediment transport rates 

In addition, they represent novel aquatic and wetland habitats that may at least 

partially compensate for—but spatially redistribute—wetlands lost to anthropogenic activities (Smith and others 2002).

Livestock watering ponds can be classified as retention reservoirs, pit reservoirs 

(i.e., dugouts), and pit retention reservoirs (Lokomoen 1973). Retention reservoirs 

involve the construction of small dams across intermittent streams or gullies 

to intercept spring runoff or rainwater from upland slopes, while pit reservoirs 

are steep-sided and are generally supported from groundwater as well as surface runoff (Olsen 1999). The design of pit retention reservoirs is similar to pit 

reservoirs except that spoil material is placed on the downstream side as a dam 

in order to flood the shallow area around the dugout (Payne 1992). The differences in design have implications for marsh development, as emergent species 

are unable to colonize if banks are too steep, as is often the case with dugouts. 
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Morphological characteristics such as pond shape and size also influence the 

suitability of ponds for waterfowl use (Evans and Kerbs 1977; Lokomoen 1973; 

Rumble 1985). For example Lokemoen (1973) suggested that the minimum pond 

size be 0.6 ha, while Uresk and Severson (1998) found that ponds with gently 

sloping shorelines and abundant emergent vegetation support greater densities 

of birds (Uresk and Severson 1988).

Although contributing a small fraction of the region’s total marsh area, marsh communities have also been documented as occurring along regulated sections of the 

Colorado River. These systems are relatively novel features, rare in pre-settlement 

landscapes because of the effects of scouring floods (Stevens and others 1995). 

Following the completion of the Glen Canyon dam in 1963, reduction in flood 

frequency permitted widespread marsh development. These features have low 

stability in the face of large spring floods but are able to quickly redevelop after 

scouring by what are now rare high flows (Stevens and others 1995). Although 

not documented in the literature, similar processes may also allow for the marsh 

development along other larger regulated rivers in the region.

Marsh structure and function

Previously prepared HRV assessments in the region have included analyses of 

structural and compositional attributes of vegetation. The importance of these 

variables for forested cover types is clear, as large shifts in forest composition or 

stand structure can have important repercussions for processes such as fire regimes 

or wildlife habitat (Meyer and others 2003; Veblen and Donnegan 2005). In addition, there are many examples from the scientific literature of methodological 

approaches for characterizing such changes and understanding their implications 

(Morgan and others 1994; Nonaka and Spies 2005; Veblen 2003). Although 

vegetation composition and structure are key attributes for marsh ecosystems as 

well, several factors limit the analytical approaches available for characterizing 

their dynamics at the broad spatial and temporal scales of this HRV assessment.

Studies have documented increases in the amount and density of vegetation cover 

in many Great Plains marshes. For example, Root and Ryan (2004) examined 

the changes in habitat for piping plovers at two North Dakota alkaline wetland 

complexes from 1938 to 1997, observing declines in beach habitat and in increase 

in vegetation cover, suggesting that long-term changes in factors such as groundwater hydrology, livestock grazing intensity, or fire frequency may be negatively 

affecting beach availability, impairing long-term recovery of threatened piping 

plovers (Root and Ryan 2004). Changes in marsh structure resulting from the 

development of dense, closed strands of species like cattails are widely noted 

in the literature and are a principal concern of many wetland managers trying 

to improve waterfowl habitat (Kantrud 1990, 1992; Weller 1994). Changes in 

the amplitude of water level fluctuations can result in shifts in species diversity 

(Keddy 2000) (Figure 22).
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Figure 22—Reduction of community diversity associated with the stabilization of water levels. The four zones 

depicted in panel A have been reduced to only two in panel B (Figure modified from Keddy 2000).

Changes in vegetation relative to historical (i.e., pre-1850) conditions vary among 

marsh types and the mechanisms of anthropogenic influence. For example, 

Hoagland and Collins (1997) attributed differences between the species composition of playas examined in their study to that presented in earlier studies to 

differences in surrounding land use and the influence of irrigation inflows from 

surrounding areas on hydrologic regime. Elsewhere in the region, changes in 

the frequency of fire and grazing practices have been used to explain shifts in 

vegetation composition or structure (Higgins 1986; Kantrud 1990).

The introduction and spread of non-native species is widely regarded as one of 

the greatest anthropogenic impacts to native communities (Brinson and Malvarez 

2002; Mack and D’Antonio 1998; Mack and others 2000). Although generalizations are difficult to make, there are several species that are regarded as problems in wetlands such as marshes. Phalaris arundinacea (reed canary grass) is 

widespread throughout Region 2, occurring in variety of habitats in addition to 

marshes. Although there is lack of consensus as to its status as native or introduced, exotic cultivars of the species were once widely planted, a factor likely 

related to its observed aggressiveness (Merigliano and Lesica 1998). It appears 

that reed canary grass responds positively to hydrologic alteration, likely an additional factor in its spread (Galatowitsch and others 1999). Its relatively great 

height per unit of biomass and its adaptable morphology have both been cited as 

possible reasons for its high competitive ability in wetlands (Miller and Zedler 

2003). It often forms monotypic stands—crowding out other species, especially 

smaller graminoids and forbs (Barnes 1999)—and has been shown to reduce 

species diversity, particularly in disturbed wetlands (Kercher and others 2004).
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Cattails (Typha spp.) are often species of additional concern, despite that they 

are one of the principal marsh dominants globally, occurring in both coastal 

and freshwater marshes. Often, they are widely viewed as aggressive invaders 

by wetland and wildlife managers because of their propensity to form uniform 

stands. The genus includes Typha domingensis, T. angustifolia, and T. latifolia, 

all of which have been documented in the region (Smith 2005). Typha latifolia

and T. domingensis are regarded as native, the latter confined to the southern part 

of the region, but evidence suggests that T. angustifolia is introduced (Kantrud 

1992). Typha angustifolia and T. latifolia readily hybridize, and their hybrid, 

Typha x glauca, is known to be an aggressive competitor in marshes. Cattails, 

particularly T. angustifolia and its hybrids, tend to be most aggressive and problematic in disturbed wetlands and those receiving elevated levels of nutrients such 

as marshes found in agricultural or urban watersheds (Woo and Zedler 2002).

The distinctive flowering spike in Typha spp. can produce hundreds of thousands 

seeds, which are efficiently wind-dispersed and germinate on bare wet soils or 

under very shallow water (Lombardi and others 1997). These characteristics, as 

well as ability of the plants to rapidly form clones by means of rhizomes, allows 

cattails to form large, persistent, and often monospecific stands (Smith 2005). 

These stands are a problem in many prairie wetlands because they alter habitat 

structure and function, resulting in a decrease in use by wildlife species (Kantrud 

1992; Kostecke and others 2004).

Purple loosestrife (Lythrum salicaria) is another species credited with dramatically 

changing the vegetation of many North American wetlands (Galatowitsch and 

others 1999). The species appears dependent on disturbance for seed germination 

(Rachich and Reader 1999). Following invasion, purple loosestrife often forms 

dense nearly monospecific stands, often displacing native species and reducing community diversity (Thompson and others 1987). The success of purple 

loosestrife may result from its ability to tolerate a wide range of environmental 

factors, to spread vegetatively, and to tolerate increased disturbance (Gaudet 

and Keddy 1995; Thompson and others 1987), although which factors are most 

important is unknown (Fernberg 1997).

Disturbance regimes

While there has been no research specifically examining change in disturbance 

process such as fire in Region 2 marshes, changes in fire regimes have been 

extensively studied in the region’s upland ecosystems. Since fires rarely originate in marshes, but rather spread from adjacent uplands, research documenting 

changes in the fire regime of the surrounding landscape are relevant. Studies have 

documented shifts in the frequency and magnitude of fires in a variety of cover 

types since Euro-American settlement of the region, in some instances pushing 

fire regimes out of the HRV. For example, studies in the Colorado Front Range 

have found that fire frequency increased following initial Euro-American settlement, but declined below historical levels beginning in the early to mid-Twentieth 

Century as fire suppression became standard policy and improved suppression 

techniques were developed (Veblen and Lorenz 199; Veblen and others 2000).
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In the Great Plains, changes in fire regimes appear to be caused by several factors. Changes in the seasonality and number of ignitions are largely the result of 

human activities, although lightning fires are still common during the late summer 

months (Higgins 1984). Regardless of ignition source, there have been changes 

in the relative size and seasonality of fires compared to historical conditions, 

a function, in part, of cultural obstacles and practices such as roads, cultivated 

fields, and heavy grazing (Higgins 1984). Wright and Bailey (1982) estimated 

fire frequency of 5 to 10 years on level-to-rolling topography and 20 to 30 years 

in areas with more dissected topography. Fire reconstructions near Devils Tower 

suggest that prior to 1770, the mean interval between fires was 27 years, while 

from 1770 to 1900, the fire return interval declined to 14 years (Fisher and 

others 1987). In the Thunder Basin National Grasslands, the Weibull Median 

Probability Interval for the entire period of record was 7.4 years; 7.9 years for 

the non-suppression period (1565 to 1939); and 6.7 years for the suppression 

period (1940 to 1988) (Perryman and Laycock 2000). Based on the position of 

scars within annual growth rings, Perryman and Laycock (2000) also suggest that 

most fires (80 percent) occurred during the latter stages of the growing season 

or during the dormant period.

Management Opportunities and Constraints _______

Management opportunities

A variety of approaches can be taken to better manage marsh resources. One 

of the most basic is to improve understanding of existing resources, including 

marsh abundance, distribution, and condition—the latter being best interpreted 

in relation to known marsh functions. Such information, in conjunction with 

proactive planning, can be used to avoid impacts to marshes resulting from 

management actions and, if impossible, to mitigate impacts. In addition, there 

are innumerable opportunities for marsh restoration and enhancement throughout 

the region. Many existing wetlands exhibit poor function, which with relatively 

modest investments of time and money, can be markedly improved. Likewise, 

the art and science of wetland creation has improved in recent decades and a 

wide variety of resources are available to facilitate wetland creation (Campbell 

and Ogden 1999; Colorado Natural Areas Program 1998; Galatowitsch and van 

der Valk 1998; Kusler and Kentula 1990; Manci and Schneller-McDonald 1989; 

Wheeler 1995).

Although extensive wetlands mapping efforts have been conducted within the 

region, typically involving some form of remote sensing imagery such as aerial 

photographs, their utility is often limited. For example, although widely available, NWI maps are often quite dated and can poorly reflect existing conditions, 

particularly in areas with rapid land use change. Also, digital versions of NWI 

data are unavailable for many parts of Region 2, reducing their utility to planners. 

Although existing paper maps can be digitized, this can be a time-consuming 

and expensive process.
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Wetland resources on several National Forests, for example the Bighorn and 

Pike/San Isabel, have been mapped. These data can be invaluable for conducting 

broad-scale resource assessments and planning activities such as the Aquatic, 

Wetland, and Riparian Assessment efforts completed or underway in several 

Region 2 National Forests (Winters and others 2005). Data can also be used 

to identify areas for site-specific assessments of wetland condition and threats.

Restoration opportunities

Wetland mitigation, whereby wetland losses are compensated by wetland restoration, creation, or enhancement, is a central component to wetlands policy 

in the United States (Ambrose 1999; Mitsch and Gosselink 2007). Regulatory 

requirements for mitigation of impacts to exiting wetland resources have been a 

major driving force behind much of the research into marsh creation and enhancement. However, recent studies evaluating the efficacy of mitigation policies on 

achieving the goal of “no net loss” of wetlands are less than encouraging. On the 

basis of area alone, it appears that mitigation requirements have not resulted in 

equal replacement of lost wetlands (Sifneos and others 1992). When functional 

attributes are evaluated, most mitigation wetlands fare even worse relative to their 

natural counterparts (Ambrose 1999). These studies highlight the difficulty in 

effectively creating wetlands with similar functional attributes as natural systems 

and the desirability of avoiding impacts whenever possible.

Management constraints

Changes in regulatory mechanisms protecting marshes—Wetland regulations have a long history of controversy and many specific points of practice 

have been decided by courts rather than by legislators or regulators. The main 

mechanism governing impacts to wetlands has been Section 404 of the Clean 

Water Act, which has historically placed regulatory oversight on a range of activities impacting wetlands with the U.S. Army Corps of Engineers. However, 

recent Supreme Court decisions have effectively removed the Corps of Engineers’ 

regulatory oversight for most wetlands lacking clear connections to surface 

water bodies such as streams. Many marshes lack surface water connections to 

navigable waters of the United States and may therefore be considered isolated 

under U.S. Corps of Engineers jurisdiction through the Clean Water Act (Tiner 

2003; Tiner and others 2002). It is also important to note that silvicultural activities are specifically exempted from Clean Water Act Section 404 regulations.

Climate change—Unlike issues of local or regional scope, there is little that 

managers can do to prevent climate changes widely predicted to occur in coming decades and centuries. Although there are still considerable differences in 

opinion on the topic, the general consensus among climate researchers is that 

global temperatures have risen and are likely to continue to rise. Because of their 

strong dependence on watershed-scale hydrologic processes, wetlands may be 

especially sensitive to major shifts in temperature or precipitation.
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Specific impacts are difficult to predict and are dependent on the magnitude and 

direction of change in key climate parameters. Parameters of particular importance to marsh ecosystems are precipitation, including both total amounts and 

seasonality, and temperature, which affects a range of processes from the scale of 

individual plants to entire ecosystems. Some regional climate change scenarios 

call for both increased regional temperatures and precipitation, although whether 

such changes will actually occur is unknown (U.S. EPA 1998a, 1998b, 1998c; 

Wagner 2003).

Because of the underlying uncertainty regarding possible changes, it is nearly 

impossible to develop rigid mitigation plans for marshes. Managers can focus 

on identifying key marsh resources and instituting baseline data collection and 

monitoring efforts. With a relatively modest expenditure of time and money, it 

is possible to collect enough hydrologic and vegetation data to evaluate changes 

over time. Availability of such baseline data can greatly facilitate the development of adaptive management plans.
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6. Historical Range of Variation for 

Region 2 Fens

Definitions and Concepts of Fens ________________

Fens are peat-accumulating wetlands (peatlands) that receive some of their 

water supply from surrounding mineral soils (Mitsch and Gosselink 2007). 

The accumulation of peat, i.e., incompletely decomposed organic matter, is the 

primary diagnostic characteristic differentiating fens from other wetland types 

in Region 2, although there are many additional hydrologic and floristic differences. The accumulation of peat is a direct consequence of stable, elevated 

water tables, which slow decomposition rates and allow for net accumulation 

of organic matter (Belyea and Malmer 2004; Crum 1988; Glaser 1987; Gorham 

1957). In Europe, peatlands are generally referred to as mires with a variety of 

terminology used to subdivide ecosystems based on water chemistry, floristics, 

water source, or landform setting (Heathwaite and others 1993; Wheeler and 

Proctor 2000). Both peatland and mire are general terms, referring to the full 

range of peat-accumulating ecosystems; the term fen encompasses a subset of 

these wetlands.

Fens are at least partially supported hydrologically by influxes of groundwater 

enriched in mineral ions through contact with surrounding bedrock or mineral 

soils. This differentiates fens from bogs, which are peatland ecosystems that 

receive their water and nutrients exclusively from atmospheric deposition (Crum 

1988; Gorham 1957). For ombrotrophic peatlands (bogs) to form, precipitation 

must exceed evapotranspiration and create a positive water balance (Bragazza and 

others 2005; Siegel 1988). Because evapotranspiration exceeds precipitation in 

much of Region 2, no ombrotrophic peatlands occur (Windell and others 1986).

Soils in peatlands, including fens, are predominantly organic in composition 

and are classified as Histosols if organic horizons exceed 40 cm in depth within 

the top 80 cm (U.S. Department of Agriculture 1999). Peat soils have distinctly 

different physical and chemical characteristics than mineral soils, including 

low bulk density and high water holding capacity, which influence ecosystem 

functioning (Moorhead and others 2000; Mulqueen 1986). The specific composition and characteristics of peat soils can vary widely as a function of the 

botanical composition of peat, degree of decomposition, and mineral content 

can vary within a stratigraphic sequence and among different peatlands (Glaser 

1987). While soils in fens are predominantly organic, areas of mineral soil can 

occur. In addition, in calcareous fens, accumulated tufa or marl deposits can 

form (Johnson and Steingraeber 2003). Because peat accumulation rates in the 

region are slow, on the order of approximately 20 cm/1000 years in many 

areas (Chimner and others 2002), fens cannot persist in sites with high rates of 

mineral sediment influx from adjacent slopes.
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Fens are common at high latitudes, where cool temperatures, low evapotranspiration rates, and high precipitation create conditions amenable to peat formation. 

In North America, fens are widely distributed throughout Canada, Alaska, and 

the Northern Great Lakes states, as well as areas along the Atlantic coast (Mitsch 

and Gosselink 2007). However, within Region 2, sites with the necessary surplus water balance needed to maintain high water tables throughout the growing 

season and allow for peat development are generally restricted to wet and cool 

microsites in the mountains (Chimner and Cooper 2003; Cooper 1998a). Exceptions do occur, for example, fens occur in the Nebraska Sandhills (Steinauer and 

others 1996), although these represent a small portion of all fens in the region.

Fen geomorphic setting and geochemical characteristics vary across the region 

(Figure 23). These variables are important drivers of fen vegetation and are 

central to many fen classification schemes. An understanding of how these variables influence the development, structure, and function of fens is essential for 

conceptualizing the historical range of variation and possible departures from 

the HRV due to anthropogenic impacts.

Geomorphic Setting and Principal 

Ecological Drivers _____________________________

Geomorphic setting

Within Region 2, fens have formed in a limited range of geomorphic and landscape 

settings that possess the necessary hydrologic and microclimatic conditions for 

peat accumulation (Windell and others 1986). Fens form in two general physiographic settings: slopes and basins (Figure 24). Several landform configurations 

produce groundwater discharge systems capable of supporting fens in montane 

landscapes. These can include discrete hillslope springs, upwelling springs, 

closed basins, and open-basin hillslopes.

Fens may occur at discrete springs formed where preferential groundwater flow 

paths discharge to the surface, for example, at the contact between two rock strata 

or at slope discontinuities. If aquifers supporting springs are sufficiently large 

and can maintain high water tables, peat can accumulate. In many broad valleys, 

springs may occur at or near the toe of hillslopes. Here, multiple groundwater 

flow paths may coalesce and emerge at the ground surface. Spring mounds may 

be formed if vegetation completely overgrows and contains an upwelling spring, 

with the vertical hydraulic pressure of the spring driving vertical accumulation 

of peat (Figure 25).
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Figure 23—Examples of fens in the five-state area encompassing Region 2: (A) Jumbo Fen, Nebraska; (B) unnamed kettle pond fen, Routt National Forest, Colorado; (C) unnamed kettle pond, Bighorn National Forest, 

Wyoming; (D) unnamed rich fen, Yellowstone National Park; (E) Mount Emmons iron fen, Colorado; (F) Black 

Fox iron fen, Black Hills National Forest, South Dakota.
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Figure 24—Wyoming fens formed in slope (A) and basin (B) geomorphic settings (photo: D. Cooper).

Figure 25—Schematic diagram (A) illustrating a spring mound fen and an example (B) found in Yellowstone National 

Park (B) (photo: D. Cooper).

Fens also occur in depressions or basins, commonly associated with ponds or 

lakes (Figure 24b). These features are most abundant in glaciated landscapes. 

Basin fens may support floating peat mats that are capable of rising and falling as 

water levels change, allowing the peat surface to maintain contact with the water 

surface. Basin size and depth are important variables affecting the development 

and functional characteristics of basin fens (Cooper and Arp 2002).
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Fens formed in depressions may support larger, more developed peat bodies than 

those formed in slope settings. Frequently, they develop complex microtopography (e.g., strings and flarks) and water flow paths, such as flow-around channels 

as the basin fills and a gently sloping fen develops (Cooper and Arp 2002). In 

contrast, more steeply sloping fens have smaller, poorly developed peat bodies 

and may have a greater extent of mineral soil. The rate of water inflow and outflow (flux) is typically greater for fens formed on slopes than on depressions.

Key drivers and ecological processes

Several ecological drivers operating hierarchically affect fen formation and function across Region 2, with the relative importance of each varying depending 

on the spatial and temporal scale being considered. At the broadest scales, the 

regional flora, climate, and geology are most important (Bedford 1996; Major 

1951); at intermediate and fine spatial scales, chemical and hydrological gradients 

become key (Cooper and Andrus 1994; Hajkova and others 2004; Tahvanainen 

and others 2002). Within individual fens, biotic interactions, microtopography, 

and disturbance processes shape patterns of vegetation and succession (Kennedy 

and others 2003; Kotowski and van Diggelen 2004; Mulligan and Gignac 2002).

Climate—Climatic variables such as annual precipitation and maximum temperature vary along latitudinal gradients in the region, affecting patterns of fen 

distribution and abundance. Other characteristics such as the seasonality of 

precipitation also vary, in part because of the differing influence of regional climate patterns like the El Nino-Southern Oscillation (ENSO) and Pacific Decadal 

Oscillation (Hauer and others 1997; Hidalgo and Dracup 2003; Schoennagel and 

others 2005; Sheppard and others 2002).

Because fen water tables are generally high in the spring and early summer due to 

snowmelt runoff, precipitation patterns in the late summer, following snowpack 

melt-out, are important for peat accumulation patterns. For example, in the San 

Juan Mountains of Colorado, the maintenance of high water tables in fens during 

the late summer depends on the southwestern monsoon providing rain (Cooper 

and Arp 2002). In Southern Rocky Mountain fens, much of the peat accumulated 

during the Holocene may have occurred during climate periods when mid- to 

late-summer precipitation totals were consistently high (Cooper 1998a).

At any latitude, fine-scale variation in key climatic variables is also important. 

Precipitation and temperature vary as a function of elevation (Barry and Chorley 

2003; Hauer and others 1997), in part explaining the greater abundance of fens 

at high elevations. The physiography of contributing watersheds can indirectly 

influence hydrologic processes such as the depth of snow accumulation and the 

rate of melting, thereby affecting water tables in fens.
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Geology—A variety of geologic factors affect fen abundance, distribution, biota, 

and functions. Because fens receive inputs of groundwater, the mineralogy of 

watersheds can influence water chemistry, which affects biotic composition 

(Cooper 1996; Cooper and Andrus 1994; Hajek and others 2002; Lemly and 

Cooper 2011), and ecological processes such as productivity (Chapin and others 

2004; Szumigalski and Bayley 1996b; Szumigalski 1995; Thormann and Bayley 1997a) and decomposition (Szumigalski and Bayley 1996a; Thormann and 

others 2001; Turetsky and Ripley 2005; Verhoeven and Arts 1992). Watershed 

mineralogy affects groundwater chemistry by contributing minerals through 

leaching of chemicals in solution and the transport of minerals in suspension 

(Godwin and others 2002).

A particularly important factor to fens is the abundance of calcium rich rocks 

in the watershed. Areas with bedrock dominated by granitic and metamorphic 

rocks often support fens with low pH; while watersheds composed of limestone, 

dolomite, or shale produce circumneutral to high pH and higher concentrations 

of mineral ions (Tahvanainen 2004; Tahvanainen and others 2002; Windell and 

others 1986). For example, highly alkaline and mineralized water discharging 

into calcareous fens in South Park, Colorado, flows through calcareous glacial 

outwash, alluvial deposits, and deep bedrock formations (Cooper 1996; Cooper 

and Sanderson 1997; Johnson and Steingraeber 2003; Sanderson and March 1996).

Surficial geologic deposits, particularly from past glaciations, influence fen 

distribution and function (Windell and others 1986). During the Pleistocene, 

valley glaciers formed in many mountainous areas of Region 2, eroding and 

redistributing large quantities of rock and sediment from watersheds (Richmond 

1960, 1986; Thornbury 1965; Wright and Porter 1983). Many areas have experienced multiple glacial epochs, extensively reworking landscapes. The resulting landforms strongly influence contemporary patterns of fen occurrence. For 

example, fens have formed in kettles, features formed by stagnant ice blocks 

that were buried in outwash or moraines as glaciers retreated (Menzies 2002). 

Fens can also be found in drainages partially impounded by lateral or terminal 

moraines (Figure 26).

Hydrologic regime—At fine and intermediate spatial scales, hydrologic regime 

is the primary factor governing the structure and function of wetlands, including fens (Mitsch and Gosselink 2007). Relative to other wetland types such as 

marshes or riparian areas, fens have stable water supplies with water tables at or 

close to the ground surface for most of the growing season (Windell and others 

1986; Winter and others 2001). Unlike marshes, fens do not experience deep 

inundation, although some microsites such as pools and water tracks can have 

greater than 20 cm of standing water seasonally (Cooper 1990a).
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Figure 26—Two photographs, illustrating at different scales, a lateral moraine dammed basin supporting a fen in Rocky 

Mountain National Park, Colorado (photos: D. Cooper). The glacier flowed from right to left in the valley shown in the 

foreground.

