United States

Department

of Agriculture

Forest Service

Rocky Mountain

Research Station

General Technical

Report RMRS-GTR-139

September 2005

Historic Range of Variability for Upland 

Vegetation in the Medicine Bow 

National Forest, Wyoming

Gregory K. Dillon

Dennis H. Knight

Carolyn B. Meyer
Dillon, Gregory K.; Knight, Dennis H.; Meyer, Carolyn B. 2005. Historic range of variability for upland 

vegetation in the Medicine Bow National Forest, Wyoming. Gen. Tech. Rep. RMRS-GTR-139. Fort Collins, CO: 

U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 85 p.

Abstract

An approach for synthesizing the results of ecological research pertinent to land management is the analysis of the 

historic range of variability (HRV) for key ecosystem variables that are affected by management activities. This report 

provides an HRV analysis for the upland vegetation of the Medicine Bow National Forest in southeastern Wyoming. 

The variables include live tree density, dead tree (snag) density, canopy cover, abundance of coarse woody debris, 

species diversity, fire return intervals, the abundance of various diseases, the proportion of the landscape in different land cover types, and the degree of patchiness in the landscape. The variables were examined at the stand and 

landscape scales, using information available in the literature and USFS databases. High-elevation landscapes were 

considered separately from low-elevation landscapes. Much of the report pertains to forests dominated by lodgepole 

pine, subalpine fir, and Engelmann spruce at high elevations, and by ponderosa pine and aspen at lower elevations. 

We defined the HRV reference period for the MBNF as approximately 1600 to 1860.

Keywords: aspen, bark beetles, canopy gaps, coarse woody debris, comandra blister rust, dwarf mistletoe, 

Engelmann spruce, fire effects, disease in Rocky Mountain forests, forest fragmentation, insects in forests, interior forest, landscape ecology, lodgepole pine, mistletoe, mountain pine beetle, ponderosa pine, roads, Rocky 

Mountain forests, snags, spruce beetle, subalpine fir, subalpine forests, timber harvesting effects, white pine blister rust, wind in forests

The Authors

Gregory K. Dillon is a GIS specialist with the Forest Service at the Rocky Mountain Research Station, Fire Sciences 

Laboratory, Missoula, MT. His current work is focused on predictive mapping of biophysical settings in the United 

States for the LANDFIRE project. Previously, he did a variety of forest ecology and spatial analysis work at the 

University of Wyoming and for the Forest Service in the Southern Appalachians and Pacific Northwest. He received 

a B.S. degree in geography from James Madison University, Harrisonburg, Virginia, in 1995, and an M.A. degree in 

geography from the University of Wyoming in 1998.

Dennis H. Knight is professor emeritus in the Department of Botany at the University of Wyoming, Laramie, WY. 

He retired in 2001 after 35 years on the faculty, where he taught courses in ecology and forest management. Much 

of his research focused on the effects of natural and human-caused disturbances on Rocky Mountain coniferous forests. In 1999 he contracted with the Rocky Mountain Regional Office of the Forest Service to prepare historic range 

of variability analyses for the upland vegetation of the Medicine Bow, Bighorn, and Shoshone National Forests in 

Wyoming. He received a bachelors degree in biology from Augustana College, Sioux Falls, South Dakota in 1959, 

and the Ph.D. in plant ecology from the University of Wisconsin-Madison in 1964.

Carolyn B. Meyer is on the faculty of the Department of Botany at the University of Wyoming, Laramie, WY. Much 

of her research focuses on the study of spatial and temporal variation of ecological phenomena. She received a B.S. 

degree in natural resources from Humboldt State University, California in 1983, an M.S. degree in wildlife biology 

from the University of Wyoming in 1985, and a Ph.D. in ecology from the University of Wyoming in 1999. Prior to 

joining the faculty at Wyoming, she held research and management positions with the U.S. Fish and Wildlife Service 

and the National Park Service.

Cover photo: Wyoming-Colorado border south of Encampment, photo by Lane Eskew.

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Executive Summary

The challenges of sustainable land management have led to an increased emphasis on incorporating the results of 

science in the decision making process. One approach for accomplishing this objective is through the analysis of 

the historic range of variability (HRV) for key ecosystem variables that are affected by management activities. The 

rationale for HRV analyses is that the chances of sustainable forest management are greater if the variation in managed ecosystems includes the range of conditions that are expected at various scales in ecosystems relatively uninfluenced by humans. This report provides an HRV analysis for the upland vegetation of the Medicine Bow National 

Forest (MBNF).

By definition, HRV analyses require the identification of specific variables and an estimation of how those variables fluctuated, at more than one scale, prior to the advent of resource extraction and management by EuropeanAmericans. A complete list of the variables that we examined is found in table 7. Examples include live tree density, dead tree (snag) density, canopy cover, abundance of coarse woody debris, species diversity, fire return intervals, 

the abundance of various diseases, the proportion of the landscape in different land cover types, and the degree of 

patchiness in the landscape. We examined variables at two scales—the stand and the landscape—and we separated high-elevation landscapes from low-elevation landscapes. Much of the report pertains to forests dominated by 

lodgepole pine, Engelmann spruce, and subalpine fir at high elevations, and ponderosa pine and aspen at lower elevations. Our analysis emphasizes forests rather than grasslands and shrublands because more historical information is available for the forests.

Two significant challenges for HRV analyses are 1) selecting a “reference period” for comparison to present day conditions, and 2) making decisions with very little data about range of variability during the reference period. We defined the HRV reference period for the MBNF as approximately 1600 to 1850, as some data sets on fire history go 

back several centuries and the influence of European-Americans in the MBNF was minimal until about 1850. We 

also compared MBNF ecosystems to comparable ecosystems in natural areas relatively uninfluenced by management activities, such as Yellowstone National Park. When data were lacking, our approach was to make plausible 

though qualitative judgments about the conditions that must have existed in the reference period based on current 

knowledge about plant adaptations and ecosystem structure and function. This approach is an example of deductive science. There will always be some uncertainty in such conclusions, but our conclusions were evaluated and 

supported by a panel of four anonymous peer reviewers. For each of our conclusions about whether a stand or landscape variable is within the HRV or trending away from that norm in managed landscapes, and the Forest as a whole, 

we indicate if our confidence level is low, moderate or high.

Our report describes how the forests of the MBNF have evolved with regular disturbances, and how the kinds and 

frequency of disturbances have changed in some areas due to modern management practices. Over half of the forested area has been harvested for timber since the late 1800s (high- and low-elevation landscapes combined), and 

approximately 75% of the forests have been subjected to both harvesting and burning since that time.

With the information available, we have attempted to draw conclusions about whether or not certain ecosystem 

variables are now beyond their HRV for the period 1600-1850 or trending in that direction (table 7). To summarize, 

we believe there is evidence to suggest that the following forest variables are within the HRV at high-elevations:

• Extensive, stand-replacing fires

• Insect population sizes

• Abundance of diseases (including dwarf mistletoe in many areas)

• Density of trees in all size classes (though the mean density may have been lowered in some areas)

• Tree and understory plant cover and diversity (number of species)

• Root/shoot ratios

• Forest floor depth

• Proportion of the landscape in different cover types

• Ratio of forest to non-forest land

Similarly, the following forest variables appear to be within the HRV at lower elevations:

• Aspen stand abundance and stand structure

• Abundance of native insects and diseases

• Understory plant species diversity (number of species)

i
In contrast, we believe the following ecosystem variables are either beyond their HRV now or could be in the future 

if current management practices continue:

Within stands at high elevations affected by timber harvest

• Tree cover probably is lower and the number and size of canopy gaps probably is higher in harvested stands 

than the range of means (see text) for these variables during the HRV period in unmanaged forests of comparable age and site conditions (moderate confidence), due primarily to selective and shelterwood cuts but 

also due to fewer standing-dead trees.

• Where timber harvesting has occurred on the MBNF, snag density and the amount of coarse woody debris 

is lower than the range of means for these variables during the HRV reference period for unmanaged stands 

of comparable age and site conditions (high confidence). Whole-tree yarding can accentuate this variation 

by concentrating coarse woody debris and slash at landings in a way that has not occurred previously.

• Due to some methods of slash treatment, the intensity of soil scarification is beyond the HRV in some areas 

(high confidence). Burning can remove the forest floor, but the depth of soil disturbance after fire typically 

is less than disturbance following mechanical harvesting.

• Age- and size-structure of managed stands within ravines or on north or leeward slopes is skewed toward 

smaller and younger trees than would have occurred during the HRV period, with less old forest, because 

of harvesting where fires would have been less likely to burn (moderate confidence).

• Dwarf mistletoe abundance could be above its HRV in stands subjected to fire suppression and to partial 

timber harvests that did not remove infected trees (moderate confidence).

Across the landscape at high elevations

• Due to the timber harvests and human-caused fires in the late 1800s and early 1900s, the proportion of land 

in middle-aged forest (60-150 years old) is above the HRV (moderate confidence). Similarly, old-growth 

forest is less common on the landscape, especially in areas that historically had long fire-free intervals, 

such as in valley bottoms and on some north or leeward slopes (moderate confidence). Thus, ecosystem 

characteristics such as snags and coarse woody debris that depend on long periods of forest development 

may be less common across the landscape than during the reference period (high confidence).

• The estimated rotation times for timber harvest in forests judged suitable for harvesting are generally shorter, on average, than the estimated site-specific mean intervals between natural disturbances in high-elevation forests. As a result, the rate at which new patches are formed across the landscape probably exceeds 

the HRV (high confidence).

• Characteristics of the size and spatial pattern of disturbances and forest patches across high elevation landscapes exceed their HRVs. Timber harvest typically produces mid-sized patches (i.e., 1-20 ha) that have 

less variability in size and shape than patches created by natural fires. The result is a decline in average 

patch size, the development of a more simple patch shape, less connectivity among patches of old forest, 

and a higher, more uniform level of landscape patchiness than would have occurred during the reference 

period (high confidence).

• Due to roads and timber harvesting, the amount of edge is now higher and the land area of interior forests 

is lower, on average, than prior to the 1900s (high confidence).

• The proportion of forested land with high tree density is somewhat lower because of partial thinning designed to maximize rates of tree growth (low confidence).

Within stands at low elevations affected by harvest, fire suppression, and livestock grazing

• Fire suppression, livestock grazing, timber harvesting, and possibly climate change have led to the development of low-elevation forest stands with high densities of young trees, greater canopy cover, and less understory than usually occurred on many sites before 1850 (moderate confidence). Understory composition 

has also changed (low confidence) and forest floor depth has increased (high confidence).

• By removing large trees and downed wood from the system, wood harvesting has lowered the abundance 

of snags and coarse woody debris in managed stands of low-elevation forests (moderate confidence). 

Harvesting in recent decades, where it has occurred, also has led to an increase in sapling density and a 

reduction in the average diameter of trees, thus creating younger, more uniform-sized stands than existed before harvest (moderate confidence). Much of the harvesting that affected low-elevation forests on the 

MBNF occurred in the late 1800s and early 1900s.

• Mean fire return intervals have become longer than the HRV at low elevations on drier sites, but fire intensity is probably higher due to the amount and continuity of fuels that have developed in some areas, and 

which could lead to more stand-replacing fires in the future (high confidence).

• Forest floor depth may be deeper than the HRV at low elevations where fire suppression has been successful and prescribed burning has not been done (high confidence).

• White pine blister rust outbreaks exceed the HRV in affected stands, as this disease is not native (high confidence).

ii
Across the landscape at low elevations

• Fire suppression, livestock grazing, and the logging of older, more fire resistant trees, where this has occurred at low elevations, probably has increased the amount and continuity of fuels to a point that is above 

the HRV, leading to potentially more stand-replacing fires than occurred prior to 1850 (moderate confidence). Climate warming may contribute to this trend.

• Fire suppression has reduced the level of interspersion of tree stands with grasslands (moderate confidence), 

and the average tree and sapling density has increased above the HRV (moderate confidence).

Aspen forests

• In some aspen stands, grazing by livestock and native ungulates has reduced aspen densities and there has 

been an increase in the abundance of non-native species in the understory, to the point where such variables are beyond their HRVs (moderate confidence). Some aspen stands now have an understory of conifers, but this successional pattern probably occurred also during the reference period after long periods 

without fire (moderate confidence).

Non-forest vegetation

• Some meadows, grasslands, and shrublands are probably grazed more, or were in the 1900s, than during 

the HRV period. Reductions in livestock numbers since the early 1900s have been dramatic and rangeland 

condition is improving. Where heavy grazing occurred, some species probably declined in abundance, 

some probably became more abundant, and soil erosion probably occurred more rapidly (high confidence). Deviations from the HRV also have been caused by the invasion of exotic plants that displace native species and cause changes in other ecosystem properties, but thus far such effects are not well known 

on the MBNF. In general, insufficient information is available for drawing conclusions about current HRV 

deviations in non-forest vegetation.

Frequently there has been a lack of pertinent information for the ecosystem variables we addressed in this report. 

Such variables can be identified as priorities for research, especially if they appear to have exceeded their HRV for 

the reference period. Research will also be required if decisions about the severity of a problem cannot be made 

with existing information or if ways of restoring the HRV for a specific variable cannot be identified.

iii
Acknowledgments

Funding was provided by the Division of Renewable Resources, Rocky Mountain Region, U.S. Forest Service (USFS 

Agreement 1102-0003-98-043 with the University of Wyoming). Our analysis and conclusions were strengthened 

by the assistance of the following scientists and managers, and also by the comments of five peer reviewers coordinated by the Ecological Society of America.