Although fens often occur in stream valleys as part of larger wetland complexes, 

unlike riparian ecosystems, fens do not experience high-velocity surface flows 

or sediment deposition from fluvial processes. Hydrologic inputs to fens include 

direct precipitation, surface water, and groundwater. The relative importance 

of each differs among fens and can vary among microsites within individual 

wetlands (Koerselman 1989).

All fens receive some water via direct precipitation, with the relative amount 

scaled to fen area. However, because no ombrotrophic peatlands occur in Region 

2, precipitation is likely the least important contributor to fen hydrologic regimes. 

Much of water entering fens via rain may be lost through evapotranspiration 

(Gorham and Hofstetter 1971), although accumulated snow pack overlying some 

fens may be an important source of water to subalpine fens (Windell and others 

1986). Microsites, such as the upper portions of hummocks, may be effectively 

isolated from groundwater and be more dependent on precipitation.

Surface water inputs to fens can occur as sheet flow or channelized flow. Where 

measurements have been made, flow rates are relatively low, although their 

importance to fen hydrologic budgets may be significant. Large, channelized 

streams do not generally occur in fens because high stream energy can lead to 

peat erosion. Sheet flow, particularly from seasonal snowmelt, may contribute a 

significant amount of water to fens, although such inputs may be small relative 

to water stored in the peat body. In larger fens, sheet flow may coalesce to form 

water tracks, often around the margins of peat bodies (Cooper and Arp 2002; 

Glaser and others 1981; Heinselman 1970).

Groundwater is the most important source of water to fens. Because of the region’s 

relatively dry summers and the small size of many contributing fen watersheds, 

precipitation and surface water flows alone are insufficient to support fen hydrologic regimes. In addition to contributing to high water tables, groundwater 

discharging into fens is cold, further inhibiting decomposition and promoting 

peat accumulation.
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Groundwater in the peat bodies of fens is hydraulically connected to underlying 

mineral soils (Siegel 1983, 1988). Because of its direct contact with minerals in 

bedrock or alluvium, groundwater generally has higher mineral ion concentrations 

than either surface water or precipitation. The highly variable conductivity in 

peat soils and complex mixing of water sources creates significant physiochemical heterogeneity, an important driver of fine-scale community diversity in fens.

Groundwater entering fens can originate in local, intermediate, and regional 

groundwater aquifers, with the relative importance of each varying among wetlands (Chapman and others 2003; Godwin and others 2002; Winter and others 

2001). For example, calcareous fens in South Park, Colorado, receive groundwater 

from shallow outwash and alluvial aquifers as well as from deep bedrock aquifers 

(Johnson and Steingraeber 2003). The specific flow paths taken by groundwater 

are often complex and influenced by factors such as the geological makeup and 

physiographic characteristics of contributing watersheds. For example, coarse 

glacial deposits with high permeability can link recharge and discharge areas, 

transmitting groundwater to fens (Almendinger and Leete 1998).

Hydrologic regimes within individual fens can be variable, and differences in the 

amplitude and timing of water table fluctuations can occur among sites such as 

fen margins, spring discharge zones, or floating mats. Because floating mats rise 

and fall with changes in pond levels, plants rooted in these environments such 

as Carex limosa, C. lasiocarpa, and Menyanthes trifoliata may experience little 

seasonal variance in relative water tables even though water levels in surrounding 

portions of the fen may drop significantly (Figure 27) (Gage and Cooper 2006b).

Figure 27—Hydrographs for Buckbean Fen, San Juan National Forest, Colorado, illustrating variability of water table depths in different 

portions of the fen (Cooper: unpublished data). Well 8 is located within 

a floating mat.
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Hydrologic processes in eolian environments like the Nebraska Sandhills differ 

from those in montane areas. Because of the high infiltration rates characteristic of sandy soils, little surface runoff occurs following precipitation events 

(McCarraher 1977). Instead, water infiltrates the ground and contributes to local 

and regional aquifers. These aquifers discharge to the surface in interdunal areas 

and support lakes, ponds, and wetlands, including fens (Borgmann-Ingwersen 

1998; Steinauer and others 1996). Fens may form when the hydraulic pressure of 

the groundwater exceeds surface resistance, for example, when groundwater is 

“mounded” under a dune adjacent to a valley (Bleed 1998; Borgmann-Ingwersen 

1998). Hydrologic regimes in these settings are generally quite stable, and radiocarbon dating of peat deposits in Sandhills fens suggests that some may exceed 

12,000 years in age (Ponte 1995).

While hydrologic processes vary widely among fens as a function of factors such 

as geologic and geomorphic setting, overall hydrologic regimes are relatively 

similar across floristic and geographic gradients. As a pre-condition for peat 

accumulation, all fens, whether in the Black Hills of South Dakota or the San 

Juan Mountains of Colorado, have water tables at or near the surface for a most 

or all of the growing season. Although it is common for water tables to drop in 

the mid- to late-summer in many fens, rarely do they drop to a meter below the 

ground surface (Figure 27). During extreme or prolonged droughts, water tables 

may decline more steeply, but if such conditions persist for extended periods 

of time allowing peat soils to dry out, oxidation of peat will occur, eventually 

leading to the loss of the peat body.

Likewise, hydrologic regimes in fens with differing floristic composition and 

water chemistry often share similar hydrologic regimes. For example, hydrologic 

data from studies of iron fens and extreme rich fens, which exist at near opposite 

ends of the pH gradient, both have elevated water tables at or near the surface 

for much of the growing season (Cooper 2003; Johnson 2003).

Water and peat chemistry—Water and peat chemistry vary widely among 

fens. Several factors influence fen biogeochemistry, including biotic and abiotic 

processes (Shotyk 1988). The chemical characteristics of peatlands most commonly analyzed by ecologists include alkalinity/acidity, cation concentrations 

(e.g., Ca2+ and Mg2+), and the abundance and form of nutrients such as NH4

+ and 

NO3

-

 (Bragazza and others 1998). Variation in these chemical attributes among 

fen types include those associated with the poor to rich water chemical gradient, while those within individual fens include the mire expanse to mire margin 

gradient identified by ecologists as major controls on the vegetation of peatlands 

(Cooper 1996; Malmer 1985). Hydrochemical variability is an additional source 

of environmental heterogeneity, a factor contributing to the high species richness 

seen in many fens (Boeye and others 1994).
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As discussed for geological drivers, the mineralogy of surrounding landscapes 

is an important factor influencing fen water chemistry. In watersheds where 

relatively soluble minerals are prevalent, groundwater inputs to fens can contain 

abundant dissolved solids. In contrast, in areas composed of relatively insoluble 

minerals such as quartz, the mineral content of groundwater is typically more 

dilute (Tahvanainen 2004). The presence of minerals in groundwater also influences the pH of fens through the ability of mineral bases to neutralize organic 

acids produced by decomposing organic matter (Shotyk 1988).

Much of the spatial and temporal variability in fen water and peat chemistry can 

be attributed to differences in the relative quality and quantity of groundwater 

entering fens (Chapman and others 2003). Organic acids found in surface water 

mixes with inorganic solutes contributed by groundwater discharge, often with 

a significantly different pH (Siegel and others 2006). Complicated flow paths 

and patterns of groundwater and surface water mixing in fens can lead to high 

variation in characteristics such as pH and Ca2+ concentration. For example, two 

distinct water sources support an iron fen found near Mount Emmons, Colorado. 

Water discharging from adjacent slopes passes through pyrite-rich bedrock has 

very low pH, while water discharging from an underlying moraine has significantly higher pH (Figure 28) (Cooper 2003).

Figure 28—Schematic cross-section of the Mount Emmons iron fen in Gunnison County, 

Colorado. Groundwater discharging into the fen from bedrock of meta-sedimentary composition (right) is strongly acidic due to high concentrations of iron and sulfur. However, the fen 

has formed upon a lateral moraine composed of till containing calcareous material, from 

which it receives inputs of groundwater with higher pH.
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In addition to the external factors such as the mineralogy of contributing watersheds, autochthonous factors (i.e., those originating within fens) can also be 

important influences on water and peat chemistry, particularly at fine spatial 

scales. For example, decomposition of plant material produces strong organic 

acids, which, if unbuffered, contribute to the low pH found in some fens (Shotyk 

1988). Also important in communities or microsites dominated by Sphagnum

species are processes of active cation exchange (Clymo 1963; Gies and Lotscher 

1973; Spearing 1972). For example, the pH on the top of Sphagnum-dominated 

hummocks may be significantly lower than surrounding areas (Clymo and others 1984).

As with most terrestrial ecosystems, nitrogen is typically the limiting macronutrient in fens, although phosphorus may also be limiting in some wetlands 

(Beltman and others 1996; Mitsch and Gosselink 2007; Thormann and Bayley 

1997b). Nitrogen inputs to fens include wet and dry atmospheric deposition and 

the mineralization of organic matter (Bayley and others 2005). Nitrate reduction, 

nitrogen (N2) fixation, and denitrification are the main processes controlling the 

flux of nitrogen in fens, although the relative importance of each vary widely 

depending on site-specific hydrologic characteristics (Beltman and others 1996; 

Oien 2004). For instance, the concentration of nitrate (NO3

-

) in perennially 

saturated and anoxic sites such as floating peat mats is typically low because 

nitrifying bacteria are not present (Williams and Wheatley 1988). The importance 

of atmospheric inputs has likely increased relative to historical conditions as a 

result of increased anthropogenic inputs from industrial and agricultural activities (Baron and others 2000; Fenn and others 2003a; Fenn and others 2003b; 

Kittel and others 2002).

Although commonly used to infer nutrient status, pH gradients do not necessarily parallel nutrient gradients. Often, concentrations of nutrients such as NO3

–, 

NH4

+, and PO43- vary more in individual wetlands due to hydrologic factors than 

among fens of different types (Bedford and others 1999; Bragazza and Gerdol 

2002; Bragazza and others 2005). For example, large differences in nutrients are 

often seen between the edge and center of peatlands (Bragazza and others 2005; 

Cooper 1990a), although not in all fens (Cooper and Andrus 1994). Variation 

also can occur vertically within a peat profile (Williams and others 1979).

Microsite development—Peatlands commonly develop microtopographical 

features, including hummocks, peat mounds, ridges (strings), and pools (hollows 

or flarks) (Figure 29) (Cooper and Andrus 1994; Foster and others 1988; Glaser 

1987). The development of microtopographic features is pronounced in patterned fens (termed “aapamires” in the European literature), which are generally 

restricted to boreal and subarctic regions, although patterned fens occur throughout the Rocky Mountains (Chadde and others 1998; Windell and others 1986). 

Microtopographic features form in other fen types, especially peat hummocks.
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Microtopographic features develop as a function of variable water flow, vegetation development, and peat accumulation. Strings and flarks are typically arranged perpendicular to the predominant direction of water flow (Glaser 1987; 

Grittinger 1970). Spatial variability in hydrologic regimes drive differences in 

plant composition, productivity, and decomposition, which over time, lead to the 

development of microtopographic features (Malmer and Wallén 1999).

Hydrologic variables, pH, and cation concentrations can vary considerably among 

microsites, influencing vegetation patterns and providing a diversity of niches 

for individual species. The resulting fine-scale environmental heterogeneity is 

one factor contributing to the high species diversity found in many fens. Many 

species commonly have distinct affinities for particular microsites, particularly 

bryophytes (Bragazza 1997; Pykala 1994; Vitt and Slack 1984). Autogenic 

processes of peat accumulation and microsite development also influence the 

temporal dynamics of peatland vegetation. For example, an increase in the relative depth to the water table resulting from vertical accumulation of peat has 

been found to promote increased shrub cover and shifts in the distribution of 

monocot species (Bragazza 2006).

Figure 29—Fen on the Roosevelt National Forest, Colorado, near Left Hand Reservoir, exhibiting 

well-developed microtopography. Note the alternating pools (dark areas) and strings.
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Concentrated groundwater discharge can lead to the development of peat mounds, 

convex accumulations of peat that rise above surrounding areas. In the Nebraska 

Sandhills, these occur where hydraulic pressure pushes up against a layer of 

peat and are functionally similar to the spring mounds found in some montane 

fens (Figures 25, 30). Peat mounds may range in size from a few square meters 

to over a hectare and can rise over 2 m above adjacent fen and meadow areas 

(Borgmann-Ingwersen 1998). Peat mounds may also become hydrologically 

disconnected from groundwater and dry out (Steinauer 1992).

Disturbance processes—Disturbance regimes in fens are a relatively minor 

driver of vegetation dynamics compared to other wetlands types such as riparian areas. Particular disturbances that may be locally important to fens are fire 

and grazing, although neither is well understood. Data from both historical and 

contemporary landscapes are lacking, so inferences must be made from anecdotal 

accounts and case studies examining effects, often from research conducted 

outside of the region.

While fire regimes have been extensively studied for forested ecosystems in 

Region 2, few studies have directly addressed the issue for fens. Unlike boreal 

regions, where peatlands comprise a large proportion of the landscape, fens in 

Region 2 occur as relatively small and discrete patches in a matrix of mostly 

forested landscapes. The likelihood of ignitions occurring in fens is therefore 

small, and fire must spread from adjacent ecosystems to directly affect fens.

Figure 30—Cross section through a Nebraska Sandhills fen illustrating landscape setting 

and groundwater flow patterns (Re-drawn from Borgmann-Ingwersen 1998).
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Fire regimes vary widely within the region’s forests, influenced by cover type, 

physiography, and climate. The latter is a particularly important factor influencing 

landscape-scale patterns of fire occurrence, with regional climatic phenomena 

such as ENSO showing a strong correlation with the frequency of fires and the 

amount of area burned (Brown and Wu 2005; Kitzberger and others 2001; Veblen 

and others 2000). The link between regional climate patterns and fire frequency 

has been demonstrated for Canadian peatlands, where researchers found a positive 

relationship between the area of peatland burned and several climate variables 

calculated using a fire weather index (Turetsky and others 2004). Fire regimes 

of subalpine forests in Region 2 are often relatively long compared to boreal 

landscapes (Kasischke and Turetsky 2006; Sherriff and others 2001; Weir and 

others 2000), suggesting that fire may play a relatively small role in the dynamics of fens.

Since fens typically remain saturated for much of the fire season, their ability to 

sustain fire is low relative to drier upland areas. Although peat will readily burn 

if sufficiently dry, the presence of saturated areas will generally limit the extent 

of peat loss. As a result, the direct impact of fire to fens is likely patchy with 

relatively modest areas experiencing peat loss, except during prolonged droughts. 

The pattern of direct effects are correlated with microtopographic patterns of 

surface relief (Benscoter and others 2005a, 2005b; Benscoter and Wieder 2003).

Fire can have a variety of indirect and direct impacts on fens. Through direct 

plant mortality, fires can influence community composition, particularly as species vary in their resilience to the effects of fire (Bowles and others 1996). For 

example, fire can maintain open conditions in some fens by selectively killing 

woody species (Jacobson and others 1991). Fire in adjacent uplands may also 

indirectly affect fens by increasing water and sediment yield and affecting water 

chemistry. For example, fens may experience a flush of nutrients such as nitrogen 

following fire (Anderson and Menges 1997; Dikici and Yilmaz 2006). As the 

density and cover of upland vegetation increases over time, the magnitude of 

these effects relative to pre-fire conditions likely declines.

Another natural disturbance potentially affecting fens is native ungulate grazing. 

Effects can include plant mortality due to herbivory or trampling as well as effects on water chemistry and nutrient availability though urine and fecal deposits 

(Frank and Evans 1997; Pellerin and others 2006). While anecdotal evidence 

suggests that local impacts to vegetation can be significant, because of variable 

population density and migratory behavior, it is unlikely that native ungulates 

have been a major disturbance agent in fens at broad spatial or temporal scales.

Other disturbance events potentially affecting fens are mass wasting events such 

as landslides. These may episodically contribute pulses of sediment to fens, 

affecting ecological functions such as peat accumulation. However, because 

the physiographic and geological setting of fens is so variable, such events are 

unimportant for most fens.
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The response of fen vegetation to disturbances likely varies in relation to several 

factors, including the spatial extent and severity of disturbance and the composition of the pre-disturbance vegetation. These influence processes of colonization 

and recruitment. For example, some fen species can form persistent seeds banks, 

allowing for rapid re-colonization following disturbance (Leck and Schutz 2005; 

Schütz and Rave 1999; Sundberg and Rydin 2000). The landscape context of 

fens may also be important. For example, the proximity of undisturbed fens 

will influence the availability of propagules for re-colonization (Tackenberg 

and others 2003).

Fen development and succession—The incomplete decomposition of organic 

matter in fens is due to reduced microbial activity in anoxic, waterlogged soils. 

The physical characteristics of all peat soils differ from mineral soils, although 

there is considerable variability among peat types in characteristics such as bulk 

density, cation exchange capacity, and hydraulic conductivity (Mulqueen 1986). 

Peat soils vary depending on the botanical composition of organic material and 

the depth and degree of decomposition (Glaser 1987). For example, peat derived 

primarily from mosses differs from that derived from vascular plants (Shotyk 

1988).

Moss-derived peat is formed in Sphagnum-dominated fens as well as those 

dominated by “brown moss” genera such as Scorpidium, Callerigon, and Drepanocladus (Glaser 1987). However, within many Region 2 fens, peat soils are 

largely derived from sedge species, such as Carex aquatilis, C. utriculata, C. 

limosa, and C. lasiocarpa (Cooper 1990a). The dominant structural components 

of such soils are belowground in origin and are mainly comprised of interwoven 

roots and rhizomes of sedges. Typically, little above-ground material is incorporated into the peat body because above-ground decomposition rates exceed 

production; however, anaerobic conditions below-ground retard decomposition 

and allow for peat accumulation (Chimner and others 2002). Patterns of peat 

accumulation likely differ in fens dominated by bryophytes such as Sphagnum 

spp., where living biomass and net primary production is concentrated aboveground (Reader and Stewart 1972).

Although the predominant source of organic matter in most peatlands is from 

vegetation growing onsite, surrounding areas can contribute large amounts of 

organic matter. For example, needles, cones, and twigs from adjacent Engelmann 

spruce-dominated uplands contributed the majority of sediment influx to fens 

in the San Juan Mountains of Colorado (Cooper and Arp 2002). Large trees can 

fall into peatlands and are often abundant in peat profiles (Carrara and others 

1991; Cooper and Arp 2002). The contribution of mineral sediment to fens may 

be episodic and tied to climate. For example, the stratigraphy of Sandhills fens 

suggests that during dry climatic periods, sand may drift onto fens from the surrounding uplands, creating distinct sand layers within the peat body (Ponte 1995).
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Peat accumulation in fens is ultimately controlled by hydrologic and climatic 

variables. However, because peat development is largely an autogenic process—

driven by the relationship between production and decomposition of organic 

matter within the fen—changes in vegetation composition and microtopography 

associated with peatland development and succession can lead to non-linear peat 

accumulation patterns. For example, modeling and reconstruction of carbon sequestration in a Swedish peatland showed distinct patterns over time, with abrupt 

changes associated with shifts in vegetation and the dominant bryophyte species, 

as well as gradual decreases associated with increasing humification of newly 

formed peat (Belyea and Malmer 2004). However, some studies have found that 

long-term rates of peat accumulation are relatively steady, despite significant 

variability in the short-term rates of peat formation (Belyea and Clymo 2001). 

Such differences may be the result of feedback mechanisms between the rate 

of peat formation and the thickness of the aerobic surface layer (i.e., acrotelm), 

whereby individual microtopographic features such as hummocks, lawns, and 

pools expand or contract vertically in response to fluctuations in the position of 

the water table, compensating for climate-mediated hydrologic variation (Belyea 

and Clymo 2001).

Peat accumulation rates measured within Region 2 fens vary, ranging from approximately 1 to 5 cm/century (Table 21). Mean annual carbon accumulation 

rates for Colorado fens are approximately 25 g C/m2 year (Chimner 2000); however, rates can vary widely in relation to hydrologic and climatic factors, and are 

often negative during dry periods or in hydrologically modified fens (Chimner 

and others 2010). For example, estimated annual carbon budgets ranged from 

–142 to 180 g C/m2·year in a fen in Rocky Mountain National Park, Colorado 

(Chimner and Cooper 2003). Fen ages are also variable, as indicated by radiocarbon dating. Based on 14C dates, some fens appear to have originated soon 

after the retreat of glaciers nearly 12,000 years BP (Cooper 1990a), while others 

are younger (Madole 1976).

Maximum peat depths reported for fens in the Nebraska Sandhills are over 7 m 

in depth (Borgmann-Ingwersen 1998), and radiocarbon dates from basal peat 

cores indicate that many of the fens came into existence soon after the formation 

of the Sandhills, over 12,000 years BP (Ponte 1995). At that time, spruce forests 

dominated the region and as the glaciers retreated and the climate warmed, boreal 

species eventually died and semi-arid grasslands developed. However, several 

plant species with boreal affinities still occur in Nebraska fens (Gage and Cooper 

2006b; Steinauer and others 1996).
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Table 21—Peat depths, radiocarbon dates, and peat accumulation rates for selected Colorado fens.

Elevation Basal date Peat depth Accumulation rate

Location (m) YBP (m) (mm/yr)

Placer Gulch Bog (Carrara and others 1991) 3,600 8,790±260 0.85 0.10

High Creek Windmill Fen (Cooper 1990b) 3,010 8,270±140 0.90 0.11

Green Mountain Pond (Cooper 1990a) 2,865 11,820±170 1.5 0.13

Big Meadows (Cooper 1990a) 2,865 11,230±170 1.5 0.13

Sacramento Creek (Cooper 1990b) 3,100 9,820±150 2.13 0.22

Buffalo Pass (Madole 1980) 3,146 7,730±250 1.93 0.25

East Lost Park Fen (Cooper 1990b) 3,100 10,080±150 2.64 0.26

Iron Bog (Fall 1997) 2,290 8,260±220 2.20 0.27

Silver Lake Bog (Pennak 1963) 2,979 6,190±300 1.75 0.28

Eureka Gulch Bog (Carrara and others 1991) 3,665 6,180±160 2.40 0.29

Cottongrass Fen (Cooper and Arp 2002) 10,460±240 3.4 0.33

Carpenter’s Fen (Cooper 1990b) 3,150 9,280±180 3.20 0.34

McMaster’s Fen 3,175 9,220±110 3.33 0.36

Dome Creek Meadow (Feiler and others 1997) 7,800±100 3.62 0.46

If the necessary conditions of water surplus are present to allow peat accumulation, fens may develop through three general processes. Terrestrialization refers 

to the process by which a water body fills with sediments and peat. In contrast, 

paludification describes the conversion of uplands to peatland through increased 

waterlogging of soils as developing peat impedes drainage. Paludification can be 

initiated by climate change, beaver dams, or geomorphic changes (Mitsch and 

Gosselink 2007). Flow-through succession, also termed topogenous development, 

is intermediate in characteristics between terrestrialization and paludification. 

This process occurs in lake basins that have continuous surface water inflows 

and outflows. Gradual accumulation of sediment and organic matter in the lake 

bottom coupled with the peat accumulation from emergent marsh plants along 

the lake margins eventually result in the lake bottom rising above the water 

level, diverting stream flows around the main peat body (Mitsch and Gosselink 

2007) (Figure 31). Terrestrialization and flow-through succession appear to be 

of greatest importance within Region 2, while globally, paludification is likely 

responsible for most of the peatland area (Bauer and others 2003).

A variety of factors may influence fen development and succession. These include 

basin size, stability of the water level, slope processes, and landform morphology. The developmental trajectory of fens in larger basins can differ from those 

in smaller basins, even within the same small watershed (Figure 32) (Cooper 

and Arp 2002). Research suggests that the development of fens is individualistic 

across landscapes, with varying patterns of peat accumulations in different landform types (Heinselman 1970). The sensitivity of fens to environmental change 

likely varies, with individual responses depending on both peatland type and 

the relative stability of hydrological conditions supporting peat accumulation 

(Bauer and others 2003).
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Figure 31—Schematic cross sections (A-C) illustrating the geomorphic and successional development of a hypothetical 

small kettle pond into a sedge-dominated fen. Early in the basin’s development (A), sedges colonize the margin of the 

recently formed kettle lake, developing over time into a floating mat dominated by densely interwoven rhizomatous species such as Carex lasiocarpa, C. limosa, and Menyanthes trifoliata (B). Ultimately, as the process of terrestrialization 

continues, the open water environment characteristic of earlier stages is lost (C).

Figure 32—General illustration of the development of two fens formed in basins of different 

size over time. Arrows indicate principal direction of surface water. The main peat body in 

the small basin is formed on the basin floor, while a floating peat mat is found during the 

developmental sequence in the larger basin (re-drawn from Moore and Bellamy 1974).
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While successional processes have been extensively studied in peatlands elsewhere 

(Bauer and others 2003; Campbell and others 1997; Glaser and others 2004a; 

Heinselman 1970; Vandiggelen and others 1996), this subject has received little 

study within Region 2. In boreal regions, several successional stages have been 

described related to peatlands, with wetlands proceeding from pond to marsh, 

to rich fen, to poor fen, and finally to bog (Muller and others 2003). The widespread application of this successional model, the classical hydrosere, has been 

questioned for boreal wetlands (Klinger 1996), and it is probably not applicable 

to many Region 2 fens.

Allogenic processes such as regional changes in climate or isostatic uplift have 

been hypothesized to be the dominant control on the development of peatlands 

in boreal and subarctic regions (Glaser and others 2004a, 2004b). Other studies 

have stressed autogenic processes as key drivers of peatland development (Muller 

and others 2003). Each process is important, with the relative dominance of each 

likely varying among individual wetlands and over time (Vitt 1994). In addition, 

there may be a threshold effect in the relationship between climate and peatland 

development, whereby climate variables such as precipitation only limit peat 

accumulation and successional processes past some threshold, with autogenic 

factors controlling peatland development when climate conditions are wetter 

(Muller and others 2003).

There are few examples of successional sequences for Region 2 fens. A general 

successional pattern for basin fens in the Southern Rocky Mountains was presented by Cooper (1990), where open-water aquatic communities dominated 

by Nuphar lutea and Potamogeton gramineus proceed through a Menyanthes 

trifoliata-Carex utriculata stage, into a C. limosa stage and, ultimately, to a C. 

utriculata-C. canescens community. There is some applicability of this model 

to other basin fens, like those in Prospect Basin in the San Juan Mountains of 

Colorado; however, it will not be applicable to sloping fens and more research 

is needed to characterize the range of variation in successional processes.

Fen Classification and Gradients_________________

A number of classification schemes have been applied to wetlands, with varying 

degrees of utility for peatlands. The criteria used to classify features vary among 

classification systems. Examples include dominant vegetation, hydrologic characteristics, geomorphic setting, and water chemistry. Each approach has strengths 

and weaknesses depending on the application.

Classifications based on vegetation are available for many regions. At broad 

scales, vegetation is well suited to mapping from aerial photographs and other 

remotely sensed data; however, most geospatial datasets lack sufficient detail 

to effectively differentiate the vegetation types found in fens. An example of a 

vegetation classification is the National Vegetation Classification System used 

by Natural Heritage programs (Comer and others 2003a, 2003b).
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Classifications developed for all wetland types such as the Cowardin system used 

by the National Wetlands Inventory program are of limited utility for discriminating peatlands or among peatland types (Cowardin and others 1979). Under 

the Cowardin system, all fens fall within the palustrine system and, depending 

upon their vegetation, may be placed in the moss-lichen, forested, emergent, or 

scrub-shrub class. Fens in Region 2 would generally be classified as either (1) 

palustrine, emergent, persistent, with a saturated water regime and organic soils, 

or (2) palustrine, scrub-shrub with a saturated water regime and organic soils.

Another widely used classification scheme for wetlands is the Hydrogeomorphic 

(HGM) approach (Brinson 1993a; Cooper 1998a). HGM emphasizes physical 

variables in classification, specifically geomorphic setting and hydrologic regime. 

Because these characteristics are such important drivers of ecological processes 

in wetlands, they provide a robust framework from which to evaluate the ecological functions of individual wetlands. The original framework was intended to be 

augmented by publication of regional guides; however, relatively few of these 

have been prepared (Hauer and others 2002a, 2002b), and none deal specifically 

with fens in the region.

A preliminary classification of Colorado wetlands using the HGM approach was 

prepared for Colorado and three subclasses were identified relevant to fens. The 

Depressional 1 subclass includes mid- to high-elevation basins with peat soils as 

well as lake fringes that may or may not support peat accumulations. Wetlands in 

the Slope 1 HGM subclass include alpine and subalpine fens and wet meadows on 

saturated, non-calcareous substrates, while those placed in the Slope 2 subclass 

include Subalpine and montane fens and wet meadows on saturated calcareous 

substrates (Carsey and others 2003; Cooper 1998a). Johnson (2005) developed 

an HGM classification for wetlands within Summit County, Colorado, based 

on edaphic, hydrologic, and geomorphic characteristics as part of a landscape 

analysis of cumulative impacts to wetlands.

A variety of other classification systems have been developed specifically for 

peatlands. For example, Moore and Belamy (1974) classified peatlands based 

on processes of hydrologic characteristics, identifying three main categories: 

rheophilous, transition, and ombrophilous. Rheophilous peatlands are influenced 

by groundwater from outside the immediate watershed; transition peatlands are 

influenced by groundwater derived only from the immediate watershed; and 

ombrophilous peatlands are not influenced by groundwater (Moore and Bellamy 1974). Subdivisions within these groups are made on the basis of whether 

flow is intermittent or continuous and whether flows occur on the peat surface 

or underneath a floating vegetation mat. As noted earlier, ombrophilous systems 

do not occur in Region 2 because evapotranspiration rates are too great and 

precipitation is insufficient to support them.
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Although the original classification system developed for Scandinavian peatlands 

referred to species richness, pH is often used to differentiate peatland types 

(Figure 33), because pH is often strongly correlated with other water chemistry 

characteristics and vegetation composition (Cooper and Andrus 1994; Windell 

and others 1986). Bogs and poor fens are highly acidic, while pH values greater 

than 8 have been reported from extreme rich fens in Colorado (Johnson and 

Steingraeber 2003). Cation concentrations in peatlands are generally assumed 

to vary along pH gradients, with low concentrations of cations such as calcium 

and magnesium at low pH values (Figure 33). However, many studies have 

demonstrated that the two can often vary (Bragazza and Gerdol 2002).

The majority of fens in the region would likely be described as transitional or moderately rich fens based on water quality and floristic criteria (Cooper and Andrus 

1994; Johnson and Steingraeber 2003). While the main hydrological distinction 

between peatland types is between ombrotrophic bogs and minerotrophic fens, 

research suggests that, when gradients in chemistry and vegetation composition 

are included, the primary division among peatlands should be between acidic 

Sphagnum-dominated bogs and poor fens and alkaline, brown-moss-dominated 

rich fens (Vitt 1994).

Figure 33—General relation between pH and Ca2+ concentrations 

characteristic of different peatland types. Note that the actual relationship between variables can vary widely.
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The most rare fen types in the region are iron fens and extreme rich fens. Both 

represent extreme physiochemical environments and support species that are 

either very rare or completely unknown from other fens in the region. Iron fens 

are found in a limited number of sites in Colorado and the Black Hills (Cooper 

2003; Glisson 2003). The reduction of pyrite in groundwater entering these fens 

produces highly acidic conditions, and poor fen species such as Sphagnum angustifolium, S. fuscum, S. balticum and Drosera rotundifolia occur (Cooper and 

others 2002; Wolf and others 2006). Iron fens are unusual in having very low pH 

and supporting vegetation typical of ombrotrophic bogs and poor fens, but with 

mineral ion concentration more typical of rich fens (Cooper 2003). For example, 

water in the Redwell iron fen in Colorado was found to have conductivity of 304 

μS and very high levels of lead, zinc, cadmium, iron, aluminum, yet it had a pH 

of only 3.5 (Rocchio and others 2003).

Most mountain ranges in the region are dominated by igneous and metamorphic 

rock. However, a handful of areas such as the Mosquito Range in Colorado and 

the Gros Ventre Range in Wyoming are composed of calcareous bedrock, 

such as limestone and support extreme rich fens (Figure 34) (Cooper 1996). 

Figure 34—Aerial photo (A) and close-up shots of High Creek fen (B and C), an extreme rich fen in 

central Colorado. White areas in panel A are accumulated magnesium and sodium salts.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 142

These are among the rarest fen types in North America, having only been described in small portions of California, the Great Lakes region, Iowa, Alaska, 

and Canada (Johnson and Steingraeber 2003). These fens have extremely high 

calcium and magnesium concentrations and pH values greater than 7.0. Due in 

part to their unique chemical characteristics, these fens support a range of species and communities found nowhere else in Region 2 (Cooper 1996; Cooper 

and Sanderson 1997; Johnson 2003; Lesica 1986).

Fen Vegetation in the Region ____________________

Fens may be floristically rich, supporting numerous rare species and community types. For example, fens harbor over 10 percent of the state plant species 

of special concern in Wyoming; while on the Shoshone National Forest, they 

support 26 percent of the state species of concern (Heidel and Laursen 2003b). 

Many regionally rare species have circumboreal affinities and are at or near the 

edge of their range in Region 2 (Gage and Cooper 2006a, 2006b; Hulten 1968; 

Weber 2003).

Natureserve (2006) recognizes approximately 20 fen vegetation associations, 

the majority of which occur as part of the Rocky Mountain Subalpine-Montane 

Fen ecological system (CES306.831). A variety of additional classifications have 

been developed outside of the National Vegetation Classification System as part 

of site-specific studies of fens (Cooper 1990a; Cooper 1996; Cooper and Andrus 

1994; Cooper and others 2002; Johnson 1996; Johnson 2003). Vegetation in fens 

is typically dominated by species of Carex, Juncus, Eleocharis and Calamagrostis 

(Windell and others 1986) (Table 22), and while a variety of herbaceous dicots 

also inhabit fens, overall cover is generally low. Clonal sedges such as Carex 

utriculata and C. aquatilis are among the most common dominants, although 

a wide variety of other sedges also occur. Communities dominated by grasses 

and rushes are less common and are often most prevalent along fen margins, 

particularly where fens abut other wetland types such as wet meadows. Examples 

of common species are Calamagrostis canadensis, Calamagrostis stricta, and 

Juncus arcticus.

Although the cover of herbaceous dicots is typically low, some species such 

as Menyanthes trifoliata are dominants in certain community types. Common 

dicots include Pedicularis groenlandica, Caltha leptosepala, and Polygonum

bistortoides. Shrub-dominated communities, sometimes referred to as carrs, are 

also common in montane and subalpine environments. The most common shrub 

species in the region are Salix planifolia, S. wolfii and Betula nana. In contrast 

to boreal regions, ericaceous shrubs are often a minor component of vegetation 

in Region 2 peatlands. Forested fens (“treed fens”) are also widely distributed 

in subalpine environments, although these appear to be less common than in 

boreal regions (Johnson 1996, 1997). The most common overstory dominants 

in treed fens are Picea engelmannii, P. glauca and Pinus contorta, which have 

genotypes that can tolerate the saturated conditions found in fens.
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Table 22—Fen vegetation associations recognized under the National Vegetation Classification System in 

Colorado, Wyoming, Nebraska, South Dakota, and Kansas (NatureServe 2003; NatureServe 2005). 

NVCS ID Association Name

CEGL001562 Carex utriculata Herbaceous Vegetation

CEGL001806 Carex buxbaumii Herbaceous Vegetation

CEGL001810 Carex lasiocarpa Herbaceous Vegetation

CEGL001811 Carex limosa Herbaceous Vegetation

CEGL001825 Carex simulata Herbaceous Vegetation

CEGL001831 Dulichium arundinaceum Seasonally Flooded Herbaceous Vegetation

CEGL002739 Ledum glandulosum Shrubland (Cohen and others 2003)

CEGL002898 Carex aquatilis-Sphagnum spp. Herbaceous Vegetation

CEGL002899 Betula nana/Sphagnum spp. Shrubland

CEGL002900 Kobresia myosuroides-Thalictrum alpinum Herbaceous Vegetation

CEGL002901 Kobresia simpliciuscula-Trichophorum pumilum Saturated Herbaceous Vegetation

CEGL002922 Carex utriculata Perched Wetland Herbaceous Vegetation

CEGL005887 Betula nana/Carex spp. Shrubland

CEGL002189 Betula pumila-Salix spp. prairie fen shrubland

CEGL001188 Salix candida/Carex utriculata Shrubland

CEGL002383 Carex lasiocarpa-Calamagrostis spp.-(Eleocharis rostellata) Herbaceous Vegetation

CEGL002390 Carex interior-Eleocharis elliptica-Thelypteris palustris Herbaceous Vegetation

CEGL002041 Carex pellita-Carex spp.-Schoenoplectus tabernaemontani Fen Herbaceous Vegetation

CEGL002267 Carex prairea-Schoenoplectus pungens-Rhynchospora capillacea Herbaceous Vegetation

CEGL002268 Carex spp.-Triglochin maritima-Eleocharis quinqueflora Marl Fen Herbaceous Vegetation

Fens in the region also support extensive bryophyte cover. Sphagnum spp.-

dominated fens are uncommon in Region 2 compared to boreal regions, and are 

largely restricted to acidic peatland types such as iron fens and some nutrient poor 

fens. Sphagnum species can also occur in microenvironments, such as on hummock tops, in more minerotrophic peatlands (Figure 35) (Johnson 1996; Zoltai 

and Johnson 1985). Although Sphagnum species are most common in very acidic 

peatlands, some species, for example S. contortum, can be found with rich fen 

indicators (Janssens and Glaser 1986). Many of the Sphagnum species found in 

the region occur in just one or a few locations (Cooper and others 2002). More 

commonly in Region 2, bryophyte communities are dominated by brown-moss 

species in the genera Drepanocladus, Aulacomnium, and Tomentypnum, which 

are characteristic of rich fens.

Species diversity varies among fen types and is typically lowest in nutrient-poor 

systems such as poor fens or in microsites characterized by either extremely wet 

or acidic conditions. While fens with very acidic or alkaline pH often support 

low species richness, they are important for regional biodiversity, supporting 

species that are rare elsewhere in Region 2 (Cooper and others 2002; Cooper 

and Sanderson 1997; Rocchio 2006b). The most important factors influencing 

patterns of plant occurrence and distribution differ among species and life forms 

(i.e., moss vs. sedge) (Gignac 1994). For example, vascular plant distribution 

appears to be influenced more by nutrient levels, while mosses respond more to 

gradients of acidity and cation concentration (Chee and Vitt 1989).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 144

Figure 35—Bryophyte-dominated hummocks in fen, Rocky Mountain National Park, Colorado 

(photo: E. Gage).

The relative importance of environmental variables also differs depending on the 

range of peatland types being analyzed. For example, water and peat chemistry 

are typically the primary factors segregating habitat types when the peatlands 

that fall along the full pH range are compared, while water table depth gradients 

assume primacy when acidity-alkalinity gradients are more narrowly defined 

(Belland and Vitt 1995; Bragazza and Gerdol 1999; Cooper and Andrus 1994; 

Vitt and others 1995). For example, the wettest microsites (e.g.,, floating mats, 

pools, and water tracks) may support species such as Carex lasiocarpa or C. 

limosa, while species such as C. utriculata, S. planifolia, and Calamagrostis 

canadensis are common along fen margins (Bragazza and Gerdol 1996; Cooper 

and Andrus 1994; Slack and others 1980). Often, vegetation responds to complex 

gradients comprised of several physiochemical variables. For instance, the mire 

margin-mire expanse gradient described for many peatlands is thought to include 

several ecological factors related to the degree of minerotrophy, peat aeration, 

and hydrology (Bragazza and others 2005; Wheeler and Proctor 2000).

Vertical zonation of vegetation is common in fens with well-developed microtopography. Different Sphagnum spp. often occupy distinct niches on hummocks 

or peat mounds (Figure 36) (Andrus and others 1983; Mulligan and Gignac 

2001; Vitt and others 1975). Species often segregate along gradients of water 

table and water chemistry and are thought to be shaped in part by competitive 

interactions among species (Bragazza 1997). Although more extensively studied 

for Sphagnum species, brown-moss species also show distinct niche separation 

(Kooijman and Bakker 1995; Kooijman and Hedenas 1991).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 145

Figure 36—Relationship of five Sphagnum species to pH along a 

hummock-hollow complex in a northern Michigan peatland (re-drawn 

from Vitt et al. 1975).

Vegetation composition can vary widely among fens, even those located in close 

proximity to one another. Such differences can be caused by differences in fen 

morphology, hydrology, and chemistry, as influenced by age and geomorphic 

setting (Cooper 1996; Cooper and Andrus 1994; Cornwell and Grubb 2003; 

Locky and Bayley 2006). For instance, although located in the same watershed 

and geological setting, five different fens in the San Juan Mountains of Colorado 

had little overlap in the dominant communities present in each fen (Cooper and 

Arp 2002). In general, moderate rich fens share more species in common with 

extreme rich fens than with poor fens, although the three fen types are separated 

by surface water pH, calcium, magnesium, and conductivity (Chee and Vitt 1989).

Diversity within individual communities is often low, as many fen dominants 

are clonal and form relatively uniform stands. However, species diversity within 

entire fens can be high, particularly where significant structural complexity in the 

form of fallen trees, strings, flarks, or ponded water is present. Fen age may also 

be a contributing factor; older fens may act as refugia, supporting vegetation once 

more widespread in the region (Cooper and Arp 2002). Vegetation patterns in 

fens often exhibit distinct zonation patterns. Complex gradients broadly structure 

vegetation patterns in relation to the morphology and hydrology of the wetland 

(Cooper 1996; Cooper and Andrus 1994; Johnson 1997). The patterns of vegetation structure, as well as the relative importance of environmental processes on 

vegetation, vary depending on the scale evaluated (Hajkova and others 2004).

Among the most important factors influencing patterns of species distribution 

and abundance are variation in water table elevation, pH, degree of minerotrophy, 

water temperature, and substrate type (Bragazza and Gerdol 1999; Cooper 1990a; 

Johnson and Steingraeber 2003). Complicating interpretation, the particular 

gradients of greatest importance for vascular species and bryophytes can differ 

(Bragazza and Gerdol 1996; Gerdol 1995; Hajkova and Hájek 2004). In addition 

to within fen variability, vegetation can widely differ among fens in the same 

watershed. Differences may be the result of variation in relative age, hydrologic 

regime, and geomorphic setting of fens.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 146

HRV of Fens in Region 2 ________________________

Overview of anthropogenic impacts

Humans have altered fens in a number of ways. These activities and their associated impacts can be broadly grouped into on-site and off-site impacts, although 

hydrologic modifications can be in both groups. The extent and magnitude of 

anthropogenic impacts to fens varies widely geographically and by fen type. 

Some impacts (e.g., peat mining) are locally severe but affect a relatively small 

area of fen, while others (e.g., atmospheric deposition of pollutants) may be more 

widespread but are subtle in their ecological effects (Chimner and others 2010).

Peat mining—Peat mining, including pre-mining drainage, is the most severe 

site-level impact to a fen because the entire peat body, vegetation, and soil seed 

bank are removed (Figure 37). Peat extracted from fens in the region comprise 

a small portion of the total national supply, as most commercial peat originates 

in Canada (Crum 1988). However, because the total area of fens in the region is 

so small, impacts from even small peat mining operations can be significant. In 

addition to changing the substrates available for plant growth, peat mining may 

cause significant hydrologic changes. Fens are typically drained using ditches 

prior to mining (Crum 1988), resulting in lower water table elevations and the 

oxidation of any peat that remains (Laiho 2006). The harvesting of peat reduces 

water storage capacity and the ability of fens to trap pollutants and often results 

in negative effects on water quality (Cooper 1990b).

Figure 37—Fen in the Mosquito Range of Colorado mined for peat. Note that even after more than 

two decades, almost no plants grow from natural colonization in the mined area (Photo: D. Cooper).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 147

In Colorado, it is estimated that 200 to 500 acres of fen have been impacted by 

peat mining, representing less than 1 percent of the total fen area in the state 

(USDI FWS 1997). No estimates are available for other states in the region. 

While the total area of fen impacted is not large, most of the mining has occurred 

in just a few areas and many ecologically significant fens have been impacted. 

For example, High Creek fen in South Park, Colorado (the best representative 

of an extreme rich fen in the southern Rocky Mountains) was partially mined 

prior to its purchase by The Nature Conservancy (Cooper and MacDonald 2000; 

Johnson and Steingraeber 2003; Sanderson and March 1996).

Mineral extraction—Lode mining represents one of the most significant environmental impacts in Region 2 (Limerick and others 2005; Wohl 2005). Although mining within fens is uncommon, instances of direct impacts have been 

documented (Figure 38). Direct effects include peat removal, plant mortality, 

changes to community structure, and alteration of groundwater flow paths.

More important to fens on a regional scale are the indirect effects of mining 

operations in watersheds where fens occur. The most important of these are 

changes in the quality and quantity of surface water and groundwater entering 

fens. Frequently, the water draining from waste-rock piles, mill tailings, and 

mine adits has very low pH and elevated concentrations of heavy metals due 

to the exposure of sulfide minerals (e.g., pyrite). These effects on water quality 

often persist long after the cessation of mining activities (Balistrieri and others 

1999; USDA FS 1995).

Figure 38—Mining within an iron fen, near Ophir, Colorado (Photo: R. Chimner).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 148

Generically referred to as acid rock drainage, mine-contaminated waters can 

negatively impact aquatic organisms as well as a variety of ecosystem processes 

such as plant production and decomposition. Research in Colorado fens found 

significant differences in the decomposition rate of Carex aquatilis between 

pristine and acid rock drainage-affected fens. Decomposition rates in pristine 

wetlands were comparable with boreal rich fens, while those in polluted fens 

were more similar to poor fens and bogs (Arp and others 1999).

Direct hydrologic modifications—Direct hydrologic modification is one of 

the most widespread impacts to fens, although no statistics are available on the 

extent of such modifications. Fens have been ditched and drained for a variety 

of purposes, most of which are secondary in nature such as to increase forage 

production and accessibility for livestock. In addition, ditching typically precedes 

other activities such as peat mining or real estate development (Crum 1988; 

USDI FWS 1997).

The most direct impact to fens from ditching is lowering of the water table, 

the effects of which can persist long after the ditches are abandoned. Ditches 

through Big Meadows fen in Rocky Mountain National Park, Colorado, were 

still effectively intercepting and diverting water nearly 75 years after abandonment (Cooper 1990a; Cooper and others 1998). The resulting lower water tables 

facilitated invasion of the fen by Deschampsia cespitosa, a native grass common 

in wet meadows but not fens (Cooper and others 1998). Similar effects may 

also promote invasions by non-native species (Figure 39) and a reduction in the 

suitability of habitat for species found in wet microsites such as Carex limosa 

(Gage and Cooper 2006b). Changes to hydrologic regimes can result in the loss 

of fen obligates and changes in the successional trajectory of impacted wetlands 

(Jukaine and others 1995).

Drainage of fens also impacts the processes of peat accumulation and carbon 

storage. Effects include increased CO2 flux from peat decomposition, an effect 

exacerbated by enhancement of decomposition by the addition of easily decomposable material from root exudation (van Huissteden and others 2006). Drier 

conditions in fens can cause a shift from net accumulation of carbon to loss 

through oxidization (Chimner and Cooper 2003). Such changes may eventually 

result in the complete loss of the peat body.

In addition to ditching, groundwater pumping may deplete aquifers and represent 

an additional threat to groundwater-dependent wetlands like fens (Chimner and 

Cooper 2003). For example, groundwater pumping at Crane Flat in Yosemite 

National Park was lowering the water table in an adjacent fen (Cooper and Wolf 

2006). Hydrologic modifications such as streamflow diversion may reduce the 

amount of water entering fens, deleteriously affecting vegetation and ecological 

processes (Chimner 2000; Chimner and Cooper 2003).
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Figure 39—Diagram illustrating water table in a hypothetical fen 

before (A) and after (B) ditching. The Grand Mesa National Forest, 

Colorado fen pictured in panel (C) has been ditched, lowering the 

water table and facilitating invasion by the non-native Agrostis 

scabra, the reddish vegetation in the back of the photograph 

(photo: D. Cooper). Panel (D) is a photograph from another fen 

on the Grand Mesa National Forest, Colorado, that has recently 

been ditched (photo: G. Austin).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 150

Livestock—Although widely examined in other riparian areas (Marlow and 

Pogacnik 1986; Parsons and others 2003; Roath and Krueger 1982; Vavra and 

others 1994), livestock impacts on fens are little studied in Region 2. Anecdotal 

evidence suggests that livestock spend little time foraging in wet areas like fens. 

However, fen species can be a significant source of forage for livestock (Catling 

and others 1994; Hermann 1970), especially late in the summer or during extended drought.

The most obvious impact to fens from livestock is direct plant mortality, either 

through consumption or hoof action. The effect is not likely experienced equally 

by all plant life forms. For example, species with rhizomes may be selected against 

and taprooted and short-lived species favored under heavy grazing pressure. 

Among clonal species, plant architecture may influence response (Martin and 

Chambers 2001). Some studies have shown that species with fast spreading clonal 

organs such as stolons are favored by grazing, as they can effectively colonize 

gaps such as hoof prints (Stammel and Kiehl 2004; Stammel and others 2003).

Since establishment and recruitment from seed is thought to be relatively rare 

in fens, species incapable of rapid vegetative growth may be more vulnerable to 

trampling impacts. However, disturbance from livestock may positively benefit 

some species by creating germination niches for seeds (Smith and others 2003; 

Stammel and others 2003). However, the negative effects of soil compaction and 

altered patterns of water and light availability on diversity probably outweigh 

positive effects such as the promotion of subordinate species; overall, livestock 

use appears to negatively affect species richness in fens (Stammel and Kiehl 

2004). Additional changes include shifts in the relative dominance of species 

through effects on the competitive ability of plants (Ausden and others 2005; 

Martin and Chambers 2001; Smith and others 2003). Trampling has been linked 

to decreases in moss cover and richness and increases in bare peat (Arnesen 

1999; Rawes and Hobbs 1979).