• Claudia Regan, Regional Office, U.S. Forest Service, and

• Department of Forest Sciences, Colorado State University

• Dee Hines, Supervisor’s Office, Medicine Bow National Forest

• David Harris, Supervisor’s Office, MBNF

• Phil Krueger, Supervisor’s Office, MBNF

• Carol Tolbert, Supervisor’s Office, MBNF

• Deana Wood, Supervisor’s Office, MBNF

• Robert Mountain, Supervisor’s Office, MBNF

• Barry Parish, Regional Office, U.S. Forest Service

• Greg Hayward, Regional Office, U.S. Forest Service, and Department of Zoology and Physiology, University of 

Wyoming

• Bruce Short, Regional Office, U.S. Forest Service

• Michael Sanders, Laramie Ranger District, MBNF

• Jerod R. DeLay, Laramie Ranger District, MBNF

• Jerry Mastel, Brush Creek District, MBNF

• Joel Harper, Supervisor’s Office, MBNF

• Clarke McClung, Supervisor’s Office, MBNF

• Pat Harrison, Supervisor’s Office, MBNF

• Timothy W. Byer, Supervisor’s Office, MBNF

• Michael J. “Mick” Hood, Supervisor’s Office, MBNF

• George Wiggins, Supervisor’s Office, MBNF

• Tom Troxel, Intermountain Forest Association

• Don Heiser, retired, Supervisor’s Office, MBNF

• Howard Pickerd, Wyoming State Forester’s Office

• Judy von Ahlefeldt, retired, Supervisor’s Office, MBNF

• Kathleen Roche, Supervisor’s Office, MBNF

• Thomas T. Veblen, Department of Geography, University of Colorado

• William H. Romme, Department of Forest Sciences, Colorado State University

• William L. Baker, Department of Geography, University of Wyoming

• Jason Sibold, Department of Geography, University of Colorado

• Stephen T. Jackson, Department of Botany, University of Wyoming

• Stephen T. Gray, Department of Botany, University of Wyoming

• Christopher L. Fastie, Department of Biology, Middlebury College

iv
Contents

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

2. Methods ..........................................................................................................................4

3. The Medicine Bow National Forest .................................................................................6

3.1 Physical setting ...........................................................................................................6

3.2 Ecological setting .......................................................................................................8

3.3 Cultural setting ......................................................................... 13

4. The HRV of High-Elevation Forest ................................................................................. 18

4.1 Variation in species composition and succession ....................................................... 18

4.2 Disturbances in high-elevation forests .......................................................................22

4.2.1 Fire ....................................................................................................................24

4.2.2 Insects ..............................................................................................................28

4.2.3 Disease .............................................................................................................30

4.2.4 Wind .................................................................................................................32

4.2.5 Timber harvesting ..............................................................................................33

4.3 Stand structure of high-elevation forests ....................................................................40

4.3.1 Tree density ....................................................................................................... 41

4.3.2 Seedling and sapling density .............................................................................41

4.3.3 Percent canopy cover ........................................................................................ 41

4.3.4 Density and size of canopy gaps ........................................................................42

4.3.5 Density and cover of understory plants ..............................................................42

4.3.6 Ratio of aboveground biomass to belowground biomass ....................................42

4.3.7 Tree species and genetic diversity ......................................................................42

4.3.8 Total plant species diversity ...............................................................................42

4.3.9 Age-class structure of stands ..............................................................................43

4.3.10 Size-class structure of stands ............................................................................43

4.3.11 Forest floor depth .............................................................................................43

4.3.12 Mineral soil affected ........................................................................................44

4.3.13 Snag density ....................................................................................................44

4.3.14 Coarse woody debris .......................................................................................44

4.4 Landscape structure at high elevations ......................................................................45

4.4.1 Landscape diversity (number and proportion of land-cover types) ......................48

4.4.2 Proportion of landscape in different stand-age classes ........................................48

4.4.3 Edge, interior forest habitat, and patchiness ........................................................50

4.4.4 Ratio of forest to non-forest land ........................................................................50

4.4.5 Rate at which new patches are formed ...............................................................50

4.4.6 Proportion of land with high snag and CWD densities .......................................50

4.4.7 Proportion of land with high tree density ...........................................................50

4.5 Summary for high-elevation forests ...........................................................................50

5. The HRV of Low-Elevation Forests and Woodlands .......................................................51

5.1 Disturbances at low elevations ..................................................................................52

5.1.1 Fire ....................................................................................................................53

5.1.2 Insects ...............................................................................................................58

5.1.3 Disease ..............................................................................................................59

5.1.4 Wind .................................................................................................................60

5.1.5 Timber harvest ...................................................................................................60

v
5.1.6 Livestock grazing ............................................................................................... 61

5.2 Stand structure at low-elevations .............................................................................. 61

5.2.1 Tree density, sapling density, canopy cover, age-class 

 structure, and size-class structure ............................................................................62

5.2.2 Spatial distribution of trees .................................................................................62

5.2.3 Snags and coarse woody debris .........................................................................62

5.2.4 Forest floor depth ..............................................................................................63

5.2.5 Understory plant cover, species composition, and diversity ................................63

5.3 Landscape structure at low-elevations .......................................................................63

5.4 Summary for low-elevation forests and woodlands ....................................................64

6. The HRV of Aspen Forests and Woodlands ...................................................................64

6.1 Disturbance history, human influences and succession ..............................................66

7. The HRV of Non-Forest Vegetation ...............................................................................67

7.1 The effects of disturbances on non-forest vegetation ..................................................69

8. Summary of Probable HRV Deviations ..........................................................................70

9. Literature Cited ............................................................................................................. 74

vi
USDA Forest Service RMRS-GTR-139. 2005 1

1. Introduction

Because of new information, changing public attitudes, and various congressional mandates pertaining to 

environmental protection and biodiversity, land managers now face the challenges of ecosystem management 

(Keiter 1989; Kaufmann and others 1994; Grumbine 

1994; Covington and DeBano 1994; Christensen and 

others 1996; Vogt and others 1997; Kaufmann and others 1998; Landres and others 1999). Ecosystem management seeks to maintain the natural diversity of both 

species and ecosystems while providing resources desired by humans (Grumbine 1994). Proponents of this 

approach recognize that our current knowledge of ecosystems is incomplete (Christensen and others 1996), 

but they also contend that resource managers should use 

available scientific information more fully. Also, as we 

learn more about the effects of certain activities, management strategies can change through a process known as 

adaptive management (Walters 1986; Christensen 1996). 

One approach for synthesizing ecological information 

for an area is through analyses of the historic range of 

variability (HRV) for key ecosystem variables, which is 

what we have done in this report for the Medicine Bow 

National Forest (MBNF) in southeastern Wyoming.

An example of how resource management attitudes 

have changed in the last 50 years is the replacement of 

the word “stability” with “resilience” (Holling and Meffe 

1996). Scientists and managers now recognize that ecosystems have developed with periodic disturbances such 

as fires and insect epidemics (Botkin 1990; Landres 

and others 1999; and others), and that many ecosystems are in a state of recovery or change, not stability. 

Disturbances are believed to have perpetuated or even 

enhanced the biological diversity that many now appreciate (Attiwill 1994). In outlining the conceptual framework for ecosystem management for the Forest Service, 

Kaufmann and others (1994) concluded that “ . . . the 

most sound basis for ecosystem management is to assure 

that the variation characterizing [managed] ecosystems 

includes the range of conditions that are expected at various scales in ecosystems uninfluenced by humans.” The 

focus is on variability.

Various terms have been used when discussing variability, including “range of natural variability” (Swanson 

and others 1993), “natural variation” (Holling and Meffe 

1996), “reference conditions” (Kaufmann and others 1994), “reference variability” (Manley and others 

1995), and “historic range of variability” (Morgan and 

others 1994). We adopted “historic range of variability” because it avoids the potential problems associated 

with interpreting the word “natural.” For this report, we 

interpret the historic range of variability (HRV) as the 

spatial and temporal variation in composition, structure and function experienced in an ecosystem from 

about 1600 to 1850, when the influences of EuropeanAmericans were minimal on the MBNF.

A challenge for HRV analyses is to draw conclusions 

or inferences based on very little data. Several approaches are used. First, biologists and ecologists often are able 

to make defensible judgments about the conditions that 

likely existed in the past based on studies of plant adaptations. Thus, knowledge about the organisms present in 

an ecosystem today provides clues about environmental conditions that existed in the past. Secondly, judgments about the past can be made from different kinds of 

historical records, such as the journals and photographs 

of early explorers and the information contained in tree 

rings. Finally, valuable information is obtained by examining modern reference areas that have not been heavily 

influenced by European-Americans, such as wilderness 

areas and some National Parks. In fact, no landscape is 

entirely free from human influences. Native Americans 

have caused some changes for thousands of years, and 

even the largest National Parks have been affected by 

technological developments.

HRV analyses require careful attention to four 

principles. First, variables should be evaluated over 

at least two spatial scales. The variability that occurs in 

small tracts of land, commonly referred to as “stands,” is 

quite different from that which occurs over larger areas, 

referred to as “landscapes.” Stands are tracts of forest or 

other vegetation types that are relatively homogeneous 

in terms of environment, plant species composition, and 

disturbance history. They generally occupy an area ranging from a few hectares to several hundred hectares. In 

contrast, landscapes range from several square kilometers to several thousand square kilometers, or more, and 

they have considerable heterogeneity. The HRV of some 

variables, such as dead standing trees (snags), is typically higher at the stand scale than at the landscape scale 

(fig. 1). Of course, some variables pertain to only one of 

the two scales. For example, the proportion of a landscape in different cover types is not meaningful at the 

stand scale.

Pertinent to considerations of scale is the fact that, 

just as there will be several to numerous stands in a landscape, there are several landscape types in a National 

Forest. For this report, we separated high-elevation landscapes from low-elevation landscapes, with much of our 

discussion focused on forests and woodlands dominated 

by lodgepole pine (Pinus contorta var. latifolia), subalpine fir (Abies lasiocarpa), and Engelmann spruce 

(Picea engelmannii) at high elevations; and ponderosa 
2 USDA Forest Service RMRS-GTR-139. 2005

pine (Pinus ponderosa), limber pine (Pinus flexilis), and 

Douglas-fir (Pseudotsuga menziesii) at lower elevations. 

We also present discussions of aspen (Populus tremuloides) woodlands and several types of shrublands, grasslands, and forblands. Forests are emphasized because 

much more historical information is available for them 

than for other vegetation types. Wetlands, including the 

riparian zone, are mentioned only briefly in our report.

The second principle is that HRV analyses should 

consider variability in space as well as time. This principle is difficult to follow because historical data are rarely available for more than one watershed or landscape. 

However, it is possible to characterize spatial variability 

in vegetation patterns across management units on the 

MBNF, as it exists today, and to compare that variability to the range of expected conditions during the HRV 

period based on our current understanding of ecosystem 

dynamics. By necessity, HRV analyses must often be deductive in nature—drawing conclusions about specific 

ecosystem variables based on general knowledge about 

ecosystems.

The third principle is that “variability” can be characterized in several ways. To illustrate, variability can be 

thought of as the absolute range of an ecosystem variable or parameter, where extreme and rare conditions are 

considered (fig. 2, top). A second approach involves the 

calculation of the standard deviation, though we have 

not found data to calculate this statistic for any variable 

on the MBNF. Alternatively, the HRV can be based on 

an estimate of the range of means for a variable over 

several consecutive periods of time, an approach that ignores extremes (fig. 2, center). In this report, we use the 

range of means approach—conceptually if not quantitatively. HRV analyses based on average conditions tend 

to produce a narrower range than when based on extremes (fig. 1).

The fourth principle affecting HRV analyses pertains to climate change through time. There is no question that the climate has changed since Rocky Mountain 

glaciers began to recede about 10,000 years ago. It also 

is clear that the climate of the 1600s, 1700s, and much of 

the 1800s was different than at the present time (Millar 

and Woolfenden 1999). Working in the Bighorn Basin of 

Wyoming, Gray and others (2004) concluded from tree 

ring data that, while the mean annual precipitation was 

about the same during the 1600s and 1700s as it is today, 

there are now fewer extended wet and dry periods and 

the extremes of the wet and dry periods in the 1900s are 

less. Moreover, research suggests that the climate of the 

region began to warm in the late 1800s (Gray and others 

2004). While the HRVs of some ecosystem variables for 

the reference period 1600 to 1850 may be less applicable 

to present conditions because the climate has changed, 

we believe they still provide a useful context for management decisions. Distinguishing the effects of climate 

change from the effects of European-Americans is a difficult but important challenge for HRV reports.

To characterize the HRV, the historic time period 

must also be specified. Manley and others (1995) indicated that, ideally, the period should cover an evolutionary time scale (i.e., as long as possible within the past 

10,000 years or so). However, glaciation and the magnitude of climate change in the MBNF during the last 

10,000 years created extremely broad ranges of variability of little relevance to the kinds of decisions faced by 

forest managers today. We selected 1600 to 1850 as the 

reference period, as it reflects conditions prior to the substantial influences of European-Americans over a time 

scale relevant to forest management. This time scale of 

250 years is also relevant to the life span of dominant 

species on the MBNF; HRV analyses in other ecosystems may require a longer or shorter reference period 

based on the species present.

An evolutionary perspective is adopted in our report, 

but in a different sense than implied by Manley and others (1995). Some of our interpretations are based on the 

prevailing consensus about the environmental conditions 

that must have led to the development of certain adaptations exhibited by the dominant plants of the MBNF. 

Such adaptations surely influence where species survive 

today. For example, the thick bark of ponderosa pine 

Figure 1—Illustration of the expected decline in the HRV as the 

spatial scale increases. For example, the number of dead 

trees (snags) in a single stand could range from zero to thousands over a two hundred year period, but almost certainly dead trees would not decline to zero over a larger area 

because the synchronous death of all the trees is much less 

probable. This figure also illustrates how the HRV, estimated 

using the range of means (defined in figure 2 and the text), is 

always less than the absolute range between the maximum 

and minimum for the entire period.
USDA Forest Service RMRS-GTR-139. 2005 3

(Pinus ponderosa) suggests that the environment where 

it occurs, or has occurred in the past, is characterized by 

periodic surface fires. Similarly, the thin bark and serotinous cones of lodgepole pine (Pinus contorta) suggest 

that the species has been subjected to periodic stand-replacing crown fires.

Estimates of HRV are inevitably limited by our ability 

to interpret the nature of past ecosystems and determine 

what was “natural” or “normal” during the reference 

period (Swanson and others 1993; Holling and Meffe 

1996). Because ecosystems change dramatically over 

time scales of a thousand years or more, or even a few 

centuries, we only capture a “snapshot in time” if our 

reference period is too short (Swanson and others 1993). 

For such reasons, estimates of HRV are not a panacea. 

Rather, they serve as one of several tools to aid managers 

in achieving the societal goal of conserving biological 

diversity while providing the benefits of various natural 

resources in a sustainable way.

To be useful to land managers, HRV analyses should 

attempt to include quantitative ranges for actual variables related to ecosystem structure, composition, or 

function (Manley and others 1995). By considering 

meaningful variables during defined time periods, land 

managers have an improved context in which to evaluate their decisions. If a management activity appears to 

0

5

10

15

HRV

Existing conditions (post-European settlement) 

0

5

10

15 X =5.9

SD=3.9

Period 1

HRV

 

 

Decade

1700

1750

1800

1850

1900

19502000

0

5

10

15

Existing

Conditions

HRV 

Period

X = 3.3 X = 0.4

X =0.6

SD=0.3

Period 1

EC

X =1.3

SD=0.9

Period 3

HRV

X =1.6

SD=1.1

Period 2

HRV

X =0.5

SD=0.4

Period 3

EC

X =0.2

SD=0.2

Period 2

EC

Stand or Landscape Variable

Figure 2—Three approaches for estimating the historic range of variability 

(HRV). Top: If HRV is defined as the 

absolute range for a variable, then the 

range of existing conditions (EC) in the 

example (0-0.8) falls within the HRV 

(0-11). Center: If HRV and EC are defined as the range of means ( x ) for 

a variable over consecutive time periods, an approach which ignores extremes, the EC (0.2-0.6) falls outside 

the HRV (1.3-5.9). If HRV is the range 

of standard deviations of the variable, 

the EC (0.2-0.4) falls outside the standard deviation for the historic conditions (0.9-3.9). We used the range of 

means as the conceptual definition of 

HRV in this report. Bottom: Only one 

mean usually is reported for the historic period, and thus, no range of means 

can be determined. In such cases, to 

be considered a deviation from the 

HRV, a mean for existing conditions 

must be substantially higher or lower 

than the HRV mean. In this example, 

the HRV would have to be less than 

± 80% of the historic mean of 3.3 to 

have the EC mean (0.4) still fall outside the HRV (1.3-5.9), using the same 

data as in the center graph.
4 USDA Forest Service RMRS-GTR-139. 2005

push an ecosystem variable to a higher or lower level, 

based on the kind of evidence included in HRV reports 

such as this one, then there is reason to evaluate the environmental impacts of that activity—or even consider 

adopting a different approach. When quantitative data 

are not available, the goal of specifying variability must 

be done using qualitative information. For some observers, this quasi-quantitative approach is not satisfactory. 