Additional effects associated with livestock include hummock development, as 

rhizomes of clonal sedges are broken by hooves (Figure 40). Waterlogged peat 

soils are highly susceptible to displacement and hoof punching, and recovery 

time of sites to pre-disturbance conditions can be decades to centuries long 

(Middleton 2002; Smith and others 2003). Trampling may increase the amount of 

bare peat, leading to increases in surface runoff and erosion (Pellerin and others 

2006). Livestock can also affect fens through the addition of nitrogen via urine 

and feces (Blank and others 2006).

As with cattle, native ungulates can have a significant effect on fens. Elk, moose, 

and deer, for example, can alter nitrogen availability though urine and feces inputs 

(Frank and Evans 1997). However, the greatest impact from native ungulates is 

likely trampling. For example, trampling by deer destroyed surface vegetation and 

increased the area of bare peat in Quebec peatlands (Pellerin and others 2006). 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 151

Figure 40—Fen subjected to livestock herbivory. Note difference in vegetation at (A) fenceline, 

(B) hummock formation, and (C) hoof prints in peat body (photos: D. Cooper).

Presumably, where ungulate populations have remained within their historical range of variation, impacts to fens have been modest and localized. Where 

populations have exceeded their HRV, as has been documented in several areas 

in the region (Coughenour and Singer 1996; Lubow and others 2002; Singer and 

others 1998), impacts may be more significant.

Water development projects—Most of Region 2 has experienced water resource development, including stream diversions and storage reservoirs in the 

mountains and plains. As noted earlier, stream diversions affect many of the 

region’s streams, reducing surface water inputs to fens as well as the amount of 

water entering aquifers necessary to support wetlands. Construction of storage 

reservoirs are an additional direct impact. Because fens occur in topographic low 

points, impoundments may lead to their flooding (Figure 41). No statistics are 

available to evaluate the extent of historical impacts to fens.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 152

Figure 41—Fen in 1956 and in 2004 following construction of Left Hand Reservoir, Boulder County, 

Colorado, illustrating impacts from water resource developments. Note that in the 1956 photograph, the 

fen had already been ditched and mined in the bottom left part of the photo.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 153

Roads and transportation infrastructure—Roads and related infrastructure 

can directly and indirectly impact fens. Where roads are constructed through 

wetlands, they can directly impact vegetation and hydrologic processes (Forman 

and Sperling 2002). In addition, transportation infrastructure can indirectly affect fens through changes in watershed-scale hydrologic processes. Culverts 

and ditches associated with roads and trails can alter natural drainage patterns, 

reduce interception and infiltration rates due to the removal of vegetation and 

soil compaction, and alter the hydrologic response of basins to annual snowmelt 

runoff and storm events (Jones and others 2000). Typical hydrologic responses 

involve more rapid overland flow and hydrograph responses to precipitation events, 

potentially causing erosion or increases in the sediment flux to affected fens.

Roads may impact fens by degrading water quality, thereby negatively impacting 

plant species that are dependent on specific water chemistry (Panno and others 

1999; Wilcox 1986a, 1986b). Specific pollutants include combustion by-products 

and oil from vehicles, as well as chemicals used for dust abatement. Road salt 

used for de-icing may negatively impact fen species, particularly bryophytes 

(Pugh and others 1996; Wilcox and Andrus 1987). At sufficient concentrations, 

runoff can cause significant decreases in species richness, evenness, and abundance (Richburg and others 2001). Although reliable statistics regarding road 

abundance are available, there are no data from the region specifically examining 

the relationship between roads and fen function.

Roads may indirectly affect fens by facilitating the spread of exotic plant species. 

Disturbances from road construction or maintenance create sites conducive to 

the establishment of many exotics, while vehicles and increased human traffic 

can serve as dispersal agents for weed propagules (Gelbard and Belnap 2003; 

Parendes and Jones 2000). Off-highway vehicle (OHV) trespass onto fens occurs in many National Forests, disturbing fen vegetation and causing ruts that 

are capable of intercepting sheet flow on the surface of fens. OHV use in or 

near wetlands may also affect fens by contributing pollutants from inefficient 

combustion and engine emissions (Havlick 2002). Although such impacts may 

only affect a small number of fens, the effects are slow to heal.

Exotic species—Although a variety of exotics can invade wetlands, fens are 

not conducive to aggressive exotics such as Canada thistle (Cirsium arvense) or 

introduced pasture grasses such as timothy (Phleum pratensis). These species 

are more characteristic of wet meadows and typically only invade fens following 

hydrologic alterations, such as ditching. Under conditions of excessive nutrient 

loading, native species such as Calamagrostis canadensis or Typha latifolia 

can overtake wetlands, including fens, forming monotypic stands and reducing 

community diversity (Drexler and Bedford 2002; Zedler and Kercher 2004).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 154

Beaver trapping—Beavers may influence the form, function, and successional 

dynamics of some fens (Chadde and others 1998). In low-gradient landscapes, 

beavers may initiate fen formation by creating ponds that are subsequently invaded 

by sedges and bryophytes. They can also influence landscape-level patterns of 

diversity by maintaining a mosaic of wetlands in different successional stages 

(Chadde and others 1998; Mitchell and Niering 1993). Although their importance to riparian ecosystems is widely recognized, few studies have examined 

the influence of beavers on fens in Region 2, making estimation of the potential 

impacts of trapping impossible. Presumably, the widespread trapping that occurred in many watersheds at least indirectly affected some fens by removing a 

major source of hydrologic variability from the landscape.

Logging—Timber harvest, like fire, has the potential to affect fens largely 

through indirect changes to surrounding landscapes. Reduced interception and 

infiltration can alter surface runoff from basins, evapotranspiration rates, and 

snow accumulation patterns. For example, the amount of precipitation reaching the forest floor increased by approximately 40 percent, and peak snowpack 

water equivalent increased by more than 35 percent following canopy removal 

in a subalpine watershed in Colorado (Stottlemyer and Troendle 1999, 2001). 

Annual and peak stream flows in logged watersheds generally increase following harvest (Troendle and King 1987), although how such changes affect fens 

is difficult to predict considering the importance of groundwater recharge and 

discharge processes on fens. Because fens in Region 2 form only in physically 

stable locations with limited erosion and mineral sediment deposition, where 

large increases in sediment yields result from upland vegetation removal, fens 

can be negatively impacted.

In subalpine watersheds, the main hydrologic pathway for snowmelt is subsurface 

flow (Stottlemyer and Troendle 1999). By changing basic hydrologic processes, 

timber harvest can result in changes to meltwater chemistry (Stottlemyer and 

Troendle 1999). For example, increases in the content of Ca2+, NO3

-

, and NH4

+

in meltwater and the seasonal flux of K+, Ca2+, SO4

2-, NO3-

, and HCO3

-

 increased 

following logging in Colorado. Within decades of harvest, hillslope hydrological 

processes returned to near pre-harvest levels (Troendle and King 1985). Whether 

fens respond to changes of this type is unknown. Since many fens are over 10,000 

years old, overall effects at the landscape scale are likely limited.

Atmospheric deposition of pollutants—The atmospheric deposition of pollutants occurs throughout the world and is a major research focus for many 

scientists. Of particular concern are nitrogen and sulfur emissions that originate 

from vehicles and industrial as well as agricultural activities. Intensification of 

agricultural practices, including the introduction of commercial fertilizers and 

increases in the number of livestock in confined feedlots has occurred over the 

last 60 years (Baron and others 2000). Other activities contributing to changes 

in the N cycle include: biomass burning, land clearing and land cover conversion, and the drainage of wetlands, which are a major N sink due to the process 

of denitrification (Vitousek and others 1997).
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Oxidization of sulfur and nitrogen compounds in emissions leads to the formation 

of nitric acid and sulfuric acid, which can contribute to the acidification of lakes, 

streams, and wetlands (Battarbee and others 1990; Charles 1990; Moore 2002). 

In the Colorado Front Range, nitrogen emissions have increased dramatically 

during the Twentieth Century, driven by a growing population and intensive 

agricultural activities (Baron and others 2000). Many catchments in areas with 

high-deposition may be nitrogen saturated (Bowman and others 2006), suggesting an increasing role for nitric acid in the acidification of surface waters 

and wetlands. In addition, because nitrate is a mobile anion capable of moving 

through soils to streams and groundwater, it can pull cations with it, depleting 

the soil of minerals such as calcium (Vitousek and others 1997).

A range of ecological responses to atmospheric deposition of pollutants can 

occur, including changes in soil fertility, vegetation type, and acidification of 

lakes, streams, and wetlands (Vitousek and others 1997). Studies in calcareous 

fens in central Europe have documented increases in plant biomass and compositional shifts to more generalist species following N additions (Pauli and others 

2002). Effects and response thresholds likely differ between vascular species and 

bryophytes. For example, in Canadian bogs and fens, the response of individual 

species varied but overall, moss productivity increased in response to nitrogen 

deposition, while productivity of Betula pumila and Ledum groenlandicum, shrub 

species also examined, was unchanged (Li and Vitt 1997; Vitt and others 2003). 

Nitrogen enrichment can alter microbial community composition in peatlands 

(Gilbert and others 1998), although this has not been analyzed for Region 2 fens.

Using interpolated maps of pollutant concentrations such as nitrate and sulfate, 

Nanus and others (2003) found an increasing spatial trend in the concentration and 

deposition of pollutants from north to south in the region. The authors identified 

hot-spots of atmospheric deposition such as northern Colorado, where the highest nitrate (2.5 to 3.0 kg/ha N) and sulfate (10.0 to 12.0 kg/ha SO4) deposition 

rates were observed. Research comparing ecosystem processes and properties 

on opposite sides of the Continental Divide in the Colorado Front Range found 

significant differences between the eastern side, which is subject to elevated N 

deposition from urban, agricultural, and industrial sources, and the western side, 

which experiences lower deposition rates. Soil percent N is higher and C:N ratios 

lower east of the Divide, while lake NO3 concentrations are also significantly 

higher on the east side, which drives changes in diatom community composition toward eutrophic communities (Baron and others 2000; Rueth and Baron 

2002). High nitrogen deposition has also been documented in the Park Range 

and Medicine Bow Mountains—an area supporting a high density of fens.
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Cumulative effects—Cumulative effects include any ecosystem changes that 

are influenced by multiple land use activities (Reid 1993). In many instances, 

more than one anthropogenic factor has affected fens directly or indirectly. As 

groundwater-fed wetlands, the hydrologic function of fens is particularly dependent on surrounding watershed processes, and factors well outside the immediate 

boundaries of fen can significantly influence hydrologic and ecological processes 

(Siegel 1988). Many individual ecological stressors act synergistically; thus, 

evaluating the effects of single factors in isolation may not adequately capture 

the true impacts to fens (Preston and Bedford 1988). For example, some research 

suggests that the impacts of nitrogen deposition may be exacerbated by grazing 

(van der Wal and others 2003).

Abundance, distribution, and condition of fens across 

Region 2 landscapes

No rigorous, quantitative studies of fen abundance and distribution have been 

conducted for contemporary landscapes across the entire region, let alone for 

historical landscapes. While estimates of the status and trends in wetland area 

have been conducted (Dahl 1990; Dahl 2000; Dahl and Johnson 1991), data are 

not specific to fens and are too coarse to make strong inferences about changes 

in fen abundance or distribution. Estimates have been developed for smaller 

land areas. For example, in Colorado, the area occupied by peatlands was estimated at approximately 100,000 acres or 10 percent of Colorado’s total wetland 

areas (USDI FWS 1997). More rigorous inventories of peatland area have been 

conducted on individual National Forests, usually at the scale of eighth level 

watersheds, but too few data are available to extrapolate to the entire region.

Because peat accumulation rates are so slow, it is reasonable to assume that fen 

area and spatial distribution did not vary in the centuries preceding Euro-American 

settlement of the region. Unlike forest cover types, which may change over decades or centuries in response to disturbance, fens are relatively permanent and 

stable landscape features. The HRV of the number of fens or their cumulative area 

was narrow as few fens were created or lost through natural processes. Although 

quantitative data are not available to compare current versus historical patterns, 

rates of fen loss may be outside of the HRV for many portions of the region due 

to anthropogenic activities. Total fen abundance in more remote wilderness areas 

is likely within the HRV, but intensively used areas like the Nebraska Sandhills 

have likely experienced significant fen loss.
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Fen condition and ecological processes—With the exception of a few wellstudied fens, the functioning and condition of fens in Region 2 is poorly understood. Basic inventories of fens are lacking for the majority of the region, and 

data on fen condition are largely unavailable, at least at landscape and regional 

scales. Several approaches have been developed for wetland functional analysis 

and condition assessment, although their applicability within the region and to 

fens vary. HGM functional assessments have been developed for wetlands in 

portions of Colorado, Nebraska, and adjacent states (Arp and Cooper 2004; Hauer 

and others 2002a; Hauer and others 2002b), although regional HGM guidebooks 

for fens have not been developed. Because actual data on wetland impacts are 

lacking at the landscape scale, broad-scale assessments have largely relied on 

surrogates such as indices of road density and land use/land cover data to infer 

disturbance (Johnson 2005). Several new approaches are being developed for 

site-level assessment; however, these projects are largely in the pilot phase and 

at present, relatively few wetlands have been evaluated (Faber-Langendoen and 

others 2006; Rocchio 2006b, 2006c).

Results from a fen inventory and condition assessment in the San Juan Mountains 

of Colorado documented a range of disturbances to fens (Chimner and others 

2010). The most common types of disturbances were associated with recreational 

and large mammal (native and livestock species), followed by roads and real 

estate development. While some disturbances such as ditching are uncommon, 

they have a more severe impact on fens. Ten percent of the fens sampled were 

highly disturbed, while the majority of fens (80 percent) were judged to be in 

either good or excellent condition.

Vegetation composition and structure—Because fens vary widely in floristic 

composition and only a small percentage have been botanically inventoried, it is 

impossible to assess, in general, the potential departure of vegetation attributes 

from historical conditions. Declines in species richness have been documented 

in response to disturbances (Golinski 2005; Soro and others 1999), however, the 

specific response depends on the nature and severity of the disturbance and type 

of fen being impacted. Because of the slow rate of peat accumulation, recovery 

from particularly severe disturbance may be exceedingly slow. In Colorado, 

significant impacts to vegetation composition, species richness, and vegetative cover were documented in response to peat mining, even many years after 

the cessation of the disturbance (Johnson 2003). A study in Sweden found that 

vegetation composition in mined peatlands was still significantly different from 

undisturbed wetlands even after 50 years (Soro and others 1999). Similar results 

were reported by Cooper and MacDonald (1998), who found that Colorado fens 

subjected to peat mining had only 30 species compared with 122 species in unmined reference areas, and that even after 40 years, the sedges and willows that 

dominated the undisturbed sites were largely absent on the mined sites.
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In the Nebraska Sandhills, disturbed fens often have lower species diversity and 

fewer community types than pristine fens (Borgmann-Ingwersen 1998; Steinauer and others 1996). However, species richness in disturbed sites may not be 

different. For example, some hydrologically modified fens may be invaded by 

exotic and upland species. In such cases, the relatively unchanged species richness may obscure more significant impacts to biodiversity resulting from the 

loss of rare species.

Although data are largely lacking, we assume that many fens have been affected 

by the extensive land use changes in the region. In Ontario, Canada, land uses 

within approximately 250 to 300 m of a wetland boundary affected plant diversity (Houlahan and others 2006). With the exception of wilderness areas, most 

National Forests support extensive road networks (McGarigal and others 2001), 

some of which may have affected vegetation in fens.

A variety of anthropogenic factors can impact the diversity and spatial arrangement 

of community types within an individual fen. Microsite diversity is a primary 

factor influencing species diversity in fen communities. Thus, anthropogenic 

impacts that reduce habitat diversity may strongly impair fen biodiversity. Hydrologic modifications, such as ditches, may eliminate pools and water track 

features, which often support distinct species assemblages. Drier conditions 

created by ditches promote the invasion of facultative meadow species such as 

Deschampsia cespitosa (Cooper 1990a). Historically, willows were removed 

from riparian areas, meadows, and fens (Patterson and Cooper 2007; Peinetti 

and others 2002b). The presence of shrubs creates unique microsites for herbs 

and bryophytes, and may offer protection from native herbivores and livestock.

Fens subject to intensive agricultural uses like haying are likely well out of the 

HRV for fen structural characteristics. This would probably include most fens in 

the Sandhills, where wide expanses of vegetation have been disturbed by ditching 

and haying. One researcher characterized the aerial view of disturbed Sandhills 

fens as smooth, monotypic expanses of grasses and sedges in contrast to relatively undisturbed areas that exhibited distinct patchiness (Borgmann-Ingwersen 

1998). With the exception of fens managed by the Nature Conservancy, nearly 

all known fens are on private lands. Presumably, they are subject to intensive 

use, including ditching, heavy cattle grazing, and the introduction of non-native 

species such as pasture grasses (Borgmann-Ingwersen 1998).

Hydrologic regimes—Because fen formation and the maintenance of peat require 

stable hydrologic regimes, one of the most significant impacts to the functioning 

of fens is through direct hydrologic alteration. Ditches are common in many fens, 

even in remote areas such as National Parks and wilderness areas (Cooper 1990a, 

1998b; Patterson 2005). The effects of disturbance on hydrologic regimes are 

well documented and include increases in seasonal water table fluctuations, a 

reduction in the mean summer water table, and elimination of wet microhabitats 

such as hollows and lawns (Golinski 2005).
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In addition to direct hydrologic impacts, dewatering of fens via ditching can 

allow the invasion of burrowing animals such as pocket gophers and voles. The 

activities of these small animals may create preferential flow paths capable of 

further dewatering of peat body and hastening peat loss due to oxidation (Figure 42). Without active hydrologic restoration, such changes may be irreversible.

Stable hydrologic regimes are necessary for peat accumulation, and a water table 

decline will, over time, result in peat oxidization. Because fens vary widely in 

geomorphic setting and hydrologic functioning, and few long-term hydrologic 

data are available for different fen types, it is difficult to generalize about the 

likely response of fens to past climatic fluctuations. The HRV for key climatic 

factors important to fen hydrologic regimes and carbon accumulation dynamics 

such as the amount and seasonality of precipitation, timing of spring snowmelt, 

and temperature is broad and includes both extended periods of both wet and 

cool conditions and extended dry periods (Blasing and others 1988; Cook and 

others 1999; Woodhouse 2003; Woodhouse and Overpeck 1998).

Fens surely vary in their sensitivity to climatic fluctuations (Winter and others 

2001). Fens supported by several aquifers or with large contributing watersheds 

may be more resilient to major climate shifts than fens with smaller watersheds. 

In contrast, fens formed in more marginal settings, such as those found in association with isolated springs on hill slope locations, may experience alternating 

periods of carbon accumulation and loss.

Indirect changes to fen hydrologic processes can occur as a result of structural 

changes to surrounding uplands. For example, the reduction in forest fire frequency surrounding McIntosh Fen on the Black Hills National Forests has led to 

increased tree cover and reduced groundwater flow into the fen (Glisson 2003). 

Figure 42—Fen dewatered by installation of a ditch during the early Twentieth Century (left panel) and subsequently invaded by pocket gophers, resulting in a proliferation of new channels, further degrading the integrity of 

the peat body (right panel), Drakesbad Meadows, Lassen Volcanic National Park, California (photos: D. Cooper).
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Where key ecological processes in surrounding uplands have significantly departed from their HRV, effects on fens may also be expected to occur, although 

there are insufficient data to be conclusive.

Peat accumulation—Several studies in the region have examined the stratigraphy 

and age structure of peat deposits (Chimner and others 2002; Cooper 1990a). 

However, no studies have attempted to place these patterns into a regional context. There are several methodological issues with the dating of peat profiles, 

including temporal discontinuities, spatial variability in the basal depth and age 

of peat deposits, and the high cost of 14C dating samples. These often prevent the 

collection of large datasets, which would be useful in developing and calibrating 

peat accumulation models at regional scales.

A variety of models describing peatland dynamics have been developed in other 

regions (Belyea and Malmer 2004; Belyea and Warner 1996; Frolking and others 

2001; Korhola 1992; Robinson 2006); however, most are not applicable to Region 

2 because they focus on ombrotrophic bogs. Many models of peat accumulation 

over time incorporate the effects of environmental change on plant productivity but ignore the botanical composition of peatland communities. However, at 

times scales of decades to centuries, changes in plant communities are probably 

an important determinant of the response of fens to climate—both historically 

and into the future (Bauer 2004).

In a study from Europe, Aurela and others (2004) found that inter-annual variation of fen carbon balances was tied to variations in the duration of the snowfree period. Research in the western United States suggests that snowmelt is 

occurring earlier (Cayan and others 2001), and while current climate conditions 

may not have yet departed from the HRV, if climate predictions are accurate, it 

is reasonable to expect further increases in the snow-free period.

Water chemistry—Fens in Region 2 occur in a wide range of chemical environments, from highly acidic iron fens on one end to extreme rich fens on the 

other. However, these fen types are rare, and the majority of fens are classified as 

transitional rich fens with intermediate pH values and ion concentrations (Cooper 

1990a; Cooper and Andrus 1994). With the exception of some microsites such 

as hummocks, there is little evidence to suggest that fens in Region 2 become 

successively more ombrotrophic through time, as has been documented for 

many boreal peatlands. In boreal regions, minerotrophic fens may develop into 

ombrotrophic bogs due to autogenic processes; however, Region 2 fens have a 

relatively stable water chemistry environment through time and do not appear 

to follow the same developmental trajectory. If this is true, it is not meaningful 

to consider a range of historical variability with regard to water chemistry at the 

site level. At landscape and regional scales, the question is whether the relative 

proportion of fen types has fluctuated over time. Since fens in the region owe 

their chemical characteristics primarily to hydrogeological setting and groundwater flow processes and because these processes change little over time, the 

answer appears to be no.
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While there was likely little variability in water chemistry historically, anthropogenic disturbances have probably altered some key fen chemical characteristics 

at both local and landscape scales. For example, water chemistry differed widely 

between disturbed and undisturbed fens on eastern Vancouver Island, with higher 

specific conductivity and cation concentrations observed in disturbed compared 

to undisturbed peatlands, which were attributed to altered drainage and changing 

land uses (Golinski 2005)

An additional anthropogenic impact without historical analogue is the increasing 

deposition of nitrogen in many portions of Region 2. Because the phenomenon 

is so recent relative to the age of fens, it is difficult to predict the effects of N 

enrichment on fen species and ecological processes. However, the variability 

in deposition rates alone across the region appear to represent a departure from 

historical patterns.

Management Opportunities and Constraints _______

Management opportunities

Increasingly, both managers and scientists are recognizing the important ecological 

functions provided by fens. However, as we have made clear in this assessment, 

there are many gaps in our knowledge of fens. More broad-scale assessments of 

fen distribution and abundance are needed. A variety of approaches should also 

be used in these assessments, including aerial photographs or other remotely 

sensed data (e.g., hyperspectral satellite imagery) to identify and map fens.

We also need detailed assessments of fens at the site level. Basic hydrology, 

water chemistry, and vegetation analyses are necessary to develop realistic 

models of fen vegetation dynamics and to understand and evaluate the effects of 

management activities on fens. Such studies may appear prohibitively expensive 

or complicated at first glance; however, installation of even a few groundwater 

monitoring wells can yield invaluable data.

More detailed studies relating geomorphology, peat stratigraphy, hydrologic regimes, and community composition are important for developing an understanding 

of the origin and development of fens in the region. Also needed are more studies examining the physiochemical and hydrologic drivers of fen formation and 

development as well as fen sensitivity to anthropogenic disturbance. Improved 

approaches for restoring degraded fens are also needed. The development of effective restoration techniques would provide managers useful tools to approach 

impaired fens.
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Restoration

Given the degraded condition of many fens in the region, there are numerous 

opportunities for restoration. Processes for evaluating options for restoration 

should incorporate ecological, technical, economic, and social aspects (Brooks 

and others 2006). Because of the complexity of these variables, there is no single 

approach for prioritizing restoration goals. However, there are several practices 

that can facilitate efficient use of limited resources and help ensure successful 

restoration outcomes.

First of all, a reliable inventory and assessment of the planning area is essential, 

as this allows for consideration of all restoration options and selection of sites 

with the greatest potential benefit and likelihood of success. It is useful to include reference sites when developing design standards for restoration projects 

(Cole and others 2006). These can be used to help formulate realistic restoration 

targets and specific design criteria for projects. Once a decision to implement a 

project has been made, a conceptual plan for restoration must be prepared that 

is based on the information collected during the site selection process. The proposed project should be mapped in both plan and profile views with the intended 

hydrologic source clearly identified (Brooks and others 2006). Restoration plans 

should specify the particular species and techniques to be used in planting and 

should identify areas where colonization by seed is to be relied upon or where 

containerized plantings or plugs will be installed. Post-construction monitoring 

protocols should also be part of any restoration plan (Brooks and others 2006).

Since hydrologic regime is such an important factor in the functioning of fens, 

every effort must be made to accurately identify all potential water sources and to 

explicitly formulate desired hydrologic outcomes. Hydrologic processes should 

be addressed at the earliest stages of project development, as the ultimate success or failure of a project will depend on successfully restoring water sources. 

Often, hydrologic restoration can be accomplished once the extent and nature 

of hydrologic modifications are known. For example, wetland hydrology can 

be restored to drained fens by plugging ditches to return the water table to its 

original elevation using soils or sheet metal check dams (Figure 43) (Cooper and 

others 1998; Patterson 2005). Because organic soils typically subside after drainage, caution should be exercised when planting as water tables may rise above 

the soil surface, killing newly planted vegetation (Ewing and Vepraskas 2006).

If disturbed fens retain suitable hydrologic regimes following disturbance, 

asexual propagules present in the fen may allow for some degree of recovery. 

For example, Johnson (2000) found that in areas with intact wetland hydrologic 

regimes, seeds played a minor role in revegetation, and that most new growth 

originated from vegetative structures. However, where significant drainage has 

occurred, sexual reproduction was found to be the dominant mode of plant establishment. In another experiment, of eight species seeded, only one, Triglochin 

maritima, germinated and established seedlings (Cooper and MacDonald 2000). 
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Seedlings, rhizome transplants, and stem cuttings of woody species such as willows are all potential approaches that should be evaluated for possible inclusion 

in revegetation plans, as each differs with respect to likely survival patterns in 

relation to hydrologic regimes (Cooper and MacDonald 2000).

Seedbanks are likely important following disturbance. Many graminoid species 

that are common in fens exhibit delayed germination, allowing for the formation of soil seedbanks (Leck and Schutz 2005; Schütz 1998; van den Broek and 

Beltman 2006). Some research also suggests that the spores of some bryophytes 

may remain viable for several years. Sunderg and Rydin (2000) found that under 

suitable conditions, the spores of some Sphagnum species can form a persistent 

spore bank with a half-life of between 1 and 20 years and with individual spore 

capsules capable of persisting for several decades or even centuries (Sundberg 

and Rydin 2000). Several soil seedbank studies have shown that woody plants 

may be absent or poorly represented in soil seed banks (McGraw 1987; Patterson 

2005; Rossell and Wells 1999).

Transplantation of soils from relatively pristine fens to heavily disturbed sites 

may promote colonization of desired species. In a Colorado fen, replacement of 

the upper 10 cm of native soil significantly speeded revegetation by supplying 

viable sexual and asexual propagules (Johnson 2003). It may be possible to reestablish some fen species, even if the complete loss of the peat body occurs. For 

example, Amon and others (2005) documented the production of peat composed 

of brown-mosses within three years of revegetation of gravel substrates on top 

of a buried fen. The authors also reported that naturally derived plugs were established most successfully while plants that established from seed produced the 

fewest survivors (Amon and others 2005). Where fens have been buried, the fill 

material should be removed and the sites should be revegetated using plantings 

grown in green houses from locally collected seed. For woody plants like willows, stem cuttings can be planted. This approach has been used successfully for 

large-scale fen restoration near Telluride, Colorado (Cooper, unpublished data).

Figure 43—Ditch in Big Meadows, Rocky Mountain National Park in 1989 (left panel), prior to hydrologic restoration by National Park Service personnel using sheet metal check dams (center panel). The same area in 

2006 (right panel) (photos: D. Cooper).
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Direct seeding of wetland plant species for fen restoration has generally produced 

poor results (Cooper and MacDonald 2000); however, because direct seeding 

is relatively inexpensive, it continues to be used for many restoration projects. 

The germination requirement of seeds are species specific (Isselstein and others 

2002; Tallowin and Smith 2001), leading to unpredictable results from seeding. 

Irrigation may facilitate germination and seedling establishment, while shading 

inhibits germination. Higher soil temperatures may improve germination rates. 

To ensure genotypes are well adapted to local site conditions, seeds should be 

collected from the area where restoration will occur. All seed should be cleaned 

and screened to remove the non-seed materials like fruiting bracts, floral parts, 

awns, scales, and perigynia (Dunne and others 1998). Many seeds require stratification under wet and cold conditions or temperature fluctuations mimicking 

natural diurnal rhythms to break dormancy. Seed stratification can be done in the 

lab or by dispersing seeds in the fall and allowing them to overwinter in the field.

Because of the sensitivity of peat soils, broadcast seeding is generally the 

preferred method of seeding. Hydroseeding is not appropriate, as the mulch, 

binder, and seed/seedlings may be washed away prior to germination and seedling establishment (Colorado Natural Areas Program 1998). Although the best 

depth to incorporate seeds varies by species, general guidelines are to plant to a 

minimum depth of 0.7 cm and a maximum depth of 1.3 cm for most small seed 

mixes, while a minimum depth of 1.5 cm and a maximum depth of 2.6 cm may 

be more appropriate for mixes containing larger seeds (Dunne and others 1998). 

Natural seed rain cannot generally be relied upon since many fen species are not 

capable of seed dispersal over large distances, particularly where surface water 

flow is interrupted (Bakker and others 1996). Some species, such as willows, 

can be effectively established from natural seed rain if appropriate water table 

depths are present (Cooper and Andrus 1994).

Plugs and bare root stock are other possible techniques that can be used in fen 

restoration (Pfadenhauer and Grootjans 1999). For example, Kobresia simpliciuscula, Carex utriculata, and other species were successfully established at 

a peat mined fen in Colorado from rhizomes collected from undisturbed areas 

(Cooper and MacDonald 2000). However, this approach is time consuming and 

is only suggested for relatively small restoration areas. Plants grown from fieldcollected seed in greenhouses have been successfully used in many restoration 

projects. This approach may be appropriate for Carex aquatilis, C. utriculata, 

Calamagrostis canadensis, Carex vesicaria, and Scirpus pungens, among others. 

This approach can be expensive relative to broadcast seeding but typically results 

in much higher plant survival. Once plants are established, they can provide localized centers of seeds for future plant establishment, as well as vigorously growing 

rhizomes that can spread over large areas. Planting densities of 2 to 4 plants/m2

have been used, and for rhizomatous species, a planting density of even 1 plant/

m2 or lower could be used to introduce the desired species over large areas or to 

supplement other plant propagation approaches such as direct seeding.
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Management constraints

Constraints on the management of fens can be divided into several categories. 

Resource constraints include access to funding to develop management and restoration goals. It is often useful to identify several likely funding scenarios when 

developing management objectives to help keep focus on realistic restoration 

options and to identify at what point attainment of project goals are not possible 

if resources are limiting. Political constraints are an additional factor that should 

be considered. It is usually best to identify and engage all potential stakeholders 

early on in planning, as this may prevent problems further down the road.

An external constraint that may have significant effects on fens that is outside 

the control of managers is regional climate changes. Where managers rely on 

harsh winters to control overabundance of elk populations, moderate climatic 

conditions predicted under some regional climate models may allow significant 

increases in elk herds, thereby increasing impacts to fens (Wang and others 2002). 

Surface temperature increases and the resulting increase in evapotranspiration 

rates may lead to lower water tables in fens, altering peat accumulation processes 

and species composition (Hogg and others 1992). Since soil temperature is a 

major factor affecting organic matter decomposition, warming may accelerate 

decomposition processes (Domisch and others 2006). Models of climate change 

indicate that temperatures will increase and precipitation seasonality will shift 

over the coming decades in the Rocky Mountains, leading to earlier spring runoff, lower summer and fall flows, decreased snowpack, and an increase in the 

proportion of precipitation falling as rain versus snow (Cooper and others 2006; 

Knowles and others 2006).

Increased ultraviolet-B (UV-B) radiation caused by reductions in the amount of 

stratospheric ozone could affect fens via effects on plant carbon allocation and 

root exudation, which in turn, could impact microbial communities. For example, 

research by Rinnan and others (2006) using experimental microcosms found 

that when Eriophorum angustifolium, a widespread fen species, was exposed 

to enhanced UV-B radiation simulating a 15 percent reduction in ozone, there 

was an increase in the concentration of root exudates in the rhizosphere and in 

the ratio of roots to shoots. Such responses may alter the belowground biomass 

distribution of the fen plants, leading to changes in the net efflux of root exudates 

(Rinnan and others 2006) and altered peat accumulation rates.
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7. Historical Range of Variation for 

Region 2 Salt Flats

Definitions and Concepts of Salt Flats ____________

Salt flats are wetlands characterized by high salt concentrations. As with other 

wetland types, the terminology used to describe these ecosystems is often ambiguous or imprecise, with different terms applied regionally. Other terms used 

to describe salt-affected wetlands are inland salt marsh, salt pan, sabhka, and 

playa. The key characteristic defining the wetlands discussed in this chapter is 

the presence of high soil salt concentrations. As described in subsequent sections, highly saline soils impose a significant environmental stress on plants and 

relatively few plant species are able to tolerate these conditions. Generically 

referred to as halophytes, plant species tolerant of highly saline environments 

form distinct plant communities and are important for delineating salt flats in 

the field. Some terms, like playa, also include non-saline wetlands and are not 

considered salt flats in this report. As with all wetlands, hydrologic regime is 

a key variable controlling the structure and function of salt flats; however, it is 

the presence of high salt concentrations that most clearly distinguishes salt flats 

from other ecosystem types.

Geomorphic Setting and Principal 

Ecological Drivers _____________________________

Geomorphic setting

Salt flats most commonly occur in arid and semi-arid landscapes where evapotranspiration exceeds precipitation throughout the year (Aschenbach 2006). 

However, saline wetlands are also abundant in more humid regions such as the 

prairie pothole region of the northern Great Plains (Last and Ginn 2005; Lieffers 

1984). There are no reliable statistics describing the contemporary distribution 

of salt flats in Region 2. Estimates of salt affected soils have been developed 

within individual states and maps showing the distribution of saline soils may be 

found on some county soil surveys. More than 400,000 ha of soils are classified 

as salt affected in Kansas (Aschenbach 2006). Salt flats are generally restricted 

to the Great Plains, although saline ecosystems also occur at high elevations in 

intermountain basins. For example, Ungar described halophytic communities occurring in South Park, Colorado, at elevations greater than 3000 m (Ungar 1974b).
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A variety of factors influence the accumulation of salts in soils and salt flat 

ecosystems can occur in several geomorphic settings. Within Region 2, salt 

flats most commonly occur associated with terminal sump basins or along the 

margins of ponds or lakes where evaporation may lead to the accumulation of 

high soluble salt concentrations (Ungar 1974b). Salt mobilization from edges can 

be rapid and influenced by lateral groundwater flow through surface sediments 

(Arndt and Richardson 1993). Halophytic communities also occur in association with discrete saline springs (Bolen 1964; Joeckel and Clement 1999, 2005). 

Anthropogenic features such as reservoirs or ditches also support salt flats, often 

in landscape settings that did not historically support such communities (Peck 

and Lovvorn 2001; Skarie and others 1986).

Key drivers and ecological processes

Climate—Climate is a primary driver of wetland formation and wetland functional characteristics, although the variables of greatest importance vary among 

wetland types. Variables of particular importance to salt flats are temperature and 

annual precipitation, particularly as they influence processes of evapotranspiration. The dominant form of precipitation at lower elevations, where the majority 

of salt flats occur, is rain. The frequency and magnitude of precipitation events is 

highly variable spatially and temporally, resulting in cyclical patterns of relative 

wetness and dryness. Most commonly studied in relation to freshwater marshes, 

these cycles also influence the function of salt flats (Kantrud and others 1989a; 

van der Valk and Davis 1978).

Climate can affect salt flat formation and function in several ways. The main 

hydrological input in many salt flats, particularly playa-type ecosystems, is direct 

precipitation and surface runoff from surrounding areas. Because of their reliance 

on precipitation, inter-annual and decadal fluctuations in precipitation are major 

drivers of hydrologic function and vegetation dynamics in these ecosystems.

At broad temporal scales, climate influences geomorphic processes. For example, 

the formation of landforms such as dunes appears to have been episodic, driven 

by shifts towards drier and warmer climatic conditions (Forman and others 2001; 

Muhs and Holliday 1995). The expansion and retreat of glaciers, the key process 

forming many saline wetlands in the northern Great Plains outside of Region 

2, was driven by large shifts in precipitation and temperature (Krimmel 2002; 

Munroe 2003; Munroe and Mickelson 2002).

Salt accumulation may affect microclimatic processes, including site evapotranspiration rates. A study in a playa found that high osmotic pressure and the presence of salt crusts caused most absorbed radiation to be partitioned to sensible 

heat (Malek and others 1990). Higher air temperatures above the wetland raised 

potential evapotranspiration (ET), and while evaporation rates of the two sites 

were similar following rainfall events, playa ET rates were quickly reduced as 

the osmotic potential increased.
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Geology—At both local and regional scales, geologic setting is a primary driver 

of salt flat accumulation processes and salt flat functioning. Excluding anthropogenic sources, soluble salts can originate from either marine or lithogenic sources, 

with the importance of each varying among sites (Waisel 1972). Because of the 

region’s continental setting, marine salt sources such as windborne sea spray 

are not factors influencing the formation of Region 2 salt flats, although fossil 

marine salt deposits do occur in some locations. Bedrock salt sources, created 

from the weathering of rock, are the most important source of salts in Region 2. 

The amount of soluble salts and the dominant ionic species present (e.g., sodium, 

calcium, sulfate, and chloride) can vary widely among different rock types. For 

example, sandstones typically contain little chloride, while limestone contains 

high amounts of chloride and sulfur relative to most shales (Waisel 1972).

The composition of parent materials in watersheds and aquifers supporting salt 

flats influences processes of salt accumulation. Saline soils are typically associated with sedimentary rock formations. For example, in Colorado, soil and water 

in areas formed from Mancos shale on the Colorado Plateau are often highly 

saline (Wagenet and Jurinak 1978; Whittig and others 1982). Because of low ion 

concentrations and slow weathering, ecosystems with high salt concentrations 

are rare in landscapes dominated by igneous or metamorphic rocks.

Parent material also influences salt flat formation through effects on local and 

regional hydrologic processes. Processes such as groundwater recharge and 

discharge vary as a function of the transmissivity of bedrock layers and the 

arrangement of geological strata. For example, salt marshes and springs are 

common in river valleys of central Kansas and are usually associated with the 

Dakota Formation, which can include permeable sandstones and less permeable 

siltstones and shales with high storage. Because of variation in the configuration 

of strata, hydraulic conductivity of the underlying aquifer ranges from less than 

10 ft/day to more than 50 ft/day (Macfarlane and others 1989).

Eolian processes can influence the form and function of salt flats and wind erosion has influenced the formation of playas, many of which support halophytic 

communities (Hovorka 1995). Salt crusts formed in salt-affected soils are easily 

mobilized and transported by wind (Blank and others 1999) and have been shown 

to influence solute concentration in groundwater in arid and semi-arid areas. On 

the southern high plains, approximately 4.5 x 105 kg of chloride is estimated to 

be removed annually from the basin floor of a relatively small (4.7 km2) saline 

lake (Wood and Sanford 1995). The transport of dust is temporally variable and 

influenced by seasonal patterns of soil moisture and wind intensity. Dust transport 

is typically limited to brief periods, separated by long periods of inactivity, and 

peaks during winter months when winds are moderately strong, precipitation is 

at a minimum, and vegetation is senescent (Stout 2003).
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Glaciation is a major factor influencing the formation of saline wetlands in some 

regions such as the prairie pothole region of the northern Great Plains (Dodd and 

Coupland 1966; Keller and Vanderkamp 1988), and while continental glaciers 

did reach parts of present-day South Dakota, no USFS lands in Region 2 were 

overlain by continental ice sheets. Pleistocene valley glaciers did affect many 

high mountain areas in the region, an important factor in the formation of some 

wetland types such as fens. However, with the exception of some intermountain 

valleys, saline wetlands rarely occur outside of the Great Plains because geological and climatic factors are not conducive to their formation.

Hydrologic regime—Hydrologic regime is the key factor controlling vegetation 

composition and soil development in all wetlands, including salt flats. Because 

high salinity is the primary feature used to differentiate salt flats from other wetland types, there is considerable variability in hydrologic characteristics among 

them. Some salt flats have hydrologic regimes similar to marshes. Their water 

levels may fluctuate widely both seasonally and among years, with extended 

periods of deep inundation alternating with periods where surface water may be 

completely absent. Many salt flats are associated with lacustrine features such 

as lakes, ponds, and reservoirs with widely fluctuating water levels.

Sources of water to salt flats are precipitation, surface water, and groundwater. 

The importance of each can vary among wetlands and temporally within individual ecosystems. The main hydrologic inputs to salt flats occurring in closed 

basins are precipitation and surface runoff (Cooper and Severn 1992). Salt flats 

may also be influenced by the groundwater, either through direct discharge in 

discrete springs or through capillary movement of water from seasonally high 

water tables (Bolen 1964; Joeckel and Clement 2005; Riley 2001).

High-velocity surface flows resulting from precipitation runoff may affect salt 

flats but are less important to geomorphic and ecological processes than in riparian 

ecosystems. In many wetlands, spatial and temporal patterns of salt accumulation 

in near-surface groundwater and surface water are dynamic and closely linked 

with discrete recharge events (Arndt and Richardson 1993). The frequency and 

magnitude of water table fluctuations are strong drivers of processes of salt accumulation and removal from soils, and are influenced by the amount and timing 

of precipitation events and by the presence of groundwater.

Anthropogenic factors, particularly irrigated agriculture, are responsible for 

salination in many areas of the semi-arid West. Stock ponds and other water 

storage features across Region 2 have likely led to an increase in lacustrine and 

lacustrine fringe wetlands in much of the region (Figure 44a). Saline seeps may 

form where slopes naturally break and water tables are near the surface. Contact 

salinity occurs where a permeable water-bearing surface layer that is thinning 

out above a less permeable layer (e.g., clay) forces groundwater flow close to the 

soil surface. Farming practices can allow water to move out of the root zone into 

the subsoil, dissolving and accumulating salts in transit, and salt flats commonly 
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Figure 44—Schematic illustrations of different hydrogeomorphic settings for salt flats. Lacustrine 

fringe creates salt accumulation where ponds dry seasonally or where the surface water supported 

water table is close to the soil surface (A), break in slope (B) where groundwater discharges to 

near the soil surface, downstream of a ditch (C) where groundwater reaches to near the soil 

surface, and perched water table (D) where capillary water reaches the soil surface (modified 

from Alberta Agriculture and Food 2004).

occur adjacent to irrigation ditches (Figure 44b, c). Salt flats may also form where 

layers of low permeability create a perched water table, with evaporation leading 

to salt precipitation. For example, saline seeps are more common in western than 

eastern South Dakota in part because much of the subsoil consists of bedded 

shale material with low permeability (Millar 2003) (Figure 44d).

The chemical composition of aquifers can vary. In Canada, researchers documented several subsurface water types with distinctive ionic composition and 

salt concentration (Last and Ginn 2005). Fluctuations in water table depth and 

precipitation events create temporal variation in salt accumulation (Riley 2001). 

There are large differences in the hydrologic functioning of saline wetland types. 

For example, while wetlands in the San Luis Valley of Colorado share many attributes with playas of the Southern High Plains or central California (Bolen and 

others 1989; Fort and Richards 1998; Smith 2003), they are not hydrologically 

isolated and precipitation fed; rather, they are connected to a large subsurface 

aquifer (Groeneveld and Or 1994).
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Water and soil chemistry—Considerable variation may exist in the geochemistry of salt flats. The most common salts are sulfates, carbonates, bicarbonates, 

and chlorides of calcium, magnesium, sodium, and potassium (Table 23). The 

concentration and composition of the dominant ions in soil and water varies spatially among and within individual wetlands, as well as over time (Last 1989a). 

As salt flats dry, salt crusts develop on the soil surface. Specific evaporites that 

are formed along the margin and bottom of salt flats may include mirabilite, thenardite, and bloedite (Last 1989a). Sodium sulfate occurs as thenardite (Na2SO4) 

above 32 °C and mirabilite (Na2SO4 • 10H2O) at cooler temperatures. Volume 

increases and solubility decreases when thenardite changes to mirabilite, often 

creating salt heaving (USDA NRCS 1999).

Table 23—Salt compounds and their constituent cations/anions found in 

salt flats. 

Salt compound Cation (+) Anion (-) Common name

NaCl sodium chloride halite (table salt)

Na2SO4 sodium sulfate Glauber’s salt

MgSO4 magnesium sulfate epsom salts

NaHCO3 sodium bicarbonate baking soda

Na2CO3 sodium carbonate sal soda

CaSO4 calcium sulfate Gypsum

CaCO3 calcium carbonate calcite (lime)

A standard measure of salinity used by the NRCS in soil taxonomy is electrical conductivity (EC), generally measured in decisiemens per meter (dS/m) or 

Millimhos per centimeter (mmhos/cm) (U.S. Department of Agriculture 1999). 

The NRCS recognizes five salinity classes based on EC, including non-saline 

(0 to 2 mmhos/cm), very slightly saline (2 to 4 mmhos/cm), slightly saline (4 to 

8mmhos/cm), moderately saline (8 to 16 mmhos/cm), and strongly saline (greater 

than 16 mmhos/cm) (USDA NRCS 1999). The concentration of ions can vary 

widely in relation to local and regional geological and hydrological factors.

The main classifiers of soils in salt flats are salinity, alkalinity, and sodicity. 

Salinity refers to the concentrations of both total soluble salts and exchangeable 

sodium, which is usually measured by EC (Shainberg 1975; Ungar 1974b), 

although most studies do not explicitly consider alkalinity or sodicity. Sodicity 

refers specifically to soil Na+ abundance. Sodium adsorption ratio (SAR)—calculated from the concentrations of sodium, calcium, and magnesium in a saturated 

extract—is typically used to measure the sodicity of a soil. The sodium adsorption ratio is calculated from the concentrations (in milliequivalents per liter) of 

sodium, calcium, and magnesium in the saturation extract (USDA NRCS 1999):
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SAR = (Na+)/ sqrt((Ca2++Mg2+)/2)

High exchangeable sodium occurs in sodic soils and has a significant influence 

on the physical and chemical properties of a soil. The presence of large quantities of Na+ ions leads to loss of soil structure because of the weak attraction of 

Na+ ions to the soil colloids allows for clay dispersion, which clogs soil pores, 

decreasing hydraulic conductivity and infiltration. Sodic soils have high levels 

of exchangeable sodium but low levels of total salts.

Sodic soils may impact plant growth by specific sodium toxicity, nutrient deficiencies or imbalances, high pH, and soil structure changes resulting from dispersion 

of clay particles. Sodic soils may be hard when dry, limiting water intake and 

plant growth is typically poor (Davis and others 2003). Sodium problems are most 

pronounced on soils with high clay contents, particularly where clays with high 

shrink-swell potential are present, such as in the basins and plains of Wyoming.

The pH in salt flats may range from approximately 7.0 to over 10.0 and may be 

barren of vegetation in the most alkaline sites (Winters and others 2005). Soils in 

arid and semiarid areas may become alkaline if there is insufficient precipitation 

to leach cations (e.g., Ca2+ and Mg2+) from soils. In alkaline sites, carbonic acid 

(H2CO3) that is formed from microbial and root respiration may form bicarbonate and carbonate ions. The pH of soil and water may reduce the availability of 

essential elements necessary for plant growth. Typically, plants growing in soils 

with high pH have reduced stomatal conductance and leaf nutrient concentrations 

of iron, calcium, manganese, and zinc (Brady 1996; Haukos and Smith 1996).

Disturbance processes—The role of disturbance in salt flat dynamics is largely 

unstudied in the region, although disturbance is likely a major environmental 

gradient affecting salt flat vegetation (Walker and Wehrhahn 1971). Disturbance 

in salt flats may include periodic high water periods, which can drown plants, 

as well as multi-year droughts, which lead to high plant mortality. As wetlands 

dry, evaporation increases salt concentrations, which eventually exceed plant 

tolerances. Halophytes in coastal salt marshes often exploit patches created by 

disturbance and are restricted to salt-stressed microsites because of their inability 

to compete with turf-forming perennial species (Bertness 1992; Shumway and 

Bertness 1992); however, it is unknown whether similar processes are important 

in inland ecosystems.

Disturbances from animals may be locally important to salt flats. Livestock and 

native ungulates can trample salt flat plants, causing plant mortality. Birds, particularly geese, can be disturbance agents in salt marshes. For example, geese 

can destroy salt marsh plants and expose bare sediments by grubbing for grass 

roots and rhizomes in the spring. Increased evaporation later in summer from 

disturbed sediments can cause increased soil salinities that adversely affect remaining plants (Srivastava and Jefferies 1996).
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Fire can also influence salt flats, although the specific importance of fire probably varies widely. Because salt flats are relatively discreet and isolated, ignitions within salt flats are probably very rare and any fires affecting them likely 

originate in surrounding ecosystems. The incidence of fire varies as a function 

of the fire regimes of surrounding community types. Smith and Kadlec (1985) 

found in a study of Utah marshes that vegetation regrowth was rapid following 

fire in communities dominated by Typha spp, Scirpus lacustris, and S. maritimus

but was very limited in more saline sites dominated by Distichlis spicata (Smith 

and Kadlec 1985).