However, deductive analyses can provide insights that 

are not available in any other way and which improve 

the chance that management decisions will not impede 

an ecosystem’s inherent resilience following the inevitable disturbances that will occur, whether caused by fires, 

wind, insects, pathogens, or humans.

Landres and others (1999) concluded that, properly 

done, HRV analyses provide context and guidance for 

managing ecological systems. They also emphasized, 

however, that it is appropriate and unavoidable for HRV 

reports to include professional judgments, even though 

based on limited data. Given that fact, they recommended that the authors of such reports should specify their 

assumptions and value judgments. Briefly, our assumptions and judgments are: (1) there is much yet to be 

learned about how to manage National Forests to meet 

the various mandates that have led to the emergence of 

ecosystem management, (2) there is value in the periodic evaluation of traditional management approaches to 

determine if they are appropriate for meeting an objective, (3) some parts of a National Forest are suitable for 

the careful production of commodities, while others are 

not—a policy that has been implemented for many years 

through the designation of, for example, lands suitable 

and not suitable for timber harvesting, and (4) the information synthesized in this report, in conjunction with 

site specific information and new studies as they become 

available, can provide guidance in making land management decisions.

HRV analyses also should not be viewed as a step toward creating “pristine” conditions (Swanson and others 

1993; Allen and Hoekstra 1994). It has been argued, in 

fact, that ecosystems may be sustainable outside the HRV 

of a variable if humans are willing to engineer functions 

once carried out naturally (Allen and Hoekstra 1994). 

The question remains, however, whether humans know 

enough to create and finance substitutes for natural processes that now are accomplished by hundreds of species that occupy ecosystems over large areas. Notably, 

National Forests already have some features that are unprecedented and clearly outside the HRV, such as roads 

and impoundments (Swanson and others 1993).

In this report for the MBNF, we have attempted to 

provide (1) an estimate of the HRV for key elements 

of terrestrial ecosystems, including quantitative ranges 

for specific variables whenever possible; and (2) a comparison of our HRV analysis with current conditions 

on the MBNF or conditions that have prevailed since 

European-Americans arrived. Our focus is on upland 

vegetation dynamics and plant-related ecosystem variables at the stand and landscape scales. After describing our methods, we review the physical, ecological, and 

cultural characteristics of the MBNF before proceeding 

with an HRV analysis for key variables in high-elevation landscapes, low-elevation landscapes, aspen forest, 

and non-forest vegetation. Our conclusions about apparent human-caused deviations from the HRV are summarized in the last section.

2. Methods

Estimating the HRV for biological systems should include the following steps (Manley and others 1995): (1) 

determine key ecosystem elements (i.e., components, 

structures, and processes); (2) identify measurable environmental indicators for these elements (e.g., tree 

density and mean disturbance return interval); and (3) 

estimate values for the indicators over the selected HRV 

reference period. We tried to be more inclusive than exclusive in selecting key elements, as the importance of 

some ecosystem features or processes may not become 

obvious until the ecosystem is subjected to some kind of 

stress (Holling and Meffe 1996). HRV studies are a relatively new endeavor and methods are changing rapidly 

as new information becomes available.

Our approach was to first make plausible judgments 

about the conditions that must have existed in the past 

based on current knowledge about ecosystem structure 

and function. Knowledge about the organisms that exist in an ecosystem today provides clues about previous 

environmental conditions, and the general principles 

of ecosystem structure and function provide a basis for 

drawing tentative conclusions, deductively, about a specific area. Such judgments and conclusions should be 

considered as hypotheses requiring further testing.

Second, we made judgments about the past from different kinds of historical records, such as those contained 

in tree rings and the journals and photographs of early 

explorers. Specifically, we used (1) historical records 

and studies covering approximately the last century 

that provided quantitative or semi-quantitative information; (2) historical accounts, which provided mostly 

qualitative information back to the late 1800s; and (3) 

reconstructive studies from tree rings, pollen analysis, 

and other methods that provide the only quantitative 
USDA Forest Service RMRS-GTR-139. 2005 5

information for periods prior to 1850 (Kaufmann and 

others 1998).

Third, valuable information was obtained from studies on the long-term dynamics of ecosystems in modern reference areas that have not been heavily influenced 

by humans, such as wilderness areas and some National 

Parks in the region. Yellowstone National Park (YNP) 

has been especially useful because of the research that 

has been done there on fire/vegetation interactions. The 

climate of YNP and the Medicine Bow National Forest 

are similar, though YNP is generally slightly cooler and 

more moist (based on data from Martner 1986). The 

geology of the two areas is different but the predominant vegetation is similar. Using data from any area as a 

proxy for another must be done with caution, but to ignore such pertinent information needlessly limits the applicability of research.

Our analyses are based primarily on the compilation 

and synthesis of information from published and unpublished sources relating to the history and ecology of the 

MBNF specifically and the Rocky Mountain region in 

general. Funds were inadequate for initiating new field 

work. Information was derived from all available sources, including journals, books, USDA Forest Service publications, MBNF reports, and MBNF databases. Much 

of the spatial and quantitative information for the MBNF 

is from the Resource Information System (RIS), a database maintained at the Supervisor’s Office in Laramie 

(USDA Forest Service 1989). There are some limitations to the RIS data, but this source provided the most 

comprehensive spatial information. Of particular concern are the stand-age data, which were determined with 

small sample sizes.

Several USFS reports were very useful. Specifically, 

the 1996 specialist report on biophysical and historical 

aspects of species and ecosystems on the MBNF, prepared by former MBNF employees Judy von Ahlefeldt 

and Clay Speas, was a valuable source of information. They summarized data from the RIS database, 

other internal MBNF documents (e.g., fire records), 

and the Medicine Bow Collection at the University of 

Wyoming’s American Heritage Center in a way that had 

not been done previously. Chapters in their report on the 

geographic and climatic setting, historical ecology, and 

current terrestrial vegetation were especially helpful.

Another noteworthy document was the 1986 report 

by Dennis N. Grasso, Scott Thybony, Robert Rosenberg, 

and Elizabeth Rosenberg. They wrote a cultural history 

for the MBNF, which later was condensed into a book, 

The Medicine Bows (Thybony and others 1986). More recently, M. Boglioli, K. Hejtmanek, and N. Stroup (2002) 

completed a report for the Forest Service on the history 

of land use on the Medicine Bow National Forest.

Also, several HRV assessments had been drafted for 

other National Forests in the Rocky Mountain Region 

when we began our work in 1999 (including the Routt, 

Rio Grande, San Isabel, and White River National 

Forests), and later Veblen and Brown (2001) provided an 

HRV assessment for the Arapaho and Roosevelt National 

Forests. These reports were helpful as well. Our report, 

however, is the first attempt for the MBNF.

For this report, we have drawn conclusions about 

whether existing conditions in terrestrial ecosystems 

have changed from conditions during the HRV reference period (1600 to 1850) due to the management activities of European-Americans. We defined HRV as 

the range of means of a variable for several consecutive 

time periods from 1600 to 1850 (fig. 2, middle). In practice, rarely are the data adequate to calculate the HRV in 

this way. To partially offset this deficiency, “proxy” data 

were sometimes used, as follows:

(1) We reviewed information for the HRV period resulting from studies in other places in the Rocky 

Mountains.

(2) We used the existing pattern (20th century) of “natural” disturbances, such as blowdowns and standreplacing fires, as the pattern that must have existed during the HRV period for comparison to patterns created by new kinds of disturbances such 

as extensive timber harvesting. Recognizing that 

humans have surely modified some of the natural disturbances (e.g., by suppressing some fires), 

we only used this method when the human modification of natural disturbance patterns was thought 

to be small relative to the landscape modifications 

caused by European-Americans.

(3) We assumed that existing conditions (20th century) 

in Yellowstone National Park were approximately 

the same as the conditions on the MBNF during the 

HRV period. The Park’s vegetation has been less 

strongly influenced by European-American activities and may reflect historical patterns that occurred 

in the Medicine Bow Mountains. In some cases involving fire regimes, we only used Yellowstone 

data for the period in which fire suppression was 

not attempted in the Park, thereby making the data 

more relevant to National Forests where attempts 

to control fires have been the norm.

When data were not available, we substituted information about ecosystem processes and the history of the 

area to estimate the direction and magnitude of changes 

from the HRV that seemed probable.
6 USDA Forest Service RMRS-GTR-139. 2005

For each variable, we attempted to address (1) its HRV 

at the stand and landscape scale; (2) its spatial variability across the landscape as well as its temporal variability, when such information was available; and (3) our 

confidence level (low, medium, and high) in concluding whether or not the variable under existing conditions 

probably is beyond the HRV or trending in that direction. 

Our conclusions and confidence levels are summarized in 

section 8, which includes a summary table that gives the 

page numbers where each variable is discussed (table 7).

3. The Medicine Bow National Forest

3.1. Physical setting

The Medicine Bow National Forest covers approximately 405,000 hectares (1,000,000 acres) and is 

composed of four geographically separate units on the 

north end of the Southern Rocky Mountain physiographic province (fig. 3). The two eastern units are in 

the Laramie Mountains—the small 22,267 ha (55,000 

acre) Sherman Mountains unit between Laramie and 

Cheyenne (part of the Laramie District; sometimes 

called the Pole Mountain unit), and the moderately sized 

72,874 ha (180,000 acre) Laramie Peak unit toward 

the north end of the Laramie Mountains (the Douglas 

District). To the west of Laramie and east of Saratoga is 

the Snowy Range unit of the Medicine Bow Mountains 

(part of the Brush Creek/Hayden District and most of 

the Laramie District), which is the largest unit and covers about 210,526 ha (520,000 acres). To the west of 

Encampment and east of Baggs is the Sierra Madre unit 

(part of the Brush Creek/Hayden District), a northern 

extension of the Park Range in Colorado that covers approximately 137,651 ha (340,000 acres).

MT

SD

NE

CO

UT

ID

MBNF

Wyoming

SIERRA

MADRE

Laramie

Rawlins

WY

CO

SNOWY

RANGE

LARAMIE

MTNS

SHERMAN

MTNS

Douglas

Cheyenne

Figure 3—Location of the four units of the Medicine Bow National Forest in southeastern Wyoming. The Sherman Mountains unit 

is also known as the Pole Mountain unit, and the Laramie Mountains unit is sometimes referred to as the Laramie Peak unit.
USDA Forest Service RMRS-GTR-139. 2005 7

All of the MBNF is mountainous, with elevations 

ranging from 1915 m (6300 ft) near Laramie Peak to 

3652 m (12,013 ft) on the top of Medicine Bow Peak 

in the Snowy Range (fig. 4). The Snowy Range is bordered on the east by the Laramie Basin and on the west 

by the North Platte River Valley. Both the Snowy Range 

and Sierra Madre have cores of Precambrian granite and 

gneiss, but these rock types are overlain in many areas 

by metamorphosed sedimentary rocks, quartzite, and 

schist (Houston 1968; Houston and Ebbett 1977; Price 

1973; Oviatt 1977). A mountain glacier formed in the 

vicinity of Medicine Bow Peak during the Pleistocene, 

which resulted in the formation of cirques, till fields, 

moraines, and numerous lakes. Glaciation in the Sierra 

Madre was much less extensive, though Battle Lake occurs at the base of a cirque and Bridger Peak is thought 

to have been a monadnock above the ice. No glaciers 

formed over the Laramie Mountains.

The Laramie and Sherman Mountains units comprise the northernmost extension of the Colorado Front 

WY

CO

−−−−−LEGEND−−−−−

Major Road

Interstate

1,000 Foot Elevation Bands:

5,001 − 6,000

6,001 − 7,000

7,001 − 8,000

8,001 − 9,000

9,001 − 10,000

10,001 − 11,000

11,001 − 12,000

Approximate

5 0 5 10 15 20

Miles

Figure 4—Map of elevation zones for the four units that comprise the Medicine Bow National Forest. Source: MBNF.
8 USDA Forest Service RMRS-GTR-139. 2005

Range. The topographic relief is less than the Snowy 

Range and Sierra Madre units, but they all have similar geologic origins. Metamorphosed sedimentary and 

volcanic rocks are common. Mafic intrusive rocks are 

found in the area, and the Precambrian Sherman granite 

is abundant, such as in the Vedauwoo recreational area. 

Sandstones and limestones of the Casper and Fountain 

formations are found on the west slopes in the Sherman 

Mountains. To the east are Oligocene sedimentary substrates, remnants of which form the “gangplank” that 

slopes toward Cheyenne. An outcrop of limestone occurs along parts of Horseshoe Creek.

As with other portions of the Rocky Mountains, the 

climate of the MBNF varies greatly with elevation and 

topographic position. Generally, the mean annual precipitation ranges from 38 cm (15 inches) at lower elevations to 100 cm (39 inches) or more in the subalpine zone 

(Martner 1986). Over much of the forested area above 

2600 m (8580 ft), about two-thirds of the precipitation 

falls as snow. The Laramie Mountains unit is lowest in 

elevation and the warmest, with a mean annual temperature of 7° C (45° F) and a maximum recorded range of 

-40° C (-40° F) to 43° C (110° F) (Martner 1986). In the 

subalpine and alpine zones, the mean annual temperature is less than 2° C (35° F). Frost can occur at high elevations during any month of the year, with a low of -46° 

C (-50° F) in the winter and a maximum of about 32° 

C (90° F) on very warm days at mid-elevations in the 

summer (Martner 1986; Marston and Clarendon 1988). 

South slopes are considerably warmer than north slopes 

or creek bottoms. The numerous watersheds function 

also as cold-air sheds, creating occasional temperature 

inversions and contributing to great spatial variability in 

the elevational distribution of different vegetation types. 