Salt flat soils with high clay contents can shrink and swell, forming large crack 

networks (Hovorka 1995). Salt efflorescence and precipitation features may be 

formed, which can create complex microtopography. Goodall and others (2000) 

described thick crusts, with high relief (>10 cm) and a blocky morphology and 

thin crusts, with low relief and a blister-like appearance. Working in the northern 

Great Plains, Last (1989b) described four basic types of salt occurrences: surface 

crusts and hardgrounds, massive and bedded salts, deposits associated with springs, 

and subsurface and groundwater-related accumulations. Soil deformation and 

crust formation may kill seedlings but it can also create microsites that facilitate 

seed entrapment and germination (Fort and Richards 1998).

Salt flat development and succession—Because of the extreme environment, 

salt flats are species poor—limited to halophytes. Succession in saline ecosystems may occur slowly because of a limited species pool of adapted species and 

may consist of successive species replacements with the same or similar species. Where hydrology and salinity are stable, communities may reach a sort of 

edaphic climax; however, these characteristics often vary widely.

Over broad time scales, primary successional patterns may be driven by longterm changes in water levels and flood frequency. The margins of saline lakes, 

for example, may experience a cyclical pattern of species invasion and retrogression. Long-term evolution of salt flats varies based on climatic and geological 

factors. Wetlands in the northern Great Plains are young, originating within the 

last 10,000 years, in contrast to the much older playas of southwestern United 

States (Last 1989b).

Seasonal fluctuations in soil salinity are an important factor driving vegetation 

dynamics, and life stages of a species may differ strikingly in their salinity tolerance. Many species will only germinate during periods of lower salinity (Riley 

2001), although some species (e.g., Hordeum jubatum) will germinate at salinities higher than those considered favorable to growth and reproduction (Badger 

and Ungar 1989; Egan and Ungar 1999).
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Salt Flat Classification and Gradients _____________

Under the hierarchical classification scheme used by the National Wetlands 

Inventory program, salt flats are placed into either the Lacustrine or Palustrine 

system (Cowardin and others 1979). The Lacustrine system includes wetlands 

situated in a topographic depression or dammed river channel, that exceed 8 ha 

in area, and that have less than 30 percent total cover from trees, shrubs, or persistent emergent vegetation. Also included are wetland and deepwater habitats 

less than 8 ha in area that have active wave-formed or bedrock shoreline features 

along their boundary, features not generally found in Region 2. Palustrine wetlands are less than 8 ha in size, have more than 30 percent vegetation cover, lack 

active wave-formed or bedrock shoreline features, and have water levels less 

than 2 m at their deepest location (Cowardin and others 1979). The majority of 

salt flats would be placed into the Palustrine system using the Cowardin (1979) 

classification scheme.

Under the Cowardin (1979) classification, a variety of modifiers can be applied 

to wetlands. Modifiers include water regime, soil type, and water chemistry 

parameters such as pH and salinity. Among non-tidal wetlands, water regimes 

are defined in terms of the frost-free period and include, from wettest to driest, 

the following categories: permanently flooded, intermittently exposed, semipermanently flooded, seasonally flooded, saturated, temporarily flooded, and 

intermittently flooded. Because our definition of salt flats in this assessment 

is based on salinity rather than hydrologic characteristics, salt flat examples of 

many of the Cowardin water regime types can be found, although most salt flats 

would fall under the saturated, temporarily flooded, and intermittently flooded 

categories. The Cowardin classification system also uses the artificially flooded 

modifier to describe wetlands in which the amount and duration of flooding is 

controlled by means of pumps or siphons in combination with dikes or dams, 

although wetlands within or resulting from leakage from man-made impoundments and irrigated pasture lands supplied by diversion ditches or artesian wells 

are not included (Cowardin and others 1979).

The Cowardin (1979) system also includes water chemistry modifiers related to 

salinity, which range from fresh to hypersaline (Table 24). As we define them 

in this assessment, salt flats include wetlands classified from oligosaline to hypersaline under the Cowardin system. Notably, salinity in an individual wetland 

can vary widely over time and with the degree of flooding.

Under the HGM approach (Brinson 1993a; Cooper 1998a), salt flats can be 

described as mineral flats, depressional, or slope wetlands. The original framework was intended to be augmented by publication of regional guides; however, 

relatively few of these have been prepared (Hauer and others 2002a). None deals 

with the entire range of saline wetlands, although some types are covered in 

existing guides (Stutheit and others 2004).
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Table 24—Salinity modifiers used in the National Wetland Inventory program (Cowardin 

and others 1979).

Salinity Approximate specific conductance

Inland modifiers (parts per thousand) (μMhos at 25° C)

Hypersaline >40 >60,000

Eusaline 30.0-40 45,000-60,000

Mixosaline 0.5-30 800-45,000

Polysaline 18.0-30 30,000-45,000

Mesosaline 5.0-18 8,000-30,000

Oligosaline 0.5-5 800-8,000

Fresh <0.5 <800

Several subclasses relevant to salt flats were identified in a classification of 

Colorado wetlands using the HGM approach (Cooper 1998): subclass D2—encompassing low-elevation basins with permanently or semi-permanently flooded 

hydrologic regimes such as reservoir and lake margins and marshes; subclass 

D3—low-elevation basins acting as sumps for water moving from mountains or 

as depressions in irrigated slopes; subclass D4—low-elevation sites with temporarily flooded hydrologic regimes and poorly developed vegetation; and subclass 

D5—Intermittently flooded low elevation basins that are not flooded annually 

or are largely barren of vegetation. The F1 subclass include mineral soil flats at 

low-elevation (Table 25) (Carsey and others 2003; Cooper 1998a). In practice, 

ordination data used to delineate HGM subclasses can be problematic as some 

species occur in multiple categories. Distinguishing between slope and mineral 

flat wetlands can be particularly difficult.

Table 25—Primary HGM subclasses described by Cooper (1998) relevant to salt flats.

HGM

Subclass Description Common species

Depressional 4 Temporarily flooded low-elevation basins Polygonum lapathifolium

flooded for short periods in the spring and

early summer.

Depressional 5 Intermittently flooded low-elevation basins Xanthium strumarium

that are not flooded annually or are largely

barren of vegetation.

Flats 1 Middle to low-elevation sites on mineral saline Suaeda calceoliformis,

soil (due to evaporation) with a seasonal high Puccinellia nuttalliana,

water table near the ground surface and Sarcobatus vermiculatus

occasionally shallow standing water. 

Slope 2 Subalpine and montane fens and wet Eleocharis quinqueflora,

meadows on saturated calcareous Kobresia simpliciuscula,

substrates. Carex simulata

Slope 4 Low-elevation meadows with a seasonal Hordeum jubatum

high water table near the ground surface. 

May occur on floodplains or near springs.
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Under the National Vegetation Classification System used by Natural Heritage 

programs, salt flats may be mapped or classified under the InterMountain Basins 

Playa, Inter-Mountain Basins Mixed Salt Desert Scrub, Eastern Great Plains 

Wet Meadow Prairie and Marsh, or North American Arid West Emergent Marsh 

ecological system (Comer and others 2003b). A variety of associations have been 

described, with regional variation in specific species depending on the available 

species pool (Table 26).

Salt Flat Vegetation in the Region ________________

Salt flats typically have sparse plant cover, due to high pH and salt concentrations, 

although wetlands may support high aquatic invertebrate and algae production 

when inundated (Winters and others 2005). Soil moisture and salt concentration 

gradients are the primary edaphic factors affecting plant distribution in saline 

areas (Rogel and others 2001a, 2001b), although other factors affect species 

composition and distribution in salt flats, including hydrologic regime, plant 

competition, pH, nutrient status, and seed bank response to wet and dry cycles 

(Dix and Smeins 1967; Stewart and Kantrud 1972; van der Valk and Davis 1978).

Table 26—Primary vegetation associations related to salt flats in Colorado (Carsey and others 2003).

Element code Scientific name Common name

Mineral flats 1

CEGL001770 Distichlis spicata Herbaceous Vegetation Inland saltgrass Herbaceous Vegetation 

CEGL001779 Muhlenbergia asperifolia Herbaceous Alkali muhly Herbaceous Vegetation

Vegetation 

CEGL001799 Puccinellia nuttalliana (=airoides) Nuttall’s alkaligrass Herbaceous Vegetation

Herbaceous Vegetation 

Sarcobatus vermiculatus/Barren ground Black greasewood/Barren ground Shrubland

Shrubland

CEGL001363 Sarcobatus vermiculatus/Distichlis spicata Black greasewood/Inland saltgrass Shrubland

Shrubland 

CEGL001843 Schoenoplectus maritimus (=Bolboschoenus Cosmopolitan bulrush Herbaceous

maritimus) Herbaceous Vegetation Vegetation 

Scirpus nevadensis (=Amphiscirpus Nevada bulrush Herbaceous Vegetation

nevadensis) Herbaceous Vegetation 

CEGL001588 Spartina gracilis Herbaceous Vegetation Alkali cordgrass Herbaceous Vegetation 

CEGL001685 Sporobolus airoides Southern Plains Alkali sacaton Southern Plains Herbaceous

Herbaceous Vegetation Vegetation 

Suaeda calceoliformis Herbaceous Pursh seepweed Herbaceous Vegetation

Vegetation

Triglochin maritimum-Triglochin palustris Seaside arrowgrass-Meadow arrowgrass

Herbaceous Vegetation Herbaceous Vegetation 

Depressional 4/5

CEGL001798 Hordeum (=Critesion) jubatum Herbaceous Foxtail barley Herbaceous Vegetation

Vegetation 
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Plants have a range of physiological and life history adaptations to saline environments and vary in their tolerance to these conditions (Glenn and others 

1999). The main physiological stresses experienced by plants in saline soils are 

related to alteration of their osmotic balance. Plants accumulate organic acids 

intracellularly to compensate for high levels of external soluble salts (Shainberg 

1975; Ungar 1974a, 1974c). Effects can include reduced transpiration rates and 

plant water availability, while excessive ion concentrations may reduce uptake 

of essential mineral nutrients, constraining plant growth (Shainberg 1975). Other 

adaptations include: lower selectivity against leaf sodium uptake; the ability to 

isolate sodium from metabolically active tissues; and among some perennial 

species, accumulation of salt in older leaves, which are shed at the end of each 

year (Shainberg 1975; Ungar 1974c). High conductivities affect stomatal conductance and water use, thereby affecting photosynthetic rate. Affected plants 

may exhibit a variety of phenotypic responses, including differences in leaf 

morphology and branching pattern. Additionally, the ramets of some clonal salt 

flat species may be physiologically integrated. Vegetatively reproducing species 

such as Juncus arcticus, Salicornia rubra, and Triglochin spp. may be able to 

out compete disturbance-dependent annuals (Riley 2001).

The germination of seed for many halophytes occurs in response to seasonal 

fluctuations in salinity levels. Seeds that germinate early may be more resistant 

to increased salinity later in the season because of greater stored food reserves 

(Foderaro and Ungar 1997). Persistent soil seed banks allow for rapid colonization when more favorable salinities are present and are considered important in 

maintaining populations of halophytes in many wetlands (Badger and Ungar 

1994; Egan and Ungar 2000; Gul and Weber 2001). For example, Distichlis 

spicata responds to short-term salt marsh disturbance with rapid colonization 

of bare patches from a persistent seed bank and vegetative spread (Bertness and 

Ellison 1987).

Obligate halophytes are plants with optimal growth at moderate or high salinity 

that are incapable of growth at low salinity. However, true obligate halophytes 

are rare, and many species are facultative (Barbour 1970). Many halophytes are 

in the families Chenopodiaceae and Asteraceae, although grasses (Poaceae) are 

also common.

Species diversity is typically low in salt flats (Riley 2001; Rocchio 2006a; Windell 

and others 1986), although distinct patterns of plant zonation often occur. In 

Colorado’s San Luis Valley, wide hydrologic and salinity gradients occur over 

an elevation range of less than 2.5 m (Cooper and Severn 1992). Pondweed and 

other aquatic species dominate playas that are regularly flooded, while the margins 

of wetlands or mineral flat sites may be dominated by succulent dicots such as 

Suaeda calceoliformisa and Salicornia spp. Monocots may include Puccinellia 

airoides, Spartina gracilis, Hordeum jubatum, Triglochin ssp., Sporobolus airoides, 

Muhlenbergia asperifolia, and Distichlis spicata (Redmann 1972). Salt-tolerant, 

phreatophytic shrubs such as Sarcobatus vermiculatus commonly occur adjacent 

to wetter communities (Cooper and others 2006; Rocchio 2004, 2006a).
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HRV of Salt Flats in Region 2 ____________________

Overview of anthropogenic impacts

Because plant cover and productivity are low in salt flats and only salt-tolerant 

taxa occur (Dodd and Coupland 1966; Ungar 1966), salt flats are rarely utilized 

by people. As a result, many salt flats have been spared anthropogenic impacts 

affecting other wetland types. Hydrologic modifications can result from both 

on-site and off-site impacts. Some impacts, such as livestock trampling, may be 

locally severe, but probably affect a relatively small number of sites, while others 

may be more widespread and subtle in their ecological effects. Unlike fens, no 

broad-scale analyses have been conducted to examine the relative importance 

of different impacts on salt flats in the region.

Mineral extraction—There are documented examples of historical and contemporary human exploitation of the salt accumulated in saline ecosystems. “Salt 

licks” in the eastern half of the United States were a significant local to regional 

source of commercial salt well into the Nineteenth Century. Salt springs and efflorescences in Nebraska were known to Euro-Americans as early as 1835, and 

by the late 1850s, they had become the main impetus for the settlement of the 

area (Joeckel and Clement 1999). That these operations failed within 30 years 

suggests that overall levels of exploitation are not great in the region. Similar 

“salt works” occur in South Park, Colorado, just southwest of Antero Reservoir. 

South Park was called Bayou Salado by early explorers due to the broad expanses 

of salt on the soil (Simmons 1992).

Oil and gas extraction—Soil salinization can occur associated with oil and gas 

production since salt water, consisting primarily of sodium chloride (NaCl) is 

often encountered during drilling operations. While water may be reinjected into 

disposal wells or deposited in on-site reserve pits, contamination can occur from 

accidental spills or lateral subsurface movement of reserve pit brine (Aschenbach 

2006). Contamination from brine spills can result in the death of vegetation and 

reduced forage production on rangelands or crop yields on agricultural lands. 

Common halophytes occurring in salt flats can be used in bioremediation of soils 

contaminated by brine spills. For example, one study found that nearly all species 

evaluated reduced soil salinity significantly compared with paired control plots, 

indicating that establishment of salt-accumulating halophytes on salt-affected 

sites can sufficiently remediate the soil to the point where it can be returned to 

agricultural productivity or where other native plants can invade and become 

established (Keiffer and Ungar 2002).

Regions such as the Powder River basin in Wyoming have seen a large increase 

in coal bed methane (CBM) development. Because water permeates coal beds 

and maintains the pressure needed to trap methane within the coal, CBM production requires that water first be drawn off, lowering the pressure so gas can flow 

out of the coal and into the well bore. Commonly this water is saline and poses 

a disposal issue since surface disposal can affect streams and other habitats and 

subsurface reinjection increases the cost of production (USDI GS 2000).
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Hydrologic modifications—As with other wetland types, hydrologic modifications are likely the single greatest factor altering salt flats. Hydrologic alterations 

can be on-site in origin or can result from actions off-site. Specific causes of 

alteration include direct ditching to bring water to or away from an area, seepage 

from features such as stock ponds and irrigation ditches, and direct application 

of irrigation water. Modifications may obstruct drainage, preventing water from 

flowing to existing sites, or they may result in increased drainage, flushing of 

salts, or a change of vegetation away from halophytes. Roads, culverts, and 

collection structures all can alter hydrology and may act to either increase or 

decrease salt flat vegetation.

Agriculture—Cultivation is possibly the most drastic type of disturbance capable 

of overriding other natural gradients (Walker and Coupland 1968). Alterations 

of hydrologic regimes through drainage and land use practices were identified 

as one of the main factors affecting floristic composition in the Rainwater Basin 

(Stutheit and others 2004). Irrigation is the most common anthropogenic cause 

of salinity. As water is removed from the soil by ET, salt in soil solution may 

become 4 to 10 times greater than irrigation water within 3 to 7 days (Chapman 

1975). In much of the West, the dominant salts present in irrigation water are 

chlorides; sulfates; carbonates; and bicarbonates of calcium, magnesium, sodium, 

and potassium.

Livestock—Palatability of halophytes varies among species. Distichlis spicata

(saltgrass) can be a salt-tolerant forage crop, although it requires more selection and breeding (Bustan and others 2005). The palatability of some salt flat 

species changes through development. For example, Atriplex spp. may provide 

good forage when plants are young, but old leaves become unpalatable because 

of increased concentrations of sodium, chloride, and oxalate (Waisel 1972). 

Although forage and seed products from halophytes can replace conventional 

ingredients in animal feeding systems, there are restrictions on their use due to 

high salt content and anti-nutritional compounds present in some species (Glenn 

and others 1999). Distichlis spicata is grazed readily by livestock (Bustan and 

others 2005). Livestock effects on plant diversity may depend on salinity. In 

more productive grasslands, grazing may increase plant diversity; however, in 

arid or very saline environments, diversity often is unchanged or can decrease 

(Milchunas and others 1988). For example, herbivores in tallgrass prairies on 

poor soils reduced plant diversity, whereas those on rich soils increased diversity 

(Olff and Ritchie 1998). Livestock may promote salinization in some settings 

by encouraging the formation of salt crusts through changes to evaporation 

rates following trampling (Lavado and others 1993; Lavado and Taboada 1987). 

Livestock and large native ungulates can also affect habit through impacts to soil 

structure. Trampling can cause soil compaction and reduced infiltration (Daniel 

and others 2002).
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Exotic species—A number of exotic species may occur in salt flats. The success of many Eurasian exotics is due in part to the ability of many to tolerate 

saline soils. Salt flats are susceptible to invasion by exotic perennial species 

like Cardaria latifolia and non-native annuals such as Chenopodium glaucum. 

Other common exotics are Salsola spp., Cirsium arvense, Bassia hyssopifolia, 

and Kochia scoparia (Rocchio 2006a).

Atmospheric deposition of pollutants—Practices including the use of commercial fertilizers, increases in the livestock numbers, and manufacturing and 

transportation growth have led to increases in nitrogen deposition (Baron and 

others 2000). Although unstudied in Region 2 salt flats, possible ecological 

responses to atmospheric deposition of pollutants can include changes in soil 

fertility and shifts in vegetation composition (Vitousek and others 1997). Pollutants such as SO2 may influence salt flat vegetation through both fertilizing and 

toxic effects on vegetation depending on SO2 concentration, duration of exposure, 

plant species, and environmental factors such as light, temperature, and nutrient 

regime (Milchunas and others 1981). One study found that the biweekly addition 

of nitrogen to a salt marsh led to an increase in plant biomass and cover, but no 

reduction in diversity (Traut 2005). Dominant species such as Salicornia virginica, D. spicata, and Triglochin concinna thrived with the additional nitrogen, 

but did not displace subdominant species, suggesting that competitive exclusion 

does not occur (Traut 2005).

Abundance, distribution, and condition of salt flats across 

Region 2 landscapes

Relatively few data are available as a baseline in evaluating possible changes 

in salt flat conditions or abundance. Available data include NRCS soil surveys, 

which list soils by series and describe their suitability for agriculture, among 

other uses. In soil surveys, map units are typically assigned a capability class, 

which is the broadest category in the land capability classification system. Class 

codes from 1 to 8 are used to indicate increasing limitations of soils. Map units 

can be assigned subclass codes; for example, class “s” is made up of soils that 

have soil limitations within the rooting zone, such as shallowness of the rooting zone, stones, low moisture-holding capacity, low fertility that is difficult to 

correct, and salinity or sodium content (USDA NRCS 2005). Such information 

can be used to identify areas of high salinity. Soil series descriptions in surveys 

often list depth to water table, which, in conjunction with salinity information, 

could be used to map salt flats.

Direct hydrologic alteration, especially from agricultural activities, is likely the 

most significant anthropogenic impact. Agricultural impacts were by far the 

greatest contributor to wetland loss during the late Eighteenth to the late Twentieth Century (Dahl 1990). Changes in the quality, quantity or timing of surface 

and groundwater flow into salt flats is also common due to indirect hydrologic 

alterations of surrounding landscapes, for example due to roads, storm water and 
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irrigation outflows, groundwater pumping, or plowing that changes the direction 

of surface water runoff (Cooper and others 2006; Richter and others 1996). In 

addition to probable loss of many salt flats, anthropogenic activities have created salt flats in novel landscape conditions. Features such as stock watering 

ponds and water storage reservoirs, irrigation ditches and canal-fed wetlands, 

and roadside ditches have expanded the distribution of salt flats.

Only a small fraction of salt flats in the region have been evaluated botanically, 

so it is impossible to confidently assess departures in vegetation attributes from 

historical conditions. Although declines in species richness have been documented in response to disturbances for other ecosystems, too little research has 

been completed in salt flats to make generalizations, although we assume that 

many salt flats have been altered as a result of land use changes. Hydrologic 

regimes have been extensively modified, although few data are available. Where 

longer-term data are available, for example, in Colorado’s San Luis Valley, data 

demonstrate changes in hydrologic process such as water table depth and inundation frequency (Cooper and others 2006). Chemical changes have also occurred, 

directly in response to factors like atmospheric nitrogen deposition and indirectly 

due to altered salt transport and deposition patterns from irrigation.

Management Opportunities and Constraints _______

Management opportunities

Although wetlands are generally recognized as providing important ecological 

functions, salt flats are not necessarily widely valued. This is due, in part, to 

their low productivity and diversity and to the negative impact of high salinity 

on activities such as agriculture. Little information is available on the functions 

and ecological services provided by salt flats. However, the important role of 

many saline wetlands as bird habitat has been demonstrated, providing an incentive for preservation of playas. More regional assessments of salt flats should be 

conducted. Possible tools include aerial photographs or other remotely sensed 

data to identify and map wetlands and salt affected ecosystems (Cooper and 

others 2006; Farifteh and others 2006; Metternicht and Zinck 2003). Site-level 

assessments of salt flat hydrology, water chemistry, and vegetation relationships are also needed. Modest investments of time and money for installation 

of groundwater-monitoring wells and other environmental sensors can provide 

valuable data.
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Restoration

Little research has been directed at restoring salt flats—more is known about 

methods to alleviate certain types of salinity, such as adding gypsum to leach 

sodium from soils (Davis and others 2003; Davis and others 2003; Skarie and others 1987). The restoration of inland salt-affected wetlands has received relatively 

little attention. Notable exceptions are studies of the salt-affected communities 

of Cheyenne Bottoms, Kansas. The Nature Conservancy restored native plant 

communities subject to grazing and former cropland by reestablishing sheet flow 

across disturbed areas. Results were mixed, indicating the difficulty of restoring 

heavily disturbed plant communities (Kindscher and others 2004).

As previously discussed, a key first step in any wetland restoration project is to 

gain an understanding of hydrologic regimes driving the ecological function of 

a site. Any control structures (e.g., drainage tiles or ditches) need to be identified 

and removed. Proper selection of species is important, particularly for salt flats, 

since the salt tolerance of different species varies widely (Table 27) and ecotypic 

variation should also be considered (Aschenbach 2006). Seeds for restoration 

can be collected in the field, stratified, germinated, and grown into seedlings for 

use in restoration projects. A benefit of such an approach is the ability to place 

species directly into the physical environments where they will perform best. 

Soil texture and chemistry measurements are essential.

Table 27—Summary of the hydrologic and salinity tolerances of a variety of wetland species in relation to salinity and 

hydrologic regime.

Hydrologic regime

 Perennial saturation/ Seasonal saturation/moderate deep Seasonal saturation/ no 

Salinity stable water table flooding/ fluctuating water table flooding/fluctuating water table

Fresh Calamagrostis canadensis, Beckmania syzigachne, Calamagrostis Deschampsia cespitosa,

<2 mS/cm Carex aquatilis, Carex aurea, canadensis, Calamagrostis inexpansa, Equisetum arvense,

Carex norvegica, Carex Carex utriculata, Eleocharis palustris, Juncus balticus

raymondii, Carex rostrata, Equisetum arvense, Glyceria borealis,

Carex utriculata, Eleocharis Glyceria grandis, Glyceria striata,

acicularis, Equisetum arvense, Juncus balticus, Phragmites australis,

Glyceria striata, Menyanthes Poa palustris, Scirpus microcarpus,

trifoliata, Muhlenbergia Scirpus pungens, Scirpus validus,

glomerata, Poa palustris Typha latifolia

Moderately Calamagrostis inexpansa, Beckmania syzigachne, Calamagrostis Calamagrostis inexpansa,

saline 2- Carex utriculata, Kobresia inexpansa, Eleocharis palustris, Juncus balticus, Puccinellia

15 mS/cm simpliciuscula, Muhlenbergia Juncus balticus, Phragmites australis, nuttalliana, Spartina gracilis

glomerata Scirpus pungens, Scirpus validus,

Scolochloa festucaea, Sparganium

eurycarpum Spartina pectinata,

Typha latifolia

Saline 15- Triglochin maritima, Juncus balticus, Phragmites australis, Juncus balticus, Spartina

45 mS/cm Triglochin palustris Spartina pectinata, Triglochin maritima gracilis, Puccinellia nuttalliana,

Triglochin maritima

Hypersaline Triglochin maritima Puccinellia nuttalliana,

>45 mS/cm Spartina gracilis
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Because many salt flat species are represented in the soil seed bank (Badger and 

Ungar 1990), direct placement of soils to restoration areas could potentially be 

used for establishing plant cover. Seed bank assays should be conducted to assure that the correct species are present in the soil. Typically, only the upper soil 

horizons contain viable seed. Depending on the proximity of restoration areas 

to other wetlands, natural seed rain may be effective.