The Laramie and Sherman Mountains are lower and drier than the Sierra Madre and Snowy Range units, and the 

southern half of the Snowy Range unit is drier than the 

northern half.

Historically, the Rocky Mountain region has experienced great climatic and geological changes over time 

scales of millions of years (table 1). Consequently, the 

vegetation has changed also. By the beginning of the 

Quaternary period, approximately 2 million years ago, 

many of the species present today were in place (Tidwell 

1975). From 2 million to 10,000 years ago, a period 

known as the Pleistocene epoch, several glacial advances 

and retreats affected both the topography and vegetation 

in the northern part of the Snowy Range and portions of 

the Sierra Madre (Marston and Clarendon 1988). Since 

the end of the Pleistocene around 10,000 years ago, shifts 

in climate have continued, often triggering changes in 

the spatial and elevational distribution of plant species 

(table 1; Whitlock 1993). During the last 500 years, the 

climate of the region has been characterized by extended, severe droughts as well as by extended wet periods 

(Gray and others 2004). Over shorter time scales (2-5 

years), the El Nino Southern Oscillation (ENSO) has influenced climate and subsequently fire occurrence in the 

Rocky Mountains. In Colorado, El Nino years appear to 

have been wetter than average and La Nina years have 

been drier (Veblen and others 2000). A similar pattern 

probably extends into the MBNF of southern Wyoming, 

though it weakens from south to north.

3.2. Ecological setting

The dramatic changes in climate over time have important implications for HRV analyses. Foremost, we 

must recognize that, by considering variation observed 

during our reference period (1600-1850), we are considering only a short period in the ecological history of 

MBNF ecosystems. Soils, for example, surely have developed over thousands of years, not centuries. Also, the 

mean return interval for various kinds of disturbances 

during our reference period, including fire, undoubtedly 

has changed as the climate changed.

Topographic and climatic diversity, along with periodic natural disturbances, caused a diversity of vegetation 

types to occur across the MBNF during the reference period. The zonation of vegetation on the mountain ranges 

is similar to other parts of the Rocky Mountains, but there 

is at least one important difference: The low-elevation 

woodlands of ponderosa pine and Douglas-fir common 

in Colorado and other parts of the Rocky Mountains are 

limited to small areas in the MBNF, probably because 

of the comparatively high elevations of the adjacent intermountain basins. To illustrate, foothill grasslands and 

shrublands typically grade directly into montane forests 

dominated by lodgepole pine around the Snowy Range 

and Sierra Madre units (Knight 1994a). Ponderosa pine 

is common only at lower elevations in the Sherman and 

Laramie Mountains. Another notable feature is the extensive aspen forest on the southwest slopes of the Sierra 

Madre Mountains, from the foothills up to about 2590 

m (8500 ft) elevation. Extensive aspen forests are common to the south in Colorado (Peet 1988), but they occur in much smaller groves over most of the MBNF and 

Wyoming.

Various classification systems have been applied to the 

upland vegetation (table 2). At the broadest scale, almost 

all of the MBNF is in the Northern Parks and Ranges 

Section (M331I) of the Southern Rocky Mountain 

Steppe-Open Woodland-Coniferous Forest Province 

(Bailey and others 1994, McNab and Avers 1994). The 
USDA Forest Service RMRS-GTR-139. 2005 9

Table 1—Chronology of major geologic and climatic changes in Wyoming. Mybp = million years before present; ybp = years before 

present. From literature reviewed by Knight (1994a) and von Ahlefeldt and Speas (1996).

Time Climate/geology Vegetation

350 mybp Tropical climate; submerged under saltwater Tropical

300 mybp Swamps, lagoons, tidal flats Clubmosses, horsetails, and ferns in swamps; pine, 

 spruce, and fir on upland

100 mybp Ancestral Rocky Mountains leveled, still 

 near equator

245-66 mybp Overthrust belt in western Wyoming formed Magnolia, palm, fig, breadfruit, sassafras, cinnamon, 

 (~66 mybp) sweetgum, and willow in wetlands; conifers, 

 ginkgo, and cycads on uplands

66-50 mybp Mountains of MBNF uplifted during Laramide Flowering plants more common in upland forests;

 orogeny; continued drifting apart of continents; species include alder, beech, birch, black locust,

 climate still tropical, snow rare, much wetter chestnut, cottonwood, cyprus, dogwood, elm, fir,

 than today ginkgo, hickory, maple, oak, pine, redwood, 

50-36 mybp Extensive erosion, volcanism, and basin filling spruce, sweetgum, sycamore, walnut, and willow;

10 mybp Regional uplifting creating present mountain palm and breadfruit also present

 landscape

7-5 mybp Increased aridity, creation of rainshadows Demise of some forests, giving rise to shrublands 

 and grasslands; drought tolerant species persisted 

 or immigrated; numbers of xerophytic plants 

 increased in grasslands; fires became more 

 frequent, contributing to the demise of trees and 

 the spread of grassland; woodlands restricted to 

 ravines, valley bottoms, and ridges

660,000 ybp Volcanism (creating Yellowstone plateau); climate 

 cooling, heavy snows, further cooling and 

 formation of glaciers

2 mybp - 10,000 ybp Six glacial advances and retreats; flooding, Coniferous forests, shrublands, and grasslands (many

 creation of outwash plains, deposition of loess, broad-leaved trees now regionally extinct)

 and soil development

127,000 ybp Interglacial period Douglas-fir and limber pine common in Yellowstone 

15,000 ybp Glacial advance; temperatures 10-13 ̊C colder Trees in Yellowstone confined to a narrow 

 than today; permafrost elevational band; lowlands tundra-like

11,500 ybp Mean temperatures 5-6 ̊C cooler than today; Upper treeline about 600 m lower than today in 

 retreat of glacial ice Yellowstone; gradually colonized by Engelmann 

 spruce, and later by subalpine fir and whitebark 

 pine in some areas (11,000-9500 ybp)

9500 ybp Continued warming Establishment of lodgepole pine at higher elevations

 and Douglas-fir in foothills in Yellowstone

9000-7000 ybp Upper treeline at its lowest in Colorado Front Range 

7000-4000 ybp Altithermal period, comparatively warm and Expansion of sagebrush, greasewood, juniper, and 

 dry conditions grasses; grassland found up to 2400 m on some 

 north slopes in MBNF; spruce and fir retreated to 

 higher elevations; elevational range of spruce and 

 fir reduced

4000 ybp - present Neoglacial period; gradual cooling Expansion of forests to previous elevational ranges

1350 – 1850 AD Characterized by extended and extreme wet 

 periods and droughts (Gray and others 2004)

1850 AD - present Slightly warmer than in previous centuries and Increased tree recruitment near upper treeline 

 with shorter, less extreme wet periods and (Hessl and Baker 1997), including subalpine 

 droughts (Gray and others 2004)) meadows (Jakubos and Romme 1993)
10 USDA Forest Service RMRS-GTR-139. 2005

one exception is the western portion of the Sierra Madre, 

which is part of the Central Basin and Hills Section 

(342F) of the Intermountain Semi-desert Province. At 

an intermediate scale, the forest vegetation can be classified by the dominant overstory species, i.e., cover type, 

which is the approach adopted for the MBNF database 

(fig. 5). The primary cover types are lodgepole pine, 

Engelmann spruce-subalpine fir, ponderosa pine, aspen, 

limber pine, and Douglas-fir forests, plus upland shrublands dominated by sagebrush (Artemisia spp.), serviceberry (Amelanchier alnifolia), and mountain mahogany 

(Cercocarpus montanus), and grasslands or meadows 

dominated by various species of grasses and forbs.

Most of the forested land occurs in the Snowy Range 

and Sierra Madre units, where about 78% of the land 

is forested and about 65% of the forests are dominated 

by lodgepole pine, Engelmann spruce, and subalpine fir 

Table 2—Upland vegetation types on the MBNF, classified at four levels of detail.

This report MBNF Alexander and others 1986 von Ahlefeldt and Speas 1996

High elevation forests Engelmann spruce-subalpine fir 3 subalpine fir habitat types 10 subalpine fir community types

6 subalpine fir-Engelmann spruce 

Lodgepole pine community types

Limber pine 5 lodgepole pine habitat types 3 Engelmann spruce community types

2 limber pine habitat types 13 lodgepole pine community types

5 limber pine community types

Low elevation forests Ponderosa pine 3 ponderosa pine habitat types 14 ponderosa pine community types

 and woodlands Limber pine 9 Douglas-fir community types

Douglas-fir and mixed conifer Douglas-fir series 4 Rocky Mountain juniper community

Rocky Mountain juniper Rocky Mountain juniper series types

Pinyon pine – juniper 1/

Aspen forest Aspen – birch 1/ 3 aspen habitat types 11 aspen community types

7 aspen-subalpine fir community types

1 aspen-lodgepole pine community 

 type

2 aspen-Douglas-fir community types

Non-forest vegetation Grasslands: wheatgrass, 1 tufted hairgrass type

 (grasslands, shrub- mountain muhly, needle-and- 3 danthonia types

 lands and forblands) thread, tufted hairgrass, 1 idaho fescue type

 bluegrass 1 mountain muhly type

2 needle-and-thread types

2 western wheatgrass types

1 bluebunch wheatgrass type

4 sedgeland types

2 kobresia types1/

Shrublands: serviceberry, Gambel oak series 4 serviceberry types

 sagebrush, mountain 10 sagebrush types

 mahogany, bitterbrush, alpine 5 mountain mahogany types

 dryad, shrubby cinquefoil, 1 ocean spray type

 greasewood, snowberry, 4 antelope bitterbrush types

 thinleaf alder, Gambel oak, bur 1 skunkbush sumac type

 oak, krummholz 1 western snowberry type

3 Gambel oak community types

Forblands 2 mountain dryad types

3 alpine avens types

2 sibbaldia types

1/

 Pinyon pine, kobresia, and birch trees (Betula spp.) are not known to occur on the MBNF, though birch shrubs are common in some areas.

(table 3 and figs. 5a and 5b). On the Sierra Madre 

unit, aspen occupies about 17% of the land area, much 

more than in the Snowy Range, Laramie Peak, and 

Pole Mountain units (3.8, 1.4, and 7.4%, respectively). 

Douglas-fir occupies no more than 2% of the land area 

in any of the four units (table 3 and fig. 5); ponderosa 

pine is most abundant on the Laramie Mountains unit 

(47% of the land area; fig. 5d). Limber pine is most common in the woodlands around Laramie Peak and Pole 

Mountain but occupies no more than about 5% of the 

MBNF. On the Sierra Madre and Snowy Range units, 

respectively, approximately 15% and 30% of the forests are more than 150 years old (fig. 6), suggesting that 

about 85% and 70%, respectively, of the forests in these 

units have been subjected to burning or timber harvesting since 1850. For the forests of the MBNF as a whole, 

approximately 55% of the forested land was classified as 
USDA Forest Service RMRS-GTR-139. 2005 11

Figure 5a—Land cover map for the Snowy Range unit of the MBNF. Source: MBNF.

suitable for timber harvesting in the 1985 management 

plan (the plan adopted in 2004 indicates that 86% of the 

forested land is “tentatively suitable” for timber production and 37% of the forested land is “suitable and appropriate” for timber production).

Shrublands and grasslands are common throughout 

the MBNF (table 3 and fig. 5). Big sagebrush (Artemisia 

tridentata) dominates most of the shrublands, but mountain-mahogany and Rocky Mountain juniper (Juniperus 

scopulorum) dominate small areas. Willows and thinleaf 

alder (Alnus incana) occur along the numerous creeks. 

The northern extension of Gambel oak (Quercus gambelii) occurs in the southwestern foothills of the Sierra 

Madre unit, where it is common in a small area (about 620 
12 USDA Forest Service RMRS-GTR-139. 2005

ha [1530 acres]). Grasslands are the most common vegetation type on the Sherman Mountains unit (table 3 and 

fig. 5c), where they occupy 47% of the landscape. In the 

other three units, grasslands and meadows cover no more 

than 3 to 8% of the land area. The most extensive alpine 

zone is in the Snowy Range unit (3483 ha [8601 acres]), 

but it comprises only 1.6% of the land area even there. 

Surface water (lakes, reservoirs, and streams) occupies 

less than 0.5% of the landscape in all units (table 3).

More specific forest classifications for the MBNF 

are based on both overstory and understory species. 

Alexander and others (1986) described nine principal 

forest series, based on overstory dominants, and 16 habitat types, based on overstory dominants and understory indicator species (after Daubenmire and Daubenmire 

1968; see table 2). Proponents of the habitat type approach contend that, while the spatial distribution of 

each tree species is controlled by factors such as elevation and topographic position, more specific environmental conditions (i.e., habitat types) are better reflected 

Figure 5b—Land cover map for the Sierra Madre unit of the MBNF. Source: MBNF.

by understory species. Considering both forest and nonforest vegetation, approximately 307 habitat types and 

community types have been identified on the MBNF 

(von Ahlefeldt and Speas 1996). Unfortunately, they 

have not been mapped and little information is available 

for each one.

After considering the appropriate scale to address 

HRV, we organized our discussion around four landscape types: subalpine or upper montane, high-elevation forests; montane forests at lower elevations; aspen 

forest; and non-forest vegetation (namely, upland shrublands and grasslands/meadows). There are important 

ecological distinctions between different habitat types 

and community types within each of the four landscape 

types that we examined, but the differences blur over 

the long time periods of HRV analyses. Furthermore, 

because of successional relationships among the most 

common overstory species (lodgepole pine, subalpine 

fir, and Engelmann spruce), it is difficult to discuss HRV 

in isolation for each forest type.
USDA Forest Service RMRS-GTR-139. 2005 13

Figure 5c—Land cover 

map for the Sherman 

Mountains unit of 

the MBNF. Source: 

MBNF.

3.3. Cultural setting

For centuries, humans have lived and traveled throughout the area that is today the MBNF (Grasso and others 

1981; Fawcett and Francis 1981; Thybony and others 

1986; Boglioli and others 2002). Consequently, managers 

have inherited the legacies of past land uses and are not 

starting with a “blank slate” (Romme and others 2000). 

This is an important consideration for several reasons. 

First, our ability to interpret pre-European-American 

conditions on the MBNF will be limited by events that 

have happened during the past 150 years. Also, it is difficult to distinguish events that were “natural” from those 

caused by Native Americans. Indeed, it may be necessary 

and even desirable to include some human influences as 

a factor affecting the HRV. However, many of the unprecedented human influences of European-Americans 

present unique challenges for managers now attempting 

to bring conditions within the HRV.

For centuries, Native American tribes such as the 

Northern Arapahoe, Northern Cheyenne, Ogalala Sioux, 

Ute, and Eastern Shoshone either inhabited lowlands 

surrounding the Medicine Bow Mountains or used canyons and mountain passes as travel routes (Grasso and 

others 1981; Thybony and others 1986; Boglioli and 

others 2002). Their impacts on the landscape are largely 

unknown but probably were minor. Only small amounts 

of wood, usually from small trees, would have been 
14 USDA Forest Service RMRS-GTR-139. 2005

Figure 5d—Land cover map for the 

Laramie Mountains unit of the 

MBNF. Source: MBNF.