Management constraints

Management constraints can include lack of funding to develop management 

and restoration goals and insufficient information on the nature of the resource. 

Legal and political constraints are often imposed by issues involving wetlands 

because of water rights, so it is best to identify potential stakeholders early in 

planning. Constraints outside the control of managers include regional climate 

change and changes in surrounding land use.
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8. Historical Range of Variation for 

Region 2 Wet Meadows

Definitions and Concepts of Wet Meadows ________

Wet meadows are the most widespread wetland type in Region 2, occurring from 

the plains to the alpine zone. Although wet meadows as a class are variable in 

vegetation, geomorphic setting, and hydrologic function, they differ from other 

wetland types in several ways. Wet meadows lack the perennial high water tables 

and organic soils found in fens, the large seasonal and inter-annual water table 

fluctuations characteristic of marshes, the high salt concentrations of salt flats, 

and the direct influence of high energy surface water flows characterizing riparian 

ecosystems. Wet meadows are typically highly productive and have historically 

been heavily utilized. Since the arrival of Euro-Americans, wet meadows have 

been commonly managed for livestock forage and hay production. Although wet 

meadows are abundant, particularly in intermountain basins and on the plains, 

it is one of the least studied wetland types.

Geomorphic Setting and Principal

Ecological Drivers _____________________________

Geomorphic setting

Considering that wet meadows span all elevations and ecoregions in Region 2, it 

is not surprising that they occur in a variety of geomorphic settings. However, wet 

meadows have seasonally saturated conditions and most are supported, in part, 

by groundwater and are found in landscape settings where groundwater levels 

are near the soil surface. Wet meadows may form in topographic low positions, 

such as around lakes and ponds (Figure 45). In these settings, marshes may occupy the deepest portion of the basin, with wet meadows on the margins. Such 

zonation among wetland types is common in wetlands in the Great Plains such 

as the Nebraska Sandhills and prairie pothole region (Johnson and others 1987; 

Kantrud and others 1989a; Seabloom and others 1998).

Wet meadows may also occur adjacent to riparian communities (Figure 45). In 

these settings they may be hydrologically connected to the stream but lack the 

influence of flooding that characterizes the active riparian zone. At higher elevations, the area of wet meadow surrounding riparian zones increases relative to the 

area of the active riparian influence. This is due to the greater groundwater inputs 

and smaller streams, which affect a smaller area with overbank flooding relative. 

In the alpine, wet meadows often occur in saddles (Walker and others 2001).
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Figure 45—Wet meadows from across the region: (A) Oglala National Grassland (Nebraska), (B) Black Hills 

National Forest (South Dakota), (3) Bighorn National Forest, (Wyoming), and (4) Rocky Mountain National 

Park (Colorado).

Key drivers and ecological processes

Climate—Wet meadows exist along nearly the entire elevation gradient in Region 2. Because of this, it is difficult to generalize about the influence of climate 

on wet meadows. The prevalent precipitation type (rain versus snow) and the 

direct importance of precipitation to the hydrologic functioning and vegetation 

of wet meadows varies widely. Snow and rain differentially influence processes 

of infiltration and surface water runoff and affect groundwater recharge. Consequently, the proportion of a watershed with a snow- versus rain-dominated 

precipitation regimes can substantially influence the abundance and function of 

wet meadows. At low elevations, most precipitation falls as rain, although the 

amount and seasonality varies across the region. At higher elevations, snow and 

rain-on-snow precipitation regimes become dominant, and the cooler and wetter 

climate can support more wetlands, including wet meadows.
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Temperature is the primary factor determining the length of the growing season. 

Mean temperatures generally decline with increasing elevation. However, rugged 

topography and the resulting processes of cold air drainage can create a great deal 

of spatial variability in temperatures. Site aspect is also important, particularly 

in subalpine and alpine areas where the growing season length is controlled in 

large part by the timing of snowmelt. In a study of alpine tundra wet meadow 

soils in Colorado, researchers found that temperatures were relatively stable near 

freezing for much of the spring and autumn, and water availability was limited 

by freezing rather than by drying (Costello and Schmidt 2006).

In the alpine zone, strong and variable winds and complex topography create 

highly heterogeneous patterns of snow accumulation. The timing of snow melt 

varies, with snowmelt arriving early in windward and ridge top locations (Billings and Bliss 1959). Wind-driven redeposition of snow creates substantial local 

variation in the amount of effective precipitation. Because the timing and duration 

of spring thaw determines the fraction of this precipitation that ultimately passes 

into the soil, direct measurements of precipitation may poorly reflect the effective 

precipitation affecting a site (Taylor and Seastedt 1994). For plants, the timing 

of key phenological events (e.g., the initiation of growth) is related to snowmelt 

patterns, resulting in differences among communities (Walker and others 1995).

Geology—A number of geologic factors affect the distribution and function of 

wet meadows across the region. Differences among watersheds in lithologic characteristics influence the composition and concentration of ions in groundwater as 

well as characteristics such as porosity and permeability. Because wet meadows 

may be supported by inputs of groundwater, the mineralogy of watersheds can 

influence water chemistry, biota, and ecological processes.

Geomorphic history is also important. Pleistocene glaciation in mountainous 

areas of Region 2 strongly influenced the configuration of many contemporary 

landscapes (Richmond 1960; Thornbury 1965; Wright and Porter 1983). Large 

wetland complexes supporting extensive wet meadows are common in broad, 

glaciated valleys (Windell and others 1986). For example, large wetland complexes on both the east and west side of the Continental Divide of Colorado’s 

Front Range in Rocky Mountain National Park support marshes, fens, riparian 

shrublands, and wet meadows (Cooper 1990a; Cottrell 1993).

Hydrologic regime—Hydrologic regime is the primary factor distinguishing wet 

meadows from other wetlands types and is an important influence on biota and 

geochemical functioning of wet meadows. Wet meadows occur in sites where 

soils are seasonally saturated, but perennially high water tables or seasonally 

deep water do not occur. In montane and subalpine watersheds, wet meadows 

frequently occur in stream valleys as part of larger wetland complexes. However, 

unlike riparian ecosystems, wet meadows do not experience high-velocity surface 

flows or sediment deposition from fluvial processes. Relative to marshes, wet 

meadows have stable water tables and do not experience deep inundation. 
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Mean water tables are lower in wet meadows than the threshold required for 

deep peat formation—the main characteristic that defines fens.

Sources of water to wet meadows include direct precipitation, surface water, 

and groundwater, with the relative importance of each varying among individual 

wetlands. Wet meadows receive a portion of their annual water budget through 

direct precipitation. The amount is generally scaled to wetland area, although 

in alpine and subalpine areas where extensive drifting of snow may occur, the 

relationship may not hold. Overall, precipitation is likely the least important direct contributor to hydrologic regimes in wet meadows in the Rocky Mountains. 

Direct surface water inputs to wet meadows are generally limited to sheet flow 

following extreme precipitation events or during spring snowmelt. Channelized 

flows generally do not occur in wet meadows, although wet meadows are often 

found in close proximity to streams. While the overall frequency of flooding in 

wet meadows is low, wetlands in settings such as wide, low-gradient alluvial 

valleys may be periodically flooded due to the activities of beaver.

For most wet meadows, groundwater is the most important hydrologic input; 

however, there is considerable variability in the size and characteristics of aquifers. 

Wet meadows in the Nebraska Sandhills may be supported by regional aquifers, 

while small and local hillslope aquifers support wet meadows in montane environments (Winter and others 2001). High early summer water tables maintained 

by snowmelt occur in both fens and wet meadows. The hydrologic distinction 

is that in wet meadows, the aquifer provides less stable groundwater flows and 

the water table may drop in the middle to late summer. The water table in wet 

meadows may drop to 1 m or more below the soil surface in July and August. In 

fens, with the exception of dry years, the summer water table remains within approximately 20 to 40 cm of the soil surface (Chimner and Cooper 2003; Chimner 

and others 2002; Cooper 1990a).

Water and soil chemistry—The chemical properties of wet meadow soils are 

closely related to both hydrologic regimes and plant species distributions. Water 

and soil pH and the concentration of cations and nutrients vary among wet meadows. Site-level variables such as soil particle size distribution, organic matter 

content, and groundwater source influence soil chemistry, as do watershed-scale 

characteristics such as the mineral composition of meadow soils and surrounding landscapes. For instance, where relatively soluble minerals are dominant, 

groundwater may contain abundant dissolved solids, while areas composed of 

relatively insoluble minerals will have lower concentrations of ions. The pH in 

wet meadow soils can vary widely but is typically neither extremely acidic nor 

basic. Wet meadows in the Nebraska Sandhills, for example, ranged from 5.9 to 

7.6 (Kapustka and others 1988).
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Most wet meadows have relatively fine-textured soils that are dark in color with 

high organic matter content. Anaerobic conditions lead to the formation of hydric 

soil indicators such as mottles and oxidized root channels, and these features 

can be used to delineate wet meadows from dry meadows, particularly where 

vegetation has been disturbed. Nutrients are generally more abundant than in 

adjacent uplands—a factor contributing to the high primary productivity seen 

in many wet meadows.

The dominant forms and abundance of nutrients varies among wet meadows. 

For example, subalpine meadow soils in Colorado had low NO3

-

 concentrations, 

likely the result of assimilation and denitrification in anoxic meadow soils (Clow 

and Sueker 2000). Variation in microbial processes is highly related to soil 

moisture. For example, the topographic soil moisture gradient found in alpine 

areas is a fundamental control of nitrogen turnover patterns among communities 

(Fisk and others 1998).

When present in a watershed, beavers can significantly affect biogeochemical 

pathways by shifting element storage from vegetation to sediments and soils. The 

fundamental control over subsequent alterations of biogeochemical pathways is 

the occurrence of anaerobic conditions from waterlogged soils. As ponds fill, 

sediments accumulate substantial standing stocks of chemical elements that are 

available for vegetation growth. Short-term hydrologic variations can cause 

shifts in the aerobic-anaerobic boundary in the upper soil layers of wet meadows, 

enhancing nitrogen cycling and nitrogen availability for plant growth in both 

active and abandoned ponds (Pinay and Naiman 1991).

Disturbance processes—Wet meadows are subject to a range of natural and 

anthropogenic disturbances. Relative to riparian areas or marshes, abiotic disturbances such as floods are less important, while biotic disturbances from animals 

are of greater ecological significance. Data describing disturbance frequency, 

magnitude, or return interval from wet meadows are largely non-existent in the 

region. Unlike forests, where dendrochronological approaches to reconstructing disturbance processes are available, inferences about disturbances in wet 

meadows must be made from anecdotal accounts and case studies, often from 

research conducted outside of the region.

The role of fire in wet meadow dynamics is poorly understood. Wet meadows in 

Region 2 occur as relatively small, discrete patches set within other cover types. 

Ignitions in wet meadows are likely rare and fires affecting them originate from 

outside of the wetland’s boundaries. Because fire regimes vary widely within 

the region in response to factors such as cover type, physiography, and climate, 

it is impossible to generalize on fire regime parameters for wet meadows other 

than to say that, historically, most wet meadows would have burned on some 

time scale. Fire in wet meadows may cause direct plant mortality, influencing 

community composition, and may favor herbaceous species over woody ones 
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(DeBenedetti and Parsons 1984; Jacobson and others 1991). Indirect effects may 

include increased water, sediment, and nutrient flux from surrounding hill slopes 

(Anderson and Menges 1997).

Disturbance caused by animals are locally important in many wet meadows. Native ungulates such as deer, elk, and bison, as well as domestic livestock such as 

sheep and cattle can have important effects on the abiotic and biotic environment 

in wet meadows. Because of their high productivity, wet meadows are attractive to 

grazers, and have a long evolutionary history of herbivory. In conifer-dominated 

montane areas, meadows may provide the most significant resource of graminoid 

vegetation. Large animal effects can include plant mortality from herbivory or 

trampling. In addition, animals may alter soil chemistry and nutrient availability 

though the deposition of urine and feces (Frank and Evans 1997). Historically, 

significant local impacts to vegetation likely occurred but because of variable 

population density and migratory behavior, the effects of native ungulates on 

wet meadows at broad spatial or temporal scales was not likely great.

Small mammals such as pocket gophers can have a significant influence on wet 

meadows. Through their burrowing and foraging, they can affect patterns and 

rates of soil development, nutrient availability, microtopography, demography and 

abundance of plant species, and plant diversity (Ellison and Aldous 1952; Huntly 

and Inouye 1988; Ward and Keith 1962). For example, erosion measurements 

in a Colorado alpine ecosystem indicated long-term susceptibility of gopherdisturbed soils to redistribution by water and/or wind (Sherrod and Seastedt 2001). 

Research in Arizona found that pocket gophers can act as a keystone herbivore 

by limiting aspen invasion of meadows (Cantor and Whitham 1989). Small soil 

disturbances can also be important for the establishment of many species. For 

example, cohorts of invading sagebrush seedlings in California meadows were 

preferentially associated with gopher mounds (Berlow and others 2002). Other 

mammals that may influence wet meadows are voles (Austin and Pyle 2004; 

Howe and Lane 2004) and marmots (del Moral 1984).

Although most important in riparian ecosystems, beavers may create disturbances 

important to wet meadows. Beavers are a major geomorphic driver of meadow 

formation and landscape dynamics in some montane and subalpine areas. In 

low-gradient glacial valleys, which often support wetland complexes with both 

riparian and wet meadows ecosystems, beaver dams may cause channel avulsions flooding surrounding wet meadows (Cooper and others 2006). Abandoned 

beaver ponds may develop into wet meadows or marshes depending on resulting 

water table characteristics. At broad spatial scales, the activities of beaver can 

create a complex, heterogeneous pattern of wetland types, including wet meadows, marshes, and riparian ecosystems, thereby increasing landscape patterns 

of species richness (Wright 2002).
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In many mountain watersheds, spatial and temporal variability in the location 

and duration of beaver ponds is a major factor influencing the hydrologic, geochemical, and community-level characteristics of wetland complexes (Cooper 

and others 2006; Westbrook and others 2006). Beavers can also be important 

drivers of vegetation structure and composition, acting directly though their 

selection of species for food and dam construction and indirectly through their 

hydrologic effects (Baker and others 2005a; Barnes and Mallik 2001; Naiman 

and others 1988).

Post-disturbance recovery to natural and anthropogenic disturbances varies. Alpine 

areas are, in general, slower to recover than sites at lower elevations (Ebersole 

2002). Individual species also vary widely in their response to disturbance. Some 

species respond positively or even require disturbance in order to persist. An 

example is Spiranthes dulivialis, a diminutive orchid found in riparian areas and 

adjacent wet meadows at low elevations and listed by the USFWS as threatened 

under the Endangered Species Act (Sipes and Tepedino 1995). Management 

practices employed to promote the species in some areas include active mowing. Mowing is also used in Europe to promote species diversity (Grootjans and 

others 2002; Vinther and Hald 2000).

Wet meadow development and succession—Successional patterns vary because 

a wide range of wet meadow vegetation types exists. Alpine communities often 

do not exhibit well-defined successional sequences. The relative abundance of 

species following disturbance remains nearly constant in some sites following disturbance, while at others, one or more species may increase in relative 

abundance over time (Ebersole 2002). In subalpine and montane meadows, researchers have described successional shifts towards increased shrub dominance 

(Langenheim 1962), although many meadows show no such trend and appear 

to be physiognomically stable. The mechanisms driving different successional 

trajectories among meadows are complex. For example, abandonment and drainage of beaver ponds may result in the formation of meadows, which may persist 

for decades as graminoid-dominated patches that resist conifer invasion despite 

close proximity to seed sources (Terwilliger and Pastor 1999).

Wet Meadow Classification and Gradients _________

Wet meadows have been classified in a number of ways. Specific criteria used 

by different classification approaches to differentiate wetland classes include 

dominant vegetation, hydrologic characteristics, geomorphic setting, and water 

chemistry. Under the Cowardin classification used by the National Wetland Inventory, all wet meadows fall within the palustrine system, and depending upon 

their vegetation, are placed in the emergent or scrub-shrub class. However, fens 

and salt flats may be identically classified, reducing the utility of NWI maps for 

analyzing individual wetland classes like wet meadows.
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The Hydrogeomorphic (HGM) approach used to classify wetlands (Brinson 

1993a; Cooper 1998a) emphasizes physical variables such as geomorphic setting and hydrologic regime in classification. Several subclasses relevant to wet 

meadows were identified in a preliminary classification of Colorado wetlands 

using HGM (Table 28). Wet meadows in Slope Subclasses 1 and 2 (S1/2) may 

be dominated by woody or herbaceous species (Carsey and others 2003). Slope 

Subclass 1 wetlands are very common and widespread in mountainous portions 

of Colorado, and similar ecosystems are found in Wyoming. Slope Subclass 2 

wetlands occur on calcareous substrates and have higher concentrations of mineral 

ions (e.g. Ca2+). These are less common since most Rocky Mountain ranges are 

composed of igneous or metamorphic rocks. Wetlands in Slope Subclasses 3 and 

4 (S3/4) includes wet meadows at middle elevations in the mountains. Similar 

wetlands also occur in Wyoming and the Black Hills. Slope 4 wetlands occur at 

lower elevations but have a seasonally high water table that supports herbaceous 

or occasionally shrub associations (Carsey and others 2003).

Stewart and Kantrud (1971) classified wetlands in the prairie pothole region into 

one of seven classes based on water permanence, quantity, and quality, as well 

as the vegetation communities present. Ephemeral ponds are often wet for only a 

few weeks each year, and vegetation is dominated by facultative wetland plants 

such as Poa pratensis, Anemone canadensis, and Symphoricarpos spp. Wetlands 

in the temporary pond class support wet meadow communities in the deepest 

part of the basin and are usually fringed by the low prairie zone along meadow 

margins. Common species include Poa palustris, Carex lanuginosa, Juncus 

balticus, Eleocharis spp, Hordeum jubatum, Calamagrostis stricta, Spartina 

pectinata, and Mentha arvensis. In this classification, a shallow marsh zone occurs in the center of seasonal ponds and lakes, surrounded by wet meadows and 

wet prairie. The deepest portion of semi-permanent ponds and lakes support a 

deep marsh zone, followed by the shallow marsh zone, and lastly a wet meadow 

zone (Stewart and Kantrud 1971).

Table 28—HGM subclasses relevant to wet meadows in Colorado (Carsey and others 2003).

HGM Subclass description Common species

Slope 1 Alpine and subalpine fens and wet meadows Carex aquatilis var. stans, Carex

on saturated non-calcareous substrates. scopulorum

Slope 2 Subalpine and montane fens and wet Eleocharis quinqueflora, Kobresia

meadows on saturated calcareous substrates. simpliciuscula, Carex simulata

Slope 3 Wet meadows at middle elevations in the Juncus balticus var. montanus

mountain ecoregion with a seasonal high

water table near the ground surface.

Slope 4 Low elevation meadows with a seasonal Carex nebrascensis

high water table near the ground surface.

May occur on floodplains or near springs.

Depressional 1 Lake fringes without peat soils. Carex utriculata
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Wet Meadow Vegetation in the Region ____________

Because wet meadows occur over such broad environmental gradients, they 

are highly variable in species composition and are one of the most diverse and 

productive ecosystem types in Region 2. Comer (2003) identified four primary 

ecological systems in Region 2 relevant to wet meadows (Table 29). More than 18 

wet meadow vegetation associations are recognized by Natureserve (NatureServe 

2005), although this likely misses some associations in the region (Table 30). 

Some meadow species have large elevation ranges and considerable ecotypic 

variation. For example, Deschampsia cespitosa and Poa pratensis occur from low 

elevations to the alpine (Pearcy and Ward 1972) and may occur in wet meadows 

in some zones but not in others.

Some community types are very common. For example, wet meadows dominated by Juncus balticus are found widely throughout the western United States, 

occurring from low to middle elevations. Shrubs are typically absent or their 

cover is low in wet meadows, however Dasiphora fruticosa commonly occurs in 

montane wet meadows with Juncus arcticus. The species is resistant to drought 

and can persist or expand in response to drought or disturbance such as grazing. 

Introduced perennial sod-forming grasses such as Agrostis stolonifera are often 

co-dominants (NatureServe 2005), particularly at lower elevations. In the Nebraska Sandhills, vegetation is dominated by introduced cool-season grasses and 

legumes, along with native sedge and Juncus species, in contrast to upland sites 

that are dominated by native, warm-season grasses (Volesky and others 2004).

Communities dominated by Schoenoplectus pungens are common at low elevations and, like Juncus balticus communities, are easily identified at a distance by 

their dark green to black colors. The community occurs adjacent to ponds, lakes, 

or low-gradient streams. Shrubs and trees are usually absent. Carex-dominated 

communities are also common. For example, wet meadows dominated by Carex

nebraskensis occur throughout much of the West from as low as 1000 to over 

2800 m above sea level (NatureServe 2005). Other common species are Carex 

praegracilis, Calamagrostis stricta, Deschampsia cespitosa, Eleocharis palustris, Glyceria striata, Juncus balticus, and Schoenoplectus pungens. Common 

wet meadow associations in the alpine zone include those dominated by Carex

scopulorum, Pedicularis groenlandica, Caltha leptosepala, and Rhodioloa integrifolia. Communities typically occur down-slope of snow banks. Wet meadows 

are most common in the lower alpine zone, although some rare associations 

occur at the base of rock glaciers at higher elevations (Walker and others 2001).

Table 29—Ecological systems relevant to wet meadows in Region 2 (Comer and others 2003b).

Ecological systems Hectares % Area

Eastern Great Plains Wet Meadow Prairie and Marsh 138,843 0.1

Western Great Plains Riparian/Western Great Plains Floodplain 1,488,930 1.3

Rocky Mountain Subalpine-Montane Riparian Shrubland 475,655 0.4

Rocky Mountain Subalpine Mesic Meadow 884,960 0.8
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 193

Table 30—Graminoid-dominated wet meadow alliances and vegetation associations in Region 2 (NatureServe 2005).

Alliance Association Code

Calamagrostis canadensis Seasonally Flooded Herbaceous Alliance

Calamagrostis canadensis Western Herbaceous Vegetation CEGL001559 

Carex (rostrata, utriculata) Seasonally Flooded Herbaceous Alliance

Carex rostrata-Carex lacustris-(Carex vesicaria) Herbaceous Vegetation CEGL002257

Carex utriculata Herbaceous Vegetation CEGL001562 

Carex aquatilis Seasonally Flooded Herbaceous Alliance

Carex aquatilis-Carex spp. Herbaceous Vegetation CEGL002262 

Carex atherodes Seasonally Flooded Herbaceous Alliance

Carex atherodes Herbaceous Vegetation CEGL002220 

Carex buxbaumii Seasonally Flooded Herbaceous Alliance

Carex buxbaumii Herbaceous Vegetation CEGL001806 

Carex nebrascensis Seasonally Flooded Herbaceous Alliance

Carex nebrascensis Herbaceous Vegetation CEGL001813 

Carex pellita Seasonally Flooded Herbaceous Alliance

Carex pellita-Calamagrostis stricta Herbaceous Vegetation CEGL002254 

Carex spp.-Plantago eriopoda Temporarily Flooded Herbaceous Alliance

Calamagrostis stricta-Carex sartwellii-Carex praegracilis-Plantago

eriopoda Saline Herbaceous Vegetation CEGL002255 

Carex spp. Seasonally Flooded Herbaceous Alliance

Carex diandra Wet Meadow Herbaceous Vegetation CEGL002549 

Deschampsia caespitosa Saturated Herbaceous Alliance

Deschampsia caespitosa-Caltha leptosepala Herbaceous Vegetation CEGL001882 

Eleocharis palustris Seasonally Flooded Herbaceous Alliance

Eleocharis palustris Herbaceous Vegetation Marsh CEGL001833 

Glyceria borealis Semipermanently Flooded Herbaceous Alliance

Glyceria borealis Herbaceous Vegetation CEGL001569 

Juncus balticus Seasonally Flooded Herbaceous Alliance

Juncus balticus Herbaceous Vegetation CEGL001838 

Phalaris arundinacea Seasonally Flooded Herbaceous Alliance

Phalaris arundinacea Western Herbaceous Vegetation CEGL001474 

Schoenoplectus pungens Semipermanently Flooded Herbaceous Alliance

Schoenoplectus pungens Herbaceous Vegetation CEGL001587 

Spartina pectinata Temporarily Flooded Herbaceous Alliance

Spartina pectinata-Carex spp. Herbaceous Vegetation CEGL001477

Spartina pectinata-Schoenoplectus pungens Herbaceous Vegetation CEGL001478
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 194

Hydrologic regime is generally the most important factor influencing vegetation patterns and dynamics in wet meadows, although in alpine tundra, patterns 

of snowpack accumulation and melt-out strongly control vegetation zonation 

(Billings and Bliss 1959). Individual species vary in their response to water 

table gradients, with some species (e.g., Juncus arcticus) capable of growing 

along a wider range of water table depths than others (Dwire and others 2006). 