Table 3—Acres and percent of total (in parentheses) of different land cover types on the four units of the Medicine Bow National 

Forest. To convert acres to hectares, multiply by 0.405. Source: MBNF.

Sierra Madre Snowy Range Laramie Mtns Sherman Mtns

Forested land

Aspen 59,800 (17%) 20,795 (3.8%) 2538 (1.4%) 759 (7.4%)

Cottonwood 310 (0.09) 10 (<0.01) 113 (0.06) 0

Douglas-fir 2374 (0.7 6105 (1.1) 1604 (0.9) 165 (1.6)

Limber pine 218 (0.06) 1583 (0.3) 9874 (5.4) 179 (1.8)

Juniper 178 (0.05) 70 (0.04)

Ponderosa pine 28 (0.01) 8616 (1.6) 86,167 (47) 1032 (10)

Lodgepole pine 135,181 (39) 280,141 (50) 58,318 (32) 1682 (16.5)

Spruce/fir 73,387 (21) 117,927 (21) 2478 (1.4) 18 (0.2)

TOTAL 271,476 (78) 435,178 (79) 161,162 (89) 3835 (37.6)

Shrublands

Gambel oak 1530 (0.4) 0 0 0

Sagebrush 44,595 (13) 40,989 (7.4) 8280 (4.6) 1372 (13.4)

Mountain mahogany 4176 (1.2) 1979 (0.4) 183 (0.1)

Riparian 1986 (0.6) 13,584 (2.5) 68 (0.03 213 (2.0)

Other shrublands 11,397 (3.3) 177 (<0.01) 3947 (2.2)

TOTAL 63,684 (18.4) 56,729 (10.3) 12,478 (6.9) 1585 (15.5)

Grasslands/meadows 9005 (2.6) 42,736 (7.7) 5676 (3.1) 4758 (47)

Alpine/krummholz/barren 2108 (0.6) 8601 (1.6) 2635 (1.4) 27 (0.3)

Lakes, reservoirs, streams 643 (0.18) 1401 (0.3) 5 (<0.01) 2 (0.02)

TOTAL AREA 346,916 acres 553,008 acres 181,951 acres 10,205 acres
USDA Forest Service RMRS-GTR-139. 2005 15

harvested. Fires would have been started in the vicinity 

of travel routes and campsites in the low-elevation forests 

and woodlands (Barrett and Arno 1982; Gruell 1983a), 

and some of those fires might have burned into adjacent high-elevation vegetation as well. However, Baker 

(2002a) concluded that, for Rocky Mountain landscapes 

in general, Native Americans had little effect on fire regimes in lodgepole pine and spruce-fir forests.

Beginning in the early 1800s, trappers and explorers 

from the east began to move into the Rocky Mountain 

West. Human land uses on the MBNF during this time 

probably remained about the same, primarily because 

travelers in southern Wyoming usually followed courses around rather than through the mountains (Townsend 

1978; Grasso and others 1981). Perhaps the most significant additional influence was the cascade of changes resulting from the reduction in beaver populations caused 

by trappers, plus a large fire in the southwestern Sierra 

Madre Mountains (Grasso and others 1981; Thybony and 

others 1986). In general, the effects of Native Americans 

and the early explorers must have been less conspicuous than those of European-American settlers after the 

1860s, when the harvesting of big trees became common.

By the mid-1800s, the Overland and Oregon Trails 

facilitated the immigration of European-Americans to 

southern Wyoming. In the decades that followed, millions of large trees were cut for telegraph poles, railroad 

ties, mine timbers, fuel wood, and construction (figs. 6 

and 7). Lodgepole pine trees with a diameter at breast 

height (dbh) of 28-36 cm (11-14 inches) were harvested for railroad ties, a practice known as “tie hacking” 

that continued until the 1950s (Grasso and others 1981; 

Thybony and others 1986; Wroten 1956). Human-caused 

fires burned through the slash of some of the harvested stands, often killing the trees that were not harvested. The slash often was used as a source of fuel wood 

for cooking and heating. Riparian zones fringing major creeks, like Douglas Creek, were sometimes clearcut to facilitate the movement of logs and railroad ties 

0

10000

20000

30000

40000

50000

60000

< 1850 1850-1909 1910-1949 >1949

Stand origin

Hectares

0

5

10

15

20

25

30

35

40

45

Percent of total

0

10000

20000

30000

40000

50000

< 1850 1850-1909 1910-1949 >1949

Stand origin dates

Hectares

0

10

20

30

40

50

60

70

Percent of total

Figure 6—Age distribution of lodgepole pine and spruce-fir forests on the Snowy Range and Sierra Madre 

units of the MBNF. Only areas for which age data are available are included (37% of the lodgepole pine 

and spruce-fir forests on the Snowy Range unit; 32% on the Sierra Madre unit). Source: MBNF; adapted 

from von Ahlefeldt and Speas (1996). 
16 USDA Forest Service RMRS-GTR-139. 2005

to rivers, where they were floated downstream (Young 

and others 1989, 1994). In the Sherman Mountains there 

were insufficient numbers of trees to provide an adequate 

supply of ties, so they were imported from outside the 

region (Wroten 1956). Many trees in the Pole Mountain 

area were cut, though much of the land was not forested 

during the reference period (1600-1850). To the north, 

near Laramie Peak, trees were cut for telegraph poles 

and other purposes (Thybony and others 1986).

Early European-Americans also took advantage of 

other resources. Mining for gold and other ores became 

common in portions of the Snowy Range, Sierra Madre 

and Laramie Mountains (Duncan 1990). Old photos 

show large areas of clearcut timber prior to 1880 in the 

vicinity of mines, especially in the Sierra Madre unit 

(Thybony and others 1986). Also, large herds of sheep 

and cattle were common by 1880 on mountain grasslands and shrublands (Grasso and others 1981). The 

livestock sometimes benefited from the increased forage 

production that occurs following forest fire and timber 

harvesting. Simultaneously, the native herbivores (elk, 

deer, bighorn sheep, pronghorn antelope, and bison) 

were hunted for eastern markets and sport (Dary 1986), 

driving some species to near extinction. Poisoning and 

shooting decimated populations of the native predators, especially wolves, mountain lion, and grizzly bear. 

Many parts of the MBNF are thought to have burned 

during the late 1800s, with ignitions commonly associated with wood stoves, steam-powered saw mills, and the 

Laramie Hahn’s Peak and Pacific Railroad that crossed 

the Snowy Range. Partially due to harvesting and burning in the late 1800s and early 1900s, approximately 70-

80% of the forests on the MBNF are now 90 to 130 years 

old (see fig. 6).

Concerns about the adverse effects of EuropeanAmerican activities brought about the period of National 

Forest administration. In 1902, the U.S. Government began the establishment of several forest reserves in southeastern Wyoming that would eventually become the 

MBNF (Medicine Bow National Forest 1905). With the 

establishment of the reserves, land uses such as timber harvesting and grazing were regulated and more 

Figure 7—Watersheds with streams known to have been used for railroad tie drives. Source: MBNF
USDA Forest Service RMRS-GTR-139. 2005 17

0

500

1000

1500

2000

2500

1950 55 1960 65 1970 75 1980 85 1990

Year

Hectares

clearcut partial cut

Figure 9—The land 

area in the Snowy 

Range and Sierra 

Madre units where 

clearcutting and 

partial cutting have 

occurred during the 

period 1950-1994. 

Source: MBNF.

Figure 8a—Cumulative land area affected by clearcut harvesting 

in 1950, 1970, and 1990. From Baker 1994.

Figure 8b—Cumulative land area affected directly by shelterwood harvesting by 1994, partial cuts by 1994, and all 

clearcuts, shelterwood cuts, and partial cuts by 1994. The 

depth-of-edge effect is not shown. From Baker 1994.
18 USDA Forest Service RMRS-GTR-139. 2005

careful attention was given to finding the best ways to 

use the land (Pinchot 1907; Armstrong 1935). Also, fire 

suppression became a primary concern. Fire was personified as the “enemy” to forest managers (USDI 1897, 

p. 80). Steam-powered sawmills prior to 1903, and gasoline-powered mills later, were a common source of 

ignitions.

In the mid-1930s, harvests of 65-85% of the trees in 

a stand became common. Partially because of problems 

with blowdown in thinned stands, clearcutting was initiated in the late 1940s (Lexen 1949; Alexander 1975; 

Alexander and Edminster 1980). Also, new levels of prosperity in the region after World War II created larger markets for timber than ever before and the creation of new 

stands of rapidly growing, economically valuable trees 

became a common goal for National Forests (Langston 

1995)—providing another incentive for clearcutting. 

The cutting of larger trees was encouraged (figs. 8 and 

9). Gasoline-powered sawmills, some that were portable, 

had been in use since the 1920s and were common. Also, 

road building increased after 1950, making much more of 

the forest accessible for timber harvest (fig. 10). In addition to strip cutting and dispersed clearcuts, some stands 

were partially cut by individual selection, group selection, and shelterwood silvicultural systems (Alexander 

1975; Alexander and Edminster 1980). Unlike the previous harvesting that was concentrated in low-elevation 

stands, many subalpine forests were cut for the first time 

in the 1950s and 1960s. Natural tree regeneration in harvested stands has been quite good on the MBNF, except 

on wind-swept sites at high elevations. In such environments there may be less seed, and seedling mortality during the winter can be high (probably because of winter 

desiccation that occurs when too much of the snow is 

blown into the forests downwind). Even there, new trees 

eventually become established, though their density is 

low and their growth is slow. Over most of the MBNF, 

seeding or planting of trees is not required in harvested 

areas to assure an adequate number of seedlings within the five-year period mandated by the National Forest 

Management Act of 1976 (Jerod Delay, personal communication).

Timber sales in the MBNF have declined in recent decades, from a high of ~58 million board feet per year in 

the late 1950s to ~1.3 million board feet in 1999 (MBNF 

data). Clearcut or patch-cut harvesting is still common, 

though with greater sensitivity to wildlife and aesthetic 

values. The intensity of livestock grazing has been reduced as well, from about 400,000 sheep and 200,000 

cattle in the early 1900s (Boglioli and others 2002) to 

about 20,000 sheep and 16,000 cattle at the present time 

(MBNF data).

An important new effect has been the influence of 

people from throughout the nation who are engaged 

in a wide range of recreational activities. Moreover, 

new homes have been built on acreages adjacent to the 

MBNF. Due to accessibility, the growing popularity of 

outdoor recreation, and proximity to large and rapidly 

growing human populations in Colorado, such influences have had widespread impacts. Managing outdoor 

recreation and exurban developments could become as 

controversial as timber management. As noted, the increasing demands placed on public lands and new mandates from Congress during the last 30 years have led 

to the philosophy of ecosystem management, for which 

HRV assessments are desired.

4. The HRV of High-Elevation Forest

4.1. Variation in species composition and 

succession

Forests dominated by lodgepole pine, Engelmann 

spruce, and subalpine fir are common at middle to high 

elevations on the Snowy Range and Sierra Madre units 

(figs. 11 and 12) and have been the subject of numerous studies. Lodgepole pine is the dominant species at 

elevations of approximately 2400 to 3000 m (7920 to 

9900 ft), where it occupies 66% and 51% of the forested 

area on the Snowy Range and Sierra Madre units (von 

Ahlefeldt and Speas 1996). Subalpine fir and Engelmann 

spruce are dominant in the subalpine zone (3000 to 3300 

m; 9900 to 10,890 ft), covering 27% of forested land on 

both units (von Ahlefeldt and Speas 1996). Notably, subalpine fir and Engelmann spruce are found throughout 

the elevational range of lodgepole pine and all three species often occur together (Veblen 1986). In the Laramie 

Mountains, lodgepole pine is less common (31% and 

37% of the forested land on the Sherman Mountains and 

Laramie Mountains units, respectively), and Engelmann 

spruce and subalpine fir occur only in small stands.

The distribution of tree species is affected by moisture availability, which usually is controlled by elevation and topographic position. Spruce-fir forest is 

typical on mesic sites, such as in ravine bottoms or on 

north slopes, while lodgepole pine is common on comparatively dry sites (such as on south slopes) as well 

as in more mesic habitats (Romme and Knight 1981). 

Small patches of spruce-fir forests can be found in ravines at lower elevations where cold air drainage occurs 

(Romme and Knight 1981). High-elevation landscapes 

often have abrupt changes in topography, causing many 

forests to occur in small patches (Knight 1994b; Knight 
USDA Forest Service RMRS-GTR-139. 2005 19

Figure 10—Roads within each of the four units of the Medicine Bow National Forest. All roads that exist as a feature of the presentday landscape are shown, including closed, abandoned, obliterated, and non-USFS roads. Road density in the MBNF is about 

1.3 miles per square mile of National Forest. Source: MBNF
20 USDA Forest Service RMRS-GTR-139. 2005

and Reiners 2000). Comparatively flat topography on 

some parts of the MBNF, however, enables the development of rather extensive forest stands, especially in 

the Snowy Range (Marston and Clarendon 1988; Knight 

and Reiners 2000). As discussed in later sections, this 

can have important implications for natural disturbance 

regimes and timber harvest activity.

Lodgepole pine traditionally has been considered a 

pioneer species because of its relative intolerance for the 

forest understory environment and because it grows well 

following disturbances (Baker 1949; Lotan and Perry 

1983; Lotan and others 1985; Parker and Parker 1994). 

Its seedlings are able to withstand cold temperatures 

(Cochran and Berntsen 1973). This characteristic, combined with a history punctuated by disturbances over the 

last 150 years, help explain the prevalence of lodgepole 

pine on the MBNF today. Lodgepole pine is a disturbance-adapted species, often bearing cones that remain 

closed for many years, thereby storing large numbers of 

seed on a tree. Known as serotinous or “closed” cones, 

they open primarily when exposed to higher than normal 

temperatures, such as during a fire or when the cones are 

near the soil surface (which is warmer than the tree canopy). Significantly, not all lodgepole pine produce serotinous cones and the proportion of trees with “closed” 

and “open” cones is highly variable (Lotan and Perry 

1983; Muir and Lotan 1985; Tinker and others 1994).

As mature lodgepole pine begin to die after a hundred 

years or more, they sometimes are replaced by subalpine fir and Engelmann spruce—a textbook example of 

succession that leads to the development of an unevenaged forest structure (Daubenmire 1943; Oosting and 

Reed 1952; Knowles and Grant 1983; Knight 1994a). 

However, another disturbance commonly occurs before 

Figure 11—Distribution of lodgepole pine forests on the Medicine Bow National Forest. Source: MBNF.
USDA Forest Service RMRS-GTR-139. 2005 21

Figure 12—Distribution of spruce–fir forests on the Medicine Bow National Forest. Source: MBNF.

spruce and fir become dominant. Of course, this successional pathway is just one of several possible trajectories 

following stand replacing disturbances (fig. 13 and table 

4). Relatively stable lodgepole pine forests have been reported on the MBNF (Alexander and others 1986) and 

throughout the Rocky Mountains (Moir 1969; Loope 

and Gruell 1973; Whipple and Dix 1979; Despain 1983, 

1990; Lotan and Perry 1983; Peet 1988). Such forests 

are typically found where conditions are not favorable 

for subalpine fir or Engelmann spruce seedling establishment (e.g., on sites that are relatively dry) or where 

there is no readily available seed for these and other species. Many lodgepole pine trees commonly die before 

reaching 250 years of age, with reproduction occurring in canopy gaps formed by dying trees. Aspen can 

be another important pioneer species, but like lodgepole 

pine, it too forms self-perpetuating stands in some habitats (Mueggler 1985). Elsewhere, spruce and fir may be 

pioneer species, as Doyle (1997) documented in Grand 

Teton National Park. Reviewing the full range of successional trajectories is beyond the scope of this report, but 

a consideration of succession provides a basis for evaluating potential deviations from the HRV of some stand 

and landscape variables.

Generally, the time required for a spruce-fir forest to 

develop after a fire may range from 150 to over 1000 

years (Oosting and Reed 1952; Billings 1969; Mueggler 

1976; Romme and Knight 1981). Villalba and others 

(1994) concluded that “old growth” subalpine forests in 

Colorado should have some trees greater than 300 years 

old. Mehl (1992) suggests 200 years. Two or three centuries may be required for the development of important components of old-growth forest, namely, relatively 

large trees, numerous snags, and relatively high amounts 
22 USDA Forest Service RMRS-GTR-139. 2005

of coarse woody debris (CWD) (Mehl 1992; Rebertus 

and others 1992). Franklin and Laven (1989) stressed 

that forest development is not easily predicted by simple 

models, as disturbances of different timing and severity 

create alternative pathways of succession, depending on 

environmental conditions at the time. Indeed, the time 

between subsequent disturbance events in some areas 

may be shorter than the time required for old spruce-fir 

forests to develop. On some watersheds, old spruce-fir 

forests can be restricted to places protected from fire, 

such as ravines or leeward slopes (Romme and Knight 

1981; Despain and Romme 1991). Notably, Baker and 

Kipfmueller (2000) could not find a relationship between old forests and topographic position in the 3241 

ha (8005 acre) Rock Creek watershed they studied on 

the north end of the Snowy Range unit.

It is important to recognize that secondary succession initiated today could have quite different trajectories than 200-300 years ago because of different climatic 

conditions. Similarly, the distribution of individual species may have shifted for the same reason. Investigators 

have found evidence in the pollen record, for example, 

that the relative abundance of lodgepole pine, Engelmann 

spruce, and subalpine fir at both the stand and landscape 

scales has shifted over long time periods due to climate 

change (Hansen 1940; Fall 1997). Whitlock (1993) documented altitudinal shifts of tree species distributions in 

northwest Wyoming during the last 12,000 years; and 

in recent centuries lodgepole pine, subalpine fir, and 

Engelmann spruce have grown into meadows (Patten 

1969; Schimpf and others 1980; Pearson and others 1987; Jakubos and Romme 1993; Hessl and Baker 

1997). Some of these observations have been attributed, 

at least in part, to warmer and possibly somewhat wetter 

conditions since ca 1850-1870.

4.2. Disturbances in high-elevation forests

Historically, fire has been the predominant stand-replacing disturbance, with occasional insect epidemics 

and windstorms. Disturbances caused by diseases were 

apparently less widespread. As discussed below, large 

numbers of trees can become infected with dwarf mistletoe after long periods without fire, but the effect of 

mistletoe is usually to slow growth rather than to kill 

the tree. Disturbances have not acted in isolation, but 

rather they have had interactive relationships that enhance the natural diversity of subalpine forests. For example, a windstorm that topples a large number of trees 

may increase the probability of an insect epidemic or 

fire in some areas (Schmid and Frye 1977; Lindemann 

and Baker 2001). Successional processes have been initiated over and over during the last 10,000 years, since 

the retreat of the last glaciers, creating a diverse landscape mosaic.

Today, the frequency, size, and intensity of historic 

disturbances, and the forests that develop after different 

kinds of disturbances, often have been altered by management activities. In the following sections we examine 

these changes in more detail and attempt to characterize 

the HRV for the major disturbances. Subsequently, we 

discuss how some stand and landscape variables appear 

Figure 13—Alternative successional pathways following light and severe fires in spruce-fir forests of the 

Central Rocky Mountains. Note that 

spruce and fir are sometimes pioneer species, along with lodgepole 

pine or aspen, and that severe fires 

can lead to the establishment of 

grasslands or meadows at high elevations. Adapted from Stahelin 

1943.
USDA Forest Service RMRS-GTR-139. 2005 23

Table 4—Examples of successional stage characterizations for lodgepole pine and spruce-fir forests in the region of the MBNF.

Years after standSuccessional stage replacing disturbance Characteristics

Romme and Despain 1989; LP0 0 to 30-50 Vegetative sprouting of herbaceous plants; 

 lodgepole pine seedling establishment; many 

Despain and Romme 1991 large, dead tree boles (snags and coarse 

 woody debris) that do not ignite easily, but 

 saplings can become sufficiently dense to 

 carry fire during drought years

LP1 30-50 to 150-200 Canopy closure; stands can be relatively dense

 with little understory vegetation; period 

 of rapid tree growth, development of 

 competition, and thinning as stands age; fuels

 are discontinuous from forest floor to canopy;

 does not burn readily except during periods 

 with severe drought and wind

LP2 150-200 to 200-400 Tree growth declines due to tree age and 

 competition; some tree mortality creates 

 canopy gaps; understory vegetation more 

 dense and more diverse; mixed understory of 

 lodgepole pine, subalpine fir, and Engelmann

 spruce common; woody fuels increase 

 but still somewhat resistant to fire; fires are 

 less intense in absence of wind and drought

LP3 after 200-400 Canopy mortality increases; understory trees 

 reach 2-3 m tall and penetrate canopy gaps; 

 fuels and organic materials accumulate and 

 burn easily and intensively during drought

Aplet and others 1988 Colonization phase 0 to 100-200 Initial colonization by subalpine fir and 

 Engelmann spruce simultaneously (and 

 potentially lodgepole pine at some sites, 

 although not in the study by Aplet and others 

 1988); period of initial tree growth and 

 canopy closure

Spruce exclusion 100-200 to 200-300 Dense canopy; Engelmann spruce seedling 

phase (lasting for ~100 years survival in understory is rare 

after initial colonization)

Spruce reinitiation 200-300 to 300-400 Dominant subalpine fir and Engelmann spruce

phase (lasting for ~100 years (and some old lodgepole pine) begin to 

after spruce exclusion) die, creating canopy gaps; seedlings of 

 Engelmann spruce and sometimes lodgepole 

 pine are able to survive in understory again

Second generation after 300-400 Uneven-aged canopy; cohort of spruce 

spruce-fir forest established during reinitiation phase reaches 

 canopy; high fuel accumulation; very 

 flammable during suitable summer weather 

 conditions

USFS (1985) Grass/forb/seedling lodgepole (spruce-fir) Colonization; trees <2.5 cm diameter at breast 

0 to 20 (0 to 25) height (DBH) for lodgepole pine, subalpine fir

 and Engelmann spruce

Sapling 21 to 40 (26 to 50) Most trees 2.5 – 12.5 cm DBH

Pole 41 to 80 (51 to 80) Most trees 12.5 – 23 cm DBH

Mature 81 to 140 (81 to 140) Most trees >23 cm DBH

Overmature 140+ (140+) Most trees are larger than 33 cm DBH
24 USDA Forest Service RMRS-GTR-139. 2005

to be either beyond their HRV at the present time or are 

trending in that direction.

4.2.1. Fire

Fire has shaped the vegetation mosaic for thousands 

of years by causing disturbances on a variety of scales. 

The evidence of historical fires is recorded by charcoal and pollen deposited in bog and lake sediments 

(Mehringer and others 1977; Millspaugh and Whitlock 

1995; Whitlock and Millspaugh 1996; Fall 1997), and by 

annual rings on trees that survive fire events (fire scars) 

or which become established within a few years after 

the fire. Fire intensity in high-elevation forests varies according to weather, fuel, and site conditions.

A common approach for estimating site-specific fire 

return intervals (i.e., for stands) involves age-class sampling and evaluation of stand structure and fuel accumulation (e.g., Billings 1969; Romme 1980a, 1982; 

Romme and Knight 1981; Romme and Despain 1989; 

Kipfmueller and Baker 1998a, 2000). Based on this approach, the fire-free interval for a particular stand of 

subalpine forest in the MBNF can be 25 to 700 years. 

Average fire return intervals have also been estimated by 

counting the rings on fire-scarred trees (Hawkes 1980; 

Kipfmueller 1997). Such estimates are often shorter than 

others because they consider the occurrence of more of 

the fire events, not just stand-replacing fires. The great 

variation in fire return interval estimates is most likely 

related to differences in site characteristics and species 

composition. The longest fire-free intervals probably 

are associated with mesic sites found at high elevations, 

in ravines, or on north slopes (Billings 1969; Hawkes 

1980; Romme and Knight 1981; Tinker and others 

1994)—though this pattern is not universal (Baker and 

Kipfmueller 2000).

Another way of characterizing fire frequency is 

the mean fire interval (MFI) for a large area, such as 

a landscape (e.g., Veblen and others 1994; Wadleigh 

and Jenkins 1996; Kipfmueller 1997; Kipfmueller and 

Baker 1998a). MFI is calculated simply as the arithmetic 

mean of intervals between all known fire events within 

a landscape, even though different portions of the area 

burn each time (Romme 1980b). Because this statistic is 

area dependent, larger study areas typically have more 

fire events, leading to lower MFI calculations (Arno and 

Peterson 1983; Kipfmueller 1997). Therefore, MFI estimates from different areas are difficult to compare.

Kipfmueller and Baker (2000) calculated MFI for the 

Rock Creek watershed (3241 ha [8005 acres]) on the 

north end of the Snowy Range unit for three distinct time 

periods: (1) prior to the arrival of European-Americans 

in the 1800s (i.e., the reference period); (2) after their arrival but before effective fire control; and (3) in the midto late-1900s when fire control was attempted in some 

areas. Their results suggest a shortening of the MFI with 

the arrival of European-Americans, which indicates an 

increase in ignitions (fig. 14). The short MFI for the late 

1800s may reflect a combination of European-American 

influences and a drier climate. Since the early-1900s, 

MFI has increased slightly, but it is still shorter and less 

variable than before the arrival of European-Americans, a 

trend noted in Colorado as well (Zimmerman and Laven 

1984). Therefore, while fire ignition rates were highest during the late 1800s, they probably have remained 

higher in the 1900s than they were during the reference 

period as a whole, in part due to the influence of humanFigure 14—Mean fire intervals (MFI) for a portion of 

the Snowy Range as calculated by Kipfmueller (1997). 

Values represent the time between fires anywhere within 

the 3,241 ha (8,005 acre) 

study area that Kipfmueller 

used. Lines represent the 

range between minimum 

and maximum intervals observed, while the mark on 

each line represents the 

estimated MFI. Estimates 

for the pre-1867 fires are 

less precise. Adapted from 

Kipfmueller 1997.
USDA Forest Service RMRS-GTR-139. 2005 25

caused ignitions. Interestingly, the number of total ignitions (including human-caused) have been much higher 

in the Snowy Range and Sierra Madre in recent years 

than in Yellowstone National Park, while the number of 

lightning-caused ignitions on the two MBNF units has 

been similar to total ignitions in YNP (fig. 15). During 

the same time, however, the amount of land area burned 

on the MBNF was relatively low compared to YNP (fig. 

16), suggesting that fire suppression on the MBNF has 

been effective.

Further insight into the frequency and size of fire 

events at high elevations, namely on the Snowy Range 

and Sierra Madre units, can be gained by considering 

both long-term and recent fire histories in Yellowstone 

National Park. As an area that has remained relatively undisturbed by human influences, and one that is in 

many ways composed of similar topography and vegetation as the MBNF, YNP provides a useful proxy for 

HRV analyses. Despain and Romme (1991) reconstructed the fire history for a large, 129,600 ha (320,112 acre) 

study area in YNP (approximately two-thirds the area 

of the Snowy Range unit). Although fire apparently has 

occurred in every decade, they found that large-scale, 

stand-replacing fires have occurred during four main peFigure 15—Number of fire starts per year per million hectares in lodgepole pine and spruce-fir 

forests on the MBNF, compared to comparable 

data for Yellowstone National Park. MBNF data 

are for 1945-1993 (von Ahlefeldt and Speas 

1996); YNP data are for 1972-1988 (Despain 

and Romme 1991). The dark line indicates the 

mean and bar height indicates range.

Figure 16—Hectares burned per year per million hectares in Yellowstone National Park and the Snowy Range and Sierra Madre 

units of the MBNF. The horizontal lines indicate the absolute range. Also shown are the mean and median for YNP and the mean 

for the MBNF. Data for YNP are from Despain and Romme (1991) and represent fire dynamics with no or little fire suppression, 

in the area they studied, for the period 1972-87. Data for the MBNF were compiled by von Ahlefeldt and Speas (1996) and are 

for the period 1945-93, when attempts at fire suppression were common and often successful. 
26 USDA Forest Service RMRS-GTR-139. 2005

riods since 1690 AD, separated by intervals of 20, 110, 

and 118 years. Such data do not represent intervals between fires in a specific stand, but rather between major 

fire events somewhere in their study area. During each 

major fire period, a large percentage (7-26%) of the entire area burned, initiating extensive, relatively even-aged 

patches of forest. Despain and Romme concluded that 

such even-aged cohorts become increasingly susceptible 

to large-scale burning as succession proceeds, and that 

very likely they will burn within 200-400 years, whenever weather conditions permit (table 5). Ignitions are likely to occur every summer, but most fires do not continue 

to burn, even without suppression by managers.

Recent fire observations in YNP provide information 

on the variability in size for individual fires in the absence of fire suppression. Since 1972, YNP managers 

have allowed lightning-caused fires to burn in many instances, providing a proxy for possible dynamics in the 

early and middle 1800s. Based on fires that burned from 

1972-1988, Despain and Romme (1991) observed that 

lightning fires typically burn only a few hectares or less 

under most conditions, and that weather conditions in 

only five years during a 15-year period (1972-1987) allowed fires to burn more than 100 ha (247 acres), namely, once every three years, on average. Looking at the 

area burned per year per million hectares from 1972-

1987 in YNP (fig. 16), three important trends emerge: 

(1) The area burned in a single year is highly variable, 

ranging from 0 to 10,423 ha per million ha; (2) the mean 

area burned per year per million hectares is toward the 

low end of the range at 1138 ha/year (2811 acres/year); 

and (3) the median, possibly a better indicator of a typical year, is only 12 ha/year/million ha.