In a study evaluating the response of 63 riparian grassland species to hydrologic 

variation, the best plant-frequency response curves were obtained by using the 

growing season 10 percent cumulative frequency water level, followed closely 

by the growing season seven-day moving average high water level, suggesting 

that high water levels may be more influential than mean, median, or low water 

levels. (Henszey and others 2004).

Water table dynamics are also important drivers of community-level characteristics such as species richness and total plant cover. For instance, Dwire and 

others (2006) found that wet meadow communities dominated by sedges (Carex

spp.) had lower species richness and diversity compared to dry meadow communities, which supported a mixture of grasses and herbaceous dicots. In their 

study, species richness and total plant cover were negatively correlated with 

mean water table depth and positively correlated with mean redox potential at 

10-cm and 25-cm depths (Dwire and others 2006). In the wettest wet meadow 

communities, obligate hydrophytes tolerant of seasonal flooding, shallow water table depths, and anaerobic soil conditions generally dominate. In contrast, 

competition and other biotic interactions likely play a greater role in determining 

species composition and vegetation structure in the more diverse moist and dry 

meadow communities (Dwire and others 2004).

Most wet meadows exhibit seasonal declines in their water table. This is particularly true of wet meadows in subalpine and alpine areas that receive significant 

hydrologic inputs from snow. The magnitude of water table fluctuations varies 

widely, and drawdown may or may not create moisture stress for plants. In alpine areas, average moisture values for wet meadows at the end of the growing 

season are still higher than for mesic meadows in the beginning of the season, 

and other factors such as release from snow, solar radiation, and soil nutrients 

generally limit production (Billings and Bliss 1959; Ehleringer and Miller 1975; 

Taylor and Seastedt 1994). At lower elevations, the drop in water tables may be 

more significant and may result in decreased plant physiological performance 

(Svejcar and Riegel 1998).

Although the main variable driving species responses and interactions within 

wet meadows is the water table, direct and indirect effects of disturbances such 

as livestock grazing can also play a role. For example, Carex nebrascensis and 

Poa pratensis commonly co-occur in meadows throughout the West, but have 

maximum expression at different water tables. Poa pratensis is capable of more 

rapidly responding to disturbances that remove neighbors and increase available space than C. nebrascensis, suggesting that grazing may alter the relative 

competitive ability of the two species in favor of Poa pratensis (Martin and 

Chambers 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 195

Temperature represents an additional variable influencing vegetation, particularly 

at higher elevations. Temperature is strongly influenced by slope position and 

aspect, particularly in rugged terrain. Soil temperatures within meadows may 

also vary along water table gradients. For example, sites in wet meadows were 

cooler than those in the dry and mesic meadow types and sagebrush vegetation 

types (Castelli and others 2000).

Some research suggests that, relative to marshes and salt flats, few meadow 

species form persistent seed banks (Bekker and others 1998). However, this 

may not be true for all wet meadows. The dense vegetation cover, dominance 

by perennial and clonal species, and relatively low disturbance frequency do not 

favor this life history attribute as strongly as in more environmentally stochastic 

ecosystems like marshes. Certainly, many species do exhibit some form of seed 

dormancy but, overall, seed banks appear to be less crucial to vegetation dynamics in wet meadows (Adamus 1996). Small-scale disturbances may be important 

for the establishment of many wet meadow species. For example, some studies 

have found that hoof prints may provide opportunities for plant establishment, 

although other studies have not observed this effect (Stammel and Kiehl 2004). 

Likewise, soil disturbance from small mammals like pocket gophers can provide 

microsites amenable for plant establishment (Ellison and Aldous 1952; Huntly 

and Inouye 1988; Sherrod and Seastedt 2001).

HRV of Wet Meadows in Region 2 ________________

Overview of anthropogenic impacts

The extent and magnitude of anthropogenic impacts to wet meadows varies 

across the region, with some landscapes, such as the Nebraska Sandhills, subject 

to high degrees of modification, and others, including most alpine areas, relatively unimpacted by direct anthropogenic impacts. Anthropogenic impacts to 

wet meadows include both on-site and off-site impacts, although some stressors 

such as hydrologic modifications can fall into both groups. The effects of some 

stressors are locally severe but only affect a small fraction of the wet meadows 

in the region. Others, such as the aerial deposition of pollutants, may affect large 

areas but have subtle ecological effects. Landscape and regional analyses of the 

relative importance of different stressors is lacking, although strong inferences 

can be made from knowledge of the dominant contemporary and historical land 

uses in the region.

Livestock—Livestock use is widespread in wet meadows and is a dominant land 

use in many portions of the region (Crowley 1975). Wet meadow species are 

an important source of forage, although the digestibility and nutrient value of 

native species varies. Some sedges have relatively high crude protein and acidpepsin digestibility levels, while others are poor forage (Catling and others 1994; 

Hermann 1970). In addition to their direct effects, hay production for livestock 

use has one of the most pervasive and significant effects on wet meadows. 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 196

The purposeful introduction and promotion of non-native pasture grasses has 

shifted the composition of many wet meadows outside of the HRV.

Livestock can have significant effects on wet meadows, including direct effects 

such as herbivory and trampling and indirect impacts such as nutrient enrichment 

from urine and feces. Specific effects vary depending on the seasonality, intensity, 

and duration of grazing, and the effects differ among livestock species. Among 

the possible impacts is alteration of species composition (Belsky and others 

1999). In alpine areas, sheep have been shown to shift the floristic composition 

of meadows, with species such as Polygonum bistortoides, which are eaten by 

sheep during the early part of the season, decreasing in abundance (Bonham 

1972). Sheep grazing during the Nineteenth Century has been identified as a 

major cause of change to meadows in the Sierra Nevada (Dull 1999). Seasonlong grazing by cattle has apparently been responsible for decreasing production 

and retarding plant succession in montane meadows in Idaho (Leege and others 

1981). In the Nebraska Sandhills, stocking rate did not affect the frequency of 

dominant species, but the frequency of legumes and Poa pratensis was higher 

in grazed pastures (Volesky and others 2004). Defoliation of taller grasses and 

sedges by grazing also altered the vertical structure of wet meadows, creating a 

more open canopy and allowing shorter-stature species to increase. The stocking rate affected more response variables than grazing frequency (Volesky and 

others 2004).

Grazing can affect soil physical and chemical characteristics. Grazed meadows 

had higher extractable Na+, Ca2+, Mg2+, and NO3

-

; higher pH; and lower K+ and 

NH4

+ than sites where cattle grazing was excluded. Grazing effects were most 

pronounced at the forest edge, likely because of the spatial transfer of nutrients 

via urine and feces (Blank and others 2006).

Where populations of native ungulates have exceeded their HRV, including several areas in Region 2 (Coughenour and Singer 1996; Lubow and others 2002; 

Singer and others 1998), impacts may be similar to those from livestock. Native 

ungulates such as elk can cause plant mortality from trampling and herbivory and 

can alter nitrogen availability though urine and feces inputs (Frank and Evans 

1997). In Yellowstone, the net mineralization of soil nitrogen was higher outside 

of exclosures than inside (Frank and Groffman 1998). Elk can affect the physical 

properties of soils by increasing bulk density (Binkley and others 2003). Where 

ungulate populations have remained within their HRV, the likely impacts to wet 

meadows have probably been localized.

Agriculture—Many wet meadows in the Great Plains are adjacent to actively 

cultivated land. Activities related to farming such as pesticide and fertilizer 

applications can affect adjacent wet meadows. These chemicals can reach wet 

meadows through processes of surface water runoff, groundwater flow, or aerial 

deposition. Surface soils in the wet meadow zones of wetlands surrounded by 

cultivated land had phosphorus concentrations 2.5 to 6 times greater than those 

surrounded by uncultivated grasslands (Freeland and others 1999). Subsoils also 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 197

differed, with higher organic matter content and concentrations of phosphorus 

and NO3

-

 in the wetland surrounded by cultivated land. Cultivated wetland 

basins also typically have higher erosion rates than those surrounded by intact 

grasslands (Freeland 1997).

Hay production is very common throughout the region, pursued as a feed source 

for livestock during the dormant season. Hay production is most common on 

private lands. Harvest dates in wet meadows are dependent upon the water 

table since equipment cannot access hay until after the water table has dropped 

sufficiently below the soil surface. For example, normal harvest dates for wet 

meadows in the Sandhills range from mid-June to mid-August (Reece and others 

1994; Volesky and others 2004).

Throughout the West, there is widespread use of irrigation for hay production 

(Smeal and others 2005). Many irrigated meadows were sown with cool season 

pasture grasses, thereby altering the species composition and functional characteristics of meadows. Irrigated pastures are significant contributors of phosphorus 

to streams and aquifers, with much of the phosphorus coming from applied fertilizers (Bush and Austin 2001). In nutrient addition experiments in meadows near 

Gunnison, Colorado, White et al (2003) found that the loadings of both P and 

NH4

+ were significantly greater when fertilizer was applied in the early or late 

spring as opposed to in the fall. The authors suggested that producers of hay in 

mountain meadows should apply fertilizer in the fall to improve hay yields and 

reduce impacts to water quality (White and others 2003).

Exotic species—A number of exotic species regularly invade wet meadows. 

These include noxious weeds, such as Cirsium arvensis and Phalaris arundinacea, and pasture grasses purposefully introduced to the region for livestock 

forage. Examples of the latter are Poa pratensis, Dactylis glomerata, and Phleum pretense. Native species such as Juncus balticus and Dasiphora fruticosa

may increase with hydrologic or physical site disturbances such as overgrazing. 

Vegetation types that are rich in native species are often highly vulnerable to 

invasion by nonnative plant species and are hotspots of exotic species richness 

(Stohlgren and others 1997, 1999).

Hydrologic modifications—Hydrologic modifications are one of the most common and significant impacts to wet meadows in the region. Actions can either 

increase the amount of water reaching a site, for example through irrigation or 

streamflow augmentation, or decrease water, as with flow diversion, ditching or 

groundwater pumping. Impacts can also be separated into those that occur onsite (e.g., ditching) and those that occur off-site (e.g., upstream flow diversion).

Many wet meadows occur along with riparian areas as part of wetland complexes. Although wet meadows do not experience regular flooding, water levels 

in many wet meadows are controlled, in part, by fluctuations in stream stage. 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 198

As a consequence, activities that affect streamflow such as upstream diversions 

or impoundments may alter hydrologic processes and the hydrologic function of 

wet meadows. As an example, stabilized flow regimes and reduced spring floods 

due to upstream water developments have reduced the area of wet meadow along 

the Platte River (Niemuth and others 2004).

Changes in vegetation commonly occur when hydrologic regimes have been 

significantly altered. In southeastern Wyoming, researchers evaluated changes in 

montane meadow vegetation following flow augmentation and found that after 

two years of elevated surface and groundwater levels, herbaceous vegetation had 

shifted toward more hydrophytic species. For example, the biomass of Carex

spp. increased from 337 to 456 g/m2 in wet meadows (Henszey and others 1991).

Groundwater pumping is an additional factor potentially affecting wet meadows, 

at least at local scales. A cone of depression is formed around the well during 

pumping, often creating a significant decline in water tables in the vicinity of 

pumps (Cooper and Wolf 2006). This may lead to moisture stress in vegetation 

and, if severe enough, eventually to shifts in species composition (Cooper and 

others 2006). As discussed previously, the application of irrigation water can 

also alter soil properties and available moisture, leading to shifts in vegetation 

composition.

Roads and transportation infrastructure—Transportation infrastructure such 

as roads, culverts, and ditches can affect the function of wet meadows. Impacts 

can be direct, such as when a road bisects a wetland, or indirect, through changes 

to hill slope processes in contributing watersheds (Forman and Sperling 2002). 

Transportation infrastructure can alter natural drainage patterns, reduce interception and infiltration rates due to the removal of vegetation and soil compaction, 

and alter the hydrologic response of basins to annual snowmelt runoff and storm 

events (Jones and others 2000). Roads can also contribute chemical pollutants, 

although there is no evidence to suggest that this is a widespread stressor for 

wet meadows in the region.

Beaver trapping—In low-gradient landscapes, the dam building activities of 

beavers are important drivers of hydrologic function and geomorphic development. The loss of beaver from trapping or habitat degradation can lead to significant changes in hydrology and vegetation (Cooper and others 2006; Peinetti 

and others 2002b; Westbrook and others 2006). Presumably, these impacts have 

at least indirectly affected wet meadows by reducing hydrologic variability and 

rates of patch creation (Johnston and Naiman 1990a).

Atmospheric deposition of pollutants—Increased atmospheric loading of 

pollutants such as nitrogen and sulfur has been well documented in much of 

Region 2 (Baron and others 2000; Fenn and others 2003b). Hotspots of elevated 

deposition typically occur downwind of large metropolitan centers or significant 

agricultural operations. A variety of ecological responses to nitrogen deposition 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 199

have been documented (Baron and others 2000; Fenn and others 2003a; Rueth 

and others 2003), but data specific to wet meadows is lacking.

Research in a Montana grassland exposed to elevated concentrations of sulfur 

and nitrogen found significant increases in leaf area at low SO2 concentrations 

but not at high levels, suggesting a fine balance between fertilization and toxicity 

(Milchunas and others 1981). Research in Canadian bogs and fens examining 

the response of species to nitrogen deposition found that the response varied 

among species (Li and Vitt 1997). However, the transferability of such studies 

is questionable considering the different soil, hydrologic, and biogeochemical 

environment of wet meadows. The applicability of responses observed in more 

arid grasslands to wet meadows is also unknown.

Cumulative effects—Multiple landscape stressors can affect ecological processes 

in wet meadows. Because of their dependence on groundwater, wet meadows are 

sensitive to changes in surrounding watershed processes (Siegel 1988). Ecological 

stressors can act synergistically and unpredictably (Holling 2001). For example, 

roads may provide a conduit for the dispersal of seeds from exotic species and 

they may create disturbance conducive for their establishment. Wet meadows 

typically have low to moderate sediment fluxes, and actions in the surrounding 

watershed that affect the quality and quantity of ground or surface water may affect wet meadow function. Watershed properties are critical variables for analysis 

in common wetland condition assessments (Faber-Langendoen and others 2006; 

Hauer and others 2002b; Richter and others 1996; Rocchio 2006c) and should 

be part of any cumulative effects assessment approach.

Abundance, distribution, and condition of wet meadows 

across Region 2 landscapes

No region-wide inventory of wet meadows has been conducted in Region 2. 

While nearly the entire region has been mapped as part of the NWI program, 

only a fraction of maps have been digitized, much of the data is outdated, and 

it’s not possible to separate wet meadows from other palustrine wetlands. As a 

result, no reliable statistics exist for wet meadow area across the contemporary 

landscape, let alone historical landscapes.

Estimates are available for portions of Region 2. For example, wet meadows were 

estimated to comprise approximately 10 percent of the 4.8 million-ha Nebraska 

Sandhills region (Volesky and others 2004). In their analysis of GAP vegetation 

data, Comer and others (2003) presented a value of 1,023,803 ha (0.9 percent 

of the Region 2 area) for the main Ecological System types they identified for 

wet meadows in Region 2. However, it is not possible to perfectly cross-walk 

from Ecological System type to the classification system used in this assessment.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 200

There are insufficient historical data to characterize trends in wet meadow condition. Estimates from the USFWS in its “Status and Trends” publications (Dahl 

2000; Dahl and Johnson 1991; Tiner 1984) are too imprecise to make specific 

quantitative estimates of wetland loss for wet meadows. It is likely that the 

pattern of change in wet meadow area has varied across the region. In general, 

higher rates of wetland loss would be expected on the Great Plains, particularly 

in sites suitable for crops such as the eastern part of the region. Wetland losses 

in mountain regions are likely smaller as more area is under public land management, mostly in National Parks or USFS wilderness areas.

While many wet meadows have been destroyed, others have been created as a 

result of anthropogenic activities. Examples include wet meadows created by 

irrigation of former non-wetlands, ditching and draining of fens, or fringe wet 

meadows formed around ponds and reservoirs. On balance, though, the likely 

trend for all of Region 2 has been toward a net loss of wet meadows since EuroAmerican settlement.

A separate set of questions can be asked regarding historical changes in the 

condition or integrity of wet meadows. Historical vegetation and hydrology data 

are largely unavailable, particularly at broad landscape or regional scales. Anecdotal information such as historical photographs do provide limited evidence of 

change, such as a transition from herbaceous dominated communities to shrub 

dominated ones, but it is not possible to confidently extrapolate to broader areas 

from such limited information.

Wet meadows are highly susceptible to invasion by a variety of non-native species. Although they are among the most diverse and productive ecosystems in 

the region, wet meadows also support many exotics (Stohlgren and others 1997, 

1999). In addition to the accidental and incidental spread of non-native species, 

many wet meadows have been planted to Eurasian pasture grasses, completely 

changing the composition of communities from historical conditions. Poa pratensis has been increasing in western montane meadows, although it is unknown 

whether this is a result of grazing, decreased water tables, or competitive effects 

(Kluse and Allen-Diaz 2005).

In addition to altered species composition, the vegetation structure of many 

wet meadows has been altered due to the invasion of woody species. Tree and 

shrub invasion has increased in mountain meadows in many portions of the west 

within the last century (Miller and Halpern 1998). There has been widespread 

establishment of trees in mountain meadows of the Pacific Northwest, Sierra 

Nevada, and Rocky Mountains (Dyer and Moffett 1999; Jakubos and Romme 

1993; Millar and others 2004; Schauer and others 1998), although there is still 

considerable uncertainty regarding the precise cause. Climate change and grazing have been identified as possible causes. Some research suggests that in the 

absence of disturbance, dense herbs will prevent germination of seeds. Berlow 

and others (2002) concluded that, while sagebrush expansion is traditionally associated with increased meadow aridity, sagebrush exhibits the greatest potential 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 201

for seedling germination, growth, and survival in mesic sites and that potential 

spread is dependent on the confluence of exposed soil, a nearby seed source, and 

reduction of aboveground herb biomass such as through grazing (Berlow and 

others 2002). In contrast, shrub cover in many wet meadows has declined as a 

result of anthropogenic activities. Willows have been removed from many wet 

meadows throughout the West, often to create pasture for livestock (Patterson 

2005; Peinetti 2000).

Wet meadows supported by several aquifers or with large contributing watersheds 

may be less sensitive to climate shifts than wetlands with smaller watersheds or 

that formed in more marginal settings. Indirect and direct changes to wet meadow 

hydrologic processes have occurred in many wetlands throughout the region; 

however, the data are insufficient to quantitatively assess potential changes. 

Where key ecological processes in surrounding uplands have significantly departed from their historical range of variability, effects on wet meadows may be 

expected to occur.

Management opportunities and constraints

Management opportunities—Landscape assessments of wet meadows are 

needed, because of the lack of sufficient data to establish contemporary benchmarks, let alone to make historical reconstructions. Possible data tools include 

aerial photographs or other remotely sensed data to identify and map wetlands. 

Normalized Difference Vegetation Index (NDVI) values derived from satellite 

data hold promise for remote sensing of wet meadows. Researchers in alpine areas 

of Colorado compared NDVI values with snow depth and soil moisture values 

and found a trend of increasing value from barren to fellfield to dry meadow to 

snowbed to moist meadow to wet meadow (Walker and others 1993). Researchers 

in Oregon analyzed the spectral signature of meadows using a temporal series 

of five Landsat thematic mapper (TM) images and found that the sequence of 

scenes provided ample resolution to differentiate most of the meadow plant 

communities identified in previous field studies. The authors suggested that the 

technique also may be useful to detect the location and spread of species with 

distinctive vegetative or floral phenologies (Ager and Owens 2004).

Site-level assessments of wet meadows examining basic hydrology, water chemistry, and vegetation relationships are also needed. Relatively little money or 

time is needed for installation of a few shallow groundwater-monitoring wells. 

Several tools have been developed to assess wetland condition. Methods such 

as the Vegetation Index of Biological Integrity, Floristic Quality Assessment, 

and Ecological Integrity Scorecards vary in detail but all offer potential for 

landscape-scale assessment work (Faber-Langendoen and others 2006; Lopez and 

Fennessy 2002; Rocchio 2006c). In addition to improving our understanding of 

landscape-scale patterns of wet meadow condition, such studies would be useful 

to identify sites for restoration. Also needed are studies evaluating approaches 

for restoring degraded sites.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 202

Restoration—The ecological integrity of many wet meadows in the region has 

been degraded. Whether through hydrologic alteration or invasive species, many 

wetlands no longer provide the ecological functions they once did. Restoration 

provides a means of recouping lost ecological functions. Before conducting a 

wetland restoration project, it is critical to understand site hydrology. Drainage 

tiles, ditches and water diversions need to be identified and removed, as these 

will prevent success in later stages if not addressed first.

For vegetation establishment, it is important to select the proper species since the 

hydrologic preferences of species varies. A variety of approaches are available 

for establishing vegetation, including the use of plugs, bare root stock, or containerized stock (Steed and DeWald 2003; Steed and others 2002). In addition, 

seeds can be collected in the field, stratified in a greenhouse, and germinated 

and grown into seedlings for use in restoration projects. This approach allows 

placement of species directly into the physical environments where they will perform best. The choice of specific methods should be made carefully. Additional 

techniques should be evaluated for vegetation management. For example, fire 

has been used to as a management tool for some wet meadows. Prescribed burns 

can be used to slow woody plant invasion into meadows or to set back invasive 

species (Debenedetti and Parsons 1979; 1984).

Management constraints—Constraints include lack of information regarding 

the extent and condition of wet meadow resources. Political, economic, and 

regulatory issues may also come into play when formulating management plans. 

A long-term issue outside the control of managers is regional climate change. 

Changing temperatures and precipitation patterns predicted under some global 

climate change scenarios (US EPA 1998b, 1998c; Wagner 2003) will likely affect 

the composition and function of wet meadows by shifting the competitive balance 

among species. The boundaries between wet meadows and adjacent ecosystems 

(mesic and dry meadows, uplands, different wetland types) may shift, making it 

difficult to predict net effects on wet meadow area.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 203

9. Conclusions

The ecosystem types forming the basis of this assessment—fens, marshes, wet 

meadows, salt flats, and riparian areas – differ in vegetation composition, key 

functional characteristics, and underlying drivers. Given this natural variability and the large extent of the assessment area, it is not surprising that the data 

needed to quantitatively assess many important metrics are lacking. As a result, 

the value of this assessment lies not in a spatially explicit definition of the HRV 

for any particular part of Region 2, but rather in a conceptual overview of how 

wetland and riparian ecosystems are structured and function and the myriad ways 

Euro-Americans have altered them. The information presented here is best used 

as a basis for developing more specific hypotheses regarding possible changes 

in the HRV at the scale of individual Forests and Grasslands. Although it is possible to make some generalizations, readers are cautioned to evaluate the specific 

circumstances in their particular Forest or district.

At broad scales, climate, hydrology, and geomorphic processes are key factors 

influencing the natural abundance and distribution of wetland and riparian types. 

Wetlands are generally more abundant in cooler and moister climates, where 

groundwater and precipitation are more abundant, and in complex geomorphic 

settings, such as glaciated mountain valleys. Patterns in the general distribution of wetlands are also evident in the distribution of specific wetland types. 

Groundwater-dependent wetlands such as fens and wet meadows are typically 

more abundant at higher elevations, while more hydrologically variable marshes 

and salt flats occur at lower elevations.

Recognizing the functional differences among wetland types is important for 

developing an understanding of how humans may have affected wetlands and 

riparian ecosystems. Hydrologic processes are critically important to wetland 

and riparian areas. As a consequence, when evaluating human impacts, particular attention must be paid to the various ways in which anthropogenic activities 

may affect hydrologic function. For example, where hydrologic modifications 

like dams are present in a watershed, it is likely that downstream riparian areas 

have been altered. Other important anthropogenic impacts are grazing and the 

introduction and spread of exotic species. Both are responsible for changing the 

character of many wetlands and riparian areas in the region.

In preparing of this assessment, we identified several major information and 

research needs. One of the most critical information gaps relates to the condition or integrity of wetlands and riparian areas at landscape and regional scales. 

By leveraging recent advances in technology such as remote sensing and by 

developing improved field evaluation methods such as Indices of Biotic Integrity 

(IBI), it is possible to develop more rigorous and quantitative estimates of the 

condition of wetlands and riparian areas across Region 2 (Faber-Langendoen 

and others 2006; Karr 1999; Rocchio 2006c). Site-level hydrologic data are 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 204

essential for identifying and quantifying possible deviations from historical 

conditions (Richter and others 1998, 2003). Taken together, broad-scaled and 

local assessments can provide a means of identifying impaired ecosystems and 

prioritizing restoration efforts.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-286WWW. 2013 205

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