Thus, the overall fire regime in YNP is characterized 

by numerous fires (lightning and human-caused) that are 

unable to spread over more than a few hectares, and relatively few fires that burn large areas (and have by far 

the greatest effect on the landscape mosaic, such as the 

fires of 1988)—a finding consistent with the results of 

Kipfmueller and Baker (2000) for a small watershed on 

the north end of the MBNF. The driving factors that create this regime appear to be weather conditions and the 

type and distribution of fuel (Despain and Romme 1991; 

Baker 2002b). Kipfmueller and Baker (2000) identified 

a moderately sized fire around 1680 AD that burned at 

least 2500 ha (6175 acres) in the northern Snowy Range 

(as well as another in 1743 that apparently burned >1500 

ha [3705 acres]). Much larger fires probably occurred, 

as they have in YNP. However, thus far no evidence for 

them has been found on the MBNF (William Baker, personal communication)—perhaps because a substantial 

portion of the evidence has been lost due to the extensive timber harvesting and fires that occurred after the 

arrival of European-Americans.

Fire records for the MBNF from 1945-1993 (summarized by von Ahlefeldt and Speas 1996) suggest that fires 

have played a much smaller role in the mid- to late 1900s 

than previously, as only six fires larger than 125 ha (309 

acres) burned in the Snowy Range and Sierra Madre during this 48-year time span. Significantly, these six fires 

Table 5—Examples of ranges for mean fire intervals in high-elevation forests, as estimated by various investigators. The values are 

the estimated average time between a series of successive fire events in an individual stand (point-scale).

Fire return 

interval estimate

Source Forest type / location (years) Type of fire event Basis for estimates

Billings (1969) Spruce-fir / MBNF up to 600-700 Stand-replacing Stand age structure

Arno (1980) Subalpine forest / N. Rockies 150-350 Stand-replacing Fire-scarred trees and 

 stand age structure

Romme (1980a, 1982) Lodgepole pine, fir, spruce / YNP 300+ Stand-replacing Stand age structure and

 fuel accumulation

Romme and Knight (1981) Spruce-fir / MBNF 300-500+ Stand-replacing Stand age structure and

 fuel accumulation

Romme and Despain (1989) Lodgepole pine, fir, spruce / YNP 200-400 Stand-replacing Stand age structure and

 fuel accumulation

Hawkes (1980) Lodgepole pine, fir, spruce, 90-304 1/ all Fire-scarred trees

 larch / Alberta

Kipfmueller (1997) Lodgepole pine, fir, spruce / 39-149 2/ all Fire-scarred trees

 MBNF

1/ This range is based on the means calculated for forests at different elevations and topographic positions, including: (1) lower subalpine/lower 

elevation (90-101 years); (2) upper subalpine/upper elevation (153-304 years); (3) north aspects (187 years); and (4) south, east and west aspects 

(93-104 years). 2/ This range is based on four different methods for calculating point-scale MFI: (1) the interval between stand origin date and first scar (82 years); 

(2) the interval between scars on the same tree (39 years); (3) the interval between the last scar and the present (149 years); and (4) the overall 

average of all intervals (129 years).
USDA Forest Service RMRS-GTR-139. 2005 27

accounted for almost 70% of the total land area burned 

(3831 ha of 5642 ha) (fig. 17). The largest fire recorded from 1945 to 1993 in these two areas was only 513 

ha (1267 acres), much smaller than large historical fires 

(e.g., Despain and Romme 1991; Kipfmueller and Baker 

2000), and the vast majority of fires that burned during 

this time (ca 95%) were less than 4 ha (9.9 acres). The 

total area burned during the 48-year period (1945-93) 

on the Snowy Range and Sierra Madre was only 5462 

ha (13,491 acres), approximately 1.6% of the National 

Forest land in these two areas. Most of the fires burned at 

middle elevations (2400-2850 m), where lodgepole pine 

is the dominant species.

The probable cause for this low incidence and extent of fire in the MBNF is effective fire suppression 

(Kipfmueller and Baker 2000). Notably, 1988 was 

a drought year in the MBNF as it was in YNP. Fortythree fires were ignited on the MBNF in 1988, 29 from 

lightning and 14 from human activity. All but one were 

extinguished before they had burned more than a few 

hectares; the exception burned only 400 ha (988 acres).

Unfortunately, there is insufficient information available to conclude that lodgepole pine forests at lower elevations have a shorter MFI than at higher elevations, 

though there is a tendency for younger stands to be located at lower elevations in both YNP and the MBNF 

(William Romme, personal communication). Fires in 

low-elevation forests, once started, could have spread 

into the adjacent forests during our reference period, 

thereby shortening the MFI in forests at higher elevations. However, Baker and Kipfmueller (2001) did not 

find evidence for a relationship between MFI and elevation in the Rock Creek drainage on the north end of the 

Snowy Range. Research over a larger area in the MBNF 

is needed.

Aside from fire suppression, weather probably has 

been an important factor accounting for the relatively low extent of burning in the MBNF in recent years. 

Recent research suggests that weather conditions rather than fuel accumulations (Bessie and Johnson 1995; 

Weir and others 1995; Clark 1990; Keane and others 

2002) cause large stand-replacing fires, though some investigators have emphasized the combination of weather and fuels. For example, Romme and Despain (1989) 

concluded, based on fuel availability, that extensive fires 

could have occurred in YNP after about 1930, but they 

did not occur until 1988 because of weather conditions. 

Baker (2002b) provides a detailed review of various factors affecting the frequency and extent of fires in the 

Rocky Mountain region, concluding that climate, fuels, 

and elevation are linked causes and that ongoing human 

land uses can be confounded with climatic effects.

Dealing with infrequent but large fires presents a difficult challenge to forest managers. On one hand, disturbances of the magnitude of the 1988 Yellowstone fires 

could be characteristic of our reference period for the 

MBNF and could have had very important ecological 

effects (e.g., Christensen and others 1989). Preventing 

5 1 0 12

488

121

30

0

100

200

300

400

500

600

A B C D E F G

Size class

Number of fires

0

500

1000

1500

2000

2500

3000

3500

Area burned (ha)

number of fires area burned

Figure 17—Number of fires and area burned on the Snowy Range and Sierra Madre units of the MBNF, 

1945-1993, based on MBNF data summarized by von Ahlefeldt and Speas (1996). Size classes are: 

A < 0.1 ha; B 0.1-3.9 ha; C 4-39 ha; D 40-122 ha; E 123-409 ha; F 410-2049 ha; G >2049 ha.
28 USDA Forest Service RMRS-GTR-139. 2005

such events in the future, if that is possible, could have 

consequences for biological diversity and other ecosystem features (Keane and others 2002). On the other 

hand, such disturbances also are a frightening extreme 

in the HRV that the public has encouraged managers to 

suppress (Swanson and others 1993; Manley and others 

1995). Whether or not attempts to suppress intense fires 

in high-elevation forests have been or can be successful remains controversial. Many fires, whether started by 

lightning or humans, certainly have been put out, but the 

suppressed fires might not have affected much land area 

without the wind and drought that accompanies the often uncontrollable large fires. Although modified by supression, fire dynamics is probably still within the HRV 

(low confidence).

In YNP there is little basis for arguing that fire suppression from 1872 to 1972 had a major effect on the 

magnitude of the 1988 fires over most of the subalpine 

plateaus. While evidence for comparably large fires does 

not yet exist for the MBNF, it is probable that such fires 

have occurred in the Snowy Range and Sierra Madre 

units and that they will occur again, regardless of attempts to control them. In unmanaged forests (such as 

YNP), stand replacing fires may occur in the same stand, 

on average, every 200-400 years; and they may burn over 

a large proportion (perhaps 5-25%) of the landscape approximately once each century. Where landscape patterns and stand structures are modified by humans, as 

in the MBNF, the intervals are more difficult to predict. 

Harvesting often is now done with the objective of reducing fuel loads and flammability as well as providing 

a source of wood. However, wild fires that burn large 

areas probably are determined as much by drought and 

wind as by fuel loadings, if not more so (Johnson and 

Larsen 1991; Bessie and Johnson 1995; Weir and others 

1995; Clark 1990; Baker 2002b). Thus, it appears that 

managed forests at high elevations are just as likely to 

have stand-replacing fires as unmanaged forests.

4.2.2. Insects

Mortality caused by insects has probably been the second most important form of disturbance in high-elevation 

forests, as the spruce beetle (Dendroctonus rufipennis) and mountain pine beetle (Dendroctonus ponderosae) are capable of reaching epidemic population sizes 

in stands of spruce-fir and lodgepole pine. As with fire, 

such epidemics can be detected in the annual growthring patterns of trees, providing insights into forest dynamics in the mid-1800s. No published studies of this 

nature have been done on the MBNF, and therefore our 

current understanding of the HRV for insect populations 

must be derived from general observations and studies 

done elsewhere (Klein and others 1979; Christensen and 

others 1987; Veblen and others 1994; Veblen 2000).

Based on historical reports, the mountain pine beetle 

and other associated bark beetles (e.g., Ips spp.) have 

not been a significant factor in the lodgepole pine forests 

of the MBNF during the last century. Epidemic severity appears to be highest in the northern Rockies, where 

millions of trees have been killed in a single year (Klein 

and others 1979). Cold environments, such as at high elevations in much of the Snowy Range and Sierra Madre, 

may inhibit beetle epidemics (Amman 1977; Amman 

and others 1977; Amman 1978, 1989). Only small areas of low, endemic population sizes exist in the subalpine forests of the MBNF, and no major epidemics have 

been recorded (von Ahlefeldt and Speas 1996). A mountain pine beetle epidemic did affect a large number of 

ponderosa pine at low-elevations in the Laramie Peak 

area, from 1988 to 1994, but few lodgepole pine apparently were affected at high elevations (von Ahlefeldt 

and Speas 1996). Mountain pine beetle epidemics may 

have occurred from time to time in some warmer, low 

to middle elevation lodgepole pine forests. Observations 

suggest that beetle abundance is growing at lower elevations on the northern end of the Snowy Range and in 

the extreme southeastern part of the Sierra Madre (Jerod 

DeLay, personal communication).

Mountain pine beetle epidemics typically start in large 

diameter trees that are diseased or stressed. Most outbreaks occur in stands that are at least 80 years old, with 

trees at least 8 inches (20 cm) dbh (Koch 1996; Schmid 

and Mata 1996). Typically, the larger trees are killed in 

groups of up to 100 (Schmid and Mata 1996). The result is to reduce the average diameter of living trees in 

a stand while creating openings of varying size (Schmid 

and Mata 1996). By killing large trees, beetle epidemics increase the number of standing dead trees that can 

become snags, which eventually fall and become coarse 

woody debris (CWD), thereby improving habitat quality 

for a number of species (Balda 1975; Amman and others 

1977). Also, the canopy gaps stimulate the growth of understory plants and small trees (Stone and Wolfe 1996), 

which creates a ladder of fuel from the forest floor to 

the canopy as they grow taller. Therefore, flammability increases because of the beetles but no studies have 

shown thus far that fire frequency or intensity increases. 

Research has found that the total productivity of the forest can return to pre-epidemic levels within 10-15 years 

(Romme and others 1986).

The frequency and severity of mountain pine beetle 

epidemics vary considerably in the Rocky Mountains, 

depending on environmental conditions such as elevation and climate. On the scale of an entire National 
USDA Forest Service RMRS-GTR-139. 2005 29

Forest, the range of epidemic recurrences is 20 to 40 

years, while return intervals for an individual stand are 

longer (50-100 years) (Johnson 1995; Schmid and Mata 

1996). Duration of individual epidemics may vary from 

one to 10 years, with an average of six (Schmid and Mata 

1996). Amman and others (1977) estimated that, at the 

latitudes of the MBNF, lodgepole pine forests up to 2880 

m (9500 ft) are susceptible to only 25 to 50% mortality, 

while those above 2880 m (2576 m on the Laramie Peak 

unit) would likely incur less than 25% mortality.

As with mountain pine beetles, the spruce beetle 

has not been an important cause of disturbances on the 

MBNF during the last 150 years (von Ahlefeldt and Speas 

1996, p. 6-10), though it has caused extensive mortality elsewhere in the region. The spruce beetle does exist 

at endemic levels in the spruce-fir forests of the MBNF, 

with small outbreaks reported on Green Mountain south 

of Encampment in the Sierra Madre and in the Upper 

French Creek watershed in the Snowy Range. In contrast, much of Colorado and central New Mexico have 

had severe epidemics that have caused heavy spruce 

mortality (Baker and Veblen 1990; Veblen and others 1991, 1994). The two largest recorded outbreaks in 

Colorado occurred in the mid- to late 1800s (Sudworth 

1900; Baker and Veblen 1990) and in the 1940s (Hinds 

and others 1965). Spruce beetle outbreaks can be associated with accumulated logging debris or stands that have 

been subjected to windstorms (Schmid and Frye 1977; 

Veblen 2000).

HRV estimates for spruce beetle epidemics are based 

primarily on studies and observations from Colorado 

and New Mexico. The 1940s epidemic, which killed 

99% of overstory spruce trees across more than 283,500 

ha (700,000 acres) on the White River National Forest, 

is believed to be an extreme example (Hinds and others 1965; Schmid and Mata 1996). Historically, it appears that spruce beetle epidemics kill anywhere from 

5 to 99% of large spruce trees (> 4 inches [10 cm] dbh) 

over areas ranging from small patches to thousands of 

hectares (Schmid and Mata 1996). Based on tree ring records in a part of the White River National Forest, Veblen 

and others (1994) estimated that the average return interval and rotation period for spruce beetle epidemics was 

117 years (range = 111-122 years) and 259 years, respectively, which is more frequent than fire in their study 

area. Because the spruce beetle typically does not attack 

trees less than 70 years old, epidemics are not likely in 

stands that have experienced disturbance in less than 

that amount of time (Veblen and others 1994).

Susceptibility of spruce stands on the MBNF to future spruce beetle epidemics is uncertain. Endemic 

populations are normally found in windthrown or 

otherwise fallen spruce trees (i.e., logging residue), 

and population sizes can increase rapidly after extensive blowdowns (Schmid and Frye 1977). The 1940s epidemic in Colorado, for example, was triggered by an 

extensive blowdown in 1939 (Hinds and others 1965). 

Susceptibility to spruce beetle epidemics has not been 

defined as precisely as for mountain pine beetles, but 

Schmid and Frye (1977) suggested the following four 

characteristics of highly susceptible stands: (1) average dbh >16 inches (41 cm); (2) basal area >150 ft2

 per 

acre (>34.4 m2/ha); (3) >65% spruce in the canopy; and 

(4) locations in well-drained creek bottoms. The effects 

of a spruce beetle epidemic on the MBNF would likely 

be similar to a mountain pine beetle epidemic, with the 

creation of additional snags, more CWD, and a higher 

level of fuel continuity that could increase flammability 

while simultaneously improving habitat for other species. Because spruce beetles do not attack subalpine fir, 

the fir can become more abundant rather quickly after a 

beetle epidemic (Schmid and Hinds 1974).

Numerous other insects are known to occur on the 

MBNF, such as the western spruce budworm, western balsam bark beetle, pandora moth, pine sawflies, 

Engelmann spruce weevils, and needle miners, but thus 

far their effects appear to be minor. Western spruce budworm has been an important cause of disturbance in 

the mountains of Colorado and New Mexico (Swetnam 

and Lynch 1993; Veblen 2000). Contrary to its name, 

the spruce budworm attacks primarily Douglas-fir in the 

MBNF, a species that occurs only on about 2% of the 

MBNF. The spruce budworm has killed some Douglasfir in the southwestern part of the Snowy Range unit, and 

western balsam bark beetle has recently attacked groups 

of 70 or more subalpine fir trees in the Snowy Range and 

Sierra Madre (Jerod DeLay, personal communication). 

These insects and others typically cause mortality or low 

vigor in individual trees or small groups of trees, thereby changing forest structure. The outbreak of the western balsam bark beetle that is just now developing could 

have been triggered by the drought that has been in effect since 1999. As with large fires, insect epidemics occur periodically and may be partially caused by unusual 

climatic conditions (Swetnam and Lynch 1993).

Overall, Schmid and Mata (1996) suggest that insect populations and their effects in the southern Rocky 

Mountains are currently within their HRV, a conclusion 

that appears to merit high confidence considering that 

historic epidemics have not been recorded in the last 150 

years on the MBNF. Also the insects affecting Rocky 

Mountain forests are native and probably have reached 

epidemic population sizes from time to time during the 

last several thousand years. Human influences such as 
30 USDA Forest Service RMRS-GTR-139. 2005

silvicultural treatments and fire suppression have the potential to move epidemic frequency and severity either 

below or above historic levels (Keane and others 2002), 

but periodic episodes of severe insect damage most likely will always be a part of Rocky Mountain forest dynamics.

4.2.3. Disease

Tree pathogens such as dwarf mistletoe (Arceuthobium 

americanum), comandra blister rust (Cronartium comandrae), root diseases, broom rusts, and heart rots are also 

native elements of lodgepole pine and spruce-fir forests 

(Knight 1994a; Johnson 1995; Veblen 2000), often having the effect of reducing tree growth in a stand or causing localized mortality. Consequently they are commonly 

viewed as detrimental to “forest health.” Despite negative effects on short-term productivity, plant pathogens 

have been widely recognized for their positive contributions to biological diversity and various ecosystem processes (Gill and Hawksworth 1964; Hawksworth 1975; 

Tinnen and others 1982; Dinoor and Eshed 1984; Tinnen 

1984; Zimmerman and Laven 1984; Johnson 1995). As 

with other natural causes of disturbance, the diseases 

that currently characterize the high-elevation forests of 

the MBNF probably have always played an important 

role in long-term ecosystem dynamics. Notably, the introduced white pine blister rust (Cronartium ribicola) is 

beginning to infect some limber pine trees in the Laramie 

Mountains and Sherman Mountains units of the MBNF.

The most widespread tree parasite in high-elevation 

forests on the MBNF is dwarf mistletoe, a parasitic vascular plant commonly found on lodgepole pine. The 

effects of dwarf mistletoe on the tree include slowed 

growth, spike top, witches’ broom, dead branches and 

eventual tree death (Hawksworth and Wiens 1972). 

While such effects can have negative consequences for 

silviculture and the safety of forest visitors, they also 

may be beneficial for a variety of animals (Hawksworth 

and Johnson 1989). For example, bird abundance and diversity were found to be higher in ponderosa pine stands 

infected with mistletoe in Colorado (Bennetts and others 1996), and the same could be true for lodgepole pine 

stands.

Because of its potential to reduce timber yields, dwarf 

mistletoe has received considerable attention on the 

MBNF and elsewhere. The mistletoe is not common in the 

Laramie Mountains, but the full range of infection levels 

is found on the Snowy Range and Sierra Madre, especially at relatively low, dry sites (fig. 18; Hawksworth 1958; 

von Ahlefeldt and Speas 1996; Selmants 2000). In the 

earliest surveys of the Rocky Mountains, Mason (1915) 

described the distribution of severe mistletoe damage 

as scattered and localized. Decades later, Hawksworth 

(1958) surveyed the MBNF and found infection rates 

(percentage of trees infected) of 70% in stands that had 

been tie hacked or selectively cut from the 1860s to 1950. 

He also surveyed virgin stands and regenerating burns, 

finding them to be 51% and 27% infected, respectively. 

In 1977, Johnson and others (1978) reported that 60% 

of all lodgepole pine on the MBNF were infected with 

dwarf mistletoe, but they did not describe the spatial 

variation in severity levels. Selmants (2000) found that 

as many as 25% of the lodgepole pine trees in stands developing after clearcutting in the Snowy Range could become infected within 40 years. Kipfmueller and Baker 

(1998b) found mistletoe infection in 51% of the stands 

they examined in the Snowy Range. Interestingly, they 

mapped the distribution of mistletoe severity and found 

that the majority of infected stands had low to moderate 

infection levels, with heavy infection only in scattered, 

localized infection centers. Historically, the spread of 

heavy infection centers was limited most likely by standreplacing fires. Where such fires kill all infected trees, 

they essentially eliminate dwarf mistletoe from the stand 

(Hawksworth and Wiens 1972).

The relationship between fire and dwarf mistletoe, 

however, is complex. Where infected trees survive a fire 

or timber harvest, the mistletoe can infect new seedlings, 

creating new infection centers. Notably, some old stands 

are not infected (Kipfmueller and Baker 1998b). Given 

the generally slow spread of dwarf mistletoe (less than 

one meter per year) (Johnson 1995), it is uncertain if the 

effect of fire suppression on infection rates can be detected (Kipfmueller and Baker 1998b). Overall, the severity and distribution of dwarf mistletoe infection on 

the MBNF is likely within the HRV (moderate confidence), with the possible exception of some severely infected areas associated with selective timber harvests 

that did not remove infected trees. 

Two other diseases have received a considerable 

amount of attention on the MBNF—comandra blister rust and armillaria root disease. Although their life 

cycles are different, they both cause varying amounts 

of deformity and mortality to lodgepole pine (Johnson 

1995). Lundquist and Lister (1991) found that 27-35% 

of the stands examined in the Laramie District of the 

Snowy Range unit were infected with comandra blister rust. This compares to an infection rate of over 50% 

in the Wind River Mountains (Geils and Jacobi 1984). 

Comandra blister rust requires specific site conditions 

(Geils and Jacobi 1991), including proximity to its secondary host plant, Comandra umbellata. Losses from 

armillaria root disease appear to have been minor on the 
USDA Forest Service RMRS-GTR-139. 2005 31

Figure 18a—Distribution of varying levels of mistletoe infection on the Snowy Range unit of the MBNF. Source: MBNF.
32 USDA Forest Service RMRS-GTR-139. 2005

MBNF (Johnson 1995). Armillaria is most common on 

Douglas-fir and subalpine fir.

Other pathogens present in the high-elevation forests 

of the MBNF appear to have had minor effects, though 

all diseases undoubtedly interact with other disturbances and various ecological processes (Zimmerman and 

Laven 1984). Overall, most stands are likely to have 

numerous pathogens as concurrent members of the biological community, which helps to ensure the diverse 

elements of forest structure required for the biological 

diversity that has existed for millennia (Lundquist 1993, 

1995a, 1995b). Except where exotic pathogens such as 

the white pine blister rust are found, there is no basis for 

saying that the current abundance of the native diseases 

present on the MBNF at higher elevations is beyond the 

HRV (moderate confidence),

4.2.4. Wind

Because of high elevation, latitude and location, the 

MBNF is often subject to very high winds, with recorded 

Figure 18b—Distribution of varying levels of mistletoe infection on the Sierra Madre unit of the MBNF. Source: MBNF

gusts over 160 km/hr (100 mph) (Marston and Clarendon 

1988). Such winds may topple trees in small clumps or 

over thousands of hectares (Alexander 1964; Veblen 

and others 1991). In 1987 a wind storm blew down 6000 

hectares (14,820 acres) of forest in the Teton Wilderness 

(Knight 1994a), and in 1997 winds of 200-250 km/hr 

(125-155 mph) toppled >10,000 ha (24,700 acres) of 

forest on the western slope of the Routt National Forest 

in Colorado, just south of the MBNF (Baker and others 

2001; Lindemann and Baker 2001, 2002). Two high-elevation blowdowns have been recorded on the MBNF during the past 100 years (von Ahlefeldt and Speas 1996).

Because large blowdowns are relatively rare, their effects are largely unknown. Such disturbances are even 

more stochastic than fires and an estimation of return intervals is not now possible, nor are estimates of the HRV 

for the effects of blowdowns. We can safely say, however, that strong winds will continue to be an important factor in shaping landscape patterns and the stand structure 
USDA Forest Service RMRS-GTR-139. 2005 33

of high-elevation forests. Such disturbances probably are 

within the HRV (high confidence).

Blowdowns often interact with other kinds of disturbances. For example, trees weakened by fire, insects, 

and disease may be more susceptible to breakage by 

wind; and recently downed trees (Hinds and others 1965; 

Schmid and Frye 1977) may trigger spruce beetle epidemics. Flammability surely is increased by a blowdown 

as well. Wind fall has been a common problem associated 

with timber harvest, both along clearcut edges (Alexander 

1964; Vaillancourt 1995) and in selectively cut stands 

where a high percentage of the trees are removed (von 

Ahlefeldt and Speas 1996, page 4-29 and 6-44). Many 

trees in a stand are not toppled during blowdowns, making the effect more similar to a selective harvest than a 

clearcut (Lindemann and Baker 2001).

Recent studies of the Routt-Divide blowdown in 

northcentral Colorado indicate that various features of 

the physical environment and the vegetation determine 

where blowdowns are likely to occur. Physical factors 

seemed to be most influential, especially wind exposure, 

elevation, and aspect (Lindemann and Baker 2002). 

Unexpectedly, soil permeability, soil water-holding capacity, and distance to natural edges were not significantly correlated with the blowdown pattern. The effect 

of cover type varied with elevation. At lower elevations, 

aspen forests were less susceptible to blowdown and 

spruce-fir forests were more susceptible. At higher elevations, closer to the Continental Divide, younger stands 

were less susceptible than older stands (Baker and others 

2001). These results suggest that both topography and 

vegetation structure can influence the extent and pattern of damage, but that the effects vary with elevation 

and wind intensity. In another study of the Routt-Divide 

blowdown, Lindemann and Baker (2001) found that 

blowdown patches were larger and had more compact 

shapes than patches in the surrounding forest (which 

were derived from past fires, insect epidemics, topographic influences, and possibly previous blowdowns). 

They also discussed the potential for a spruce beetle outbreak in the vicinity of the blowdown, concluding that it 

is difficult to predict the spread of the possible outbreak 

because it can continue for a decade or more and the 

beetles can fly long distances (300 m or more in Alaska; 

Werner and Holsten 1997).

4.2.5 Timber harvesting

To some degree, harvesting has effects that are similar to fires and other natural disturbances. For example, 

canopy gaps of varying sizes are created. However, timber harvesting is different in other ways. Indeed, HRV 

analyses are motivated to a large degree by the fact that 

harvesting affects an ecosystem differently than natural disturbances. To illustrate, Muir (1993) found that 

lodgepole pine in western Montana was likely to be 

more dominant after a fire than after a clearcut, and that 

within-stand variability in tree age is less after a fire than 

after clearcutting. Also, fire tends to leave much of the 

large wood, while timber harvesting commonly removes 

it. In the sections that follow, we summarize the predominant silvicultural practices on the MBNF through 

time, and then compare the effects of timber harvesting 

on stand structure and landscape structure to the effects 

of natural disturbance regimes.

Until about 1950, selective or partial cutting in lodgepole pine or mixed conifer stands was the predominant 

form of timber harvest on the MBNF. Native Americans 

selected individual trees that would meet their needs, 

such as for lodge poles. Usually they harvested small 

trees. Impacts from their harvesting were undoubtedly 

minor compared to those of European-American after 

the mid-1860s, when trees of 10-14 inches dbh (25-36 

cm) were cut for railroad ties (tie hacking) and telegraph 

poles. Tie hacking was unregulated and became widespread, especially in the Snowy Range unit (see fig. 7). 

Mullison and Lovejoy (1909) reported that the Medicine 

Bow Mountains yielded 320 million board feet (mmbf) 

of wood from 1868-1902.

Although selective harvesting is typically not standreplacing, the majority of lodgepole pine and sprucefir forests in the Snowy Range and Sierra Madre date 

back to the mid- to late-1800s (see figs. 19a and b). 

The combined effect of selective harvest and fires during this period apparently led to the initiation of many 

new stands. Young and others (1989, 1994) found evidence on the MBNF suggesting that riparian forests 

along streams commonly used for floating railroad ties 

had also been cleared. Specifically, they found evenaged, early successional lodgepole pine forests with no 

large trees or snags along such streams. Along streams 

that did not have tie drives, the stands were structurally diverse and older, with mixed age and species composition.

During the early years of National Forest administration (1902-1949), methods for harvesting became regulated and gradually evolved toward what they are today. 

As summarized by von Ahlefeldt and Speas (1996), tree 

harvesting proceeded as follows:

1902-1914—Selective removal of live and dead trees, 

taking no more than 35% of a stand, with piling 

and burning of slash to reduce fire hazard.

1914-1922—Removal of approximately 35% of the 

trees 10-15 inches dbh (25-38 cm), with continued 

slash burning.
34 USDA Forest Service RMRS-GTR-139. 2005

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Snowy Range

total lodgepole pine acres = 272,318

acres with no age data = 28,064 (10%)

Sierra Madre

total lodgepole pine acres = 136,092

acres with no age data = 21,946 (16%)

Sherman Mountains

total lodgepole pine acres = 9,036

acres with no age data = 1,194 (13%)

Laramie Mountains

total lodgepole pine acres = 58,334

acres with no age data = 37,171 (62%)

Figure 19a—Age distribution of lodgepole pine forests on the four units of the MBNF. Bars represent 10-year age classes. 

Source: MBNF.
USDA Forest Service RMRS-GTR-139. 2005 35

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Snow Range

total spruce-fir acres = 121,373

acres with no age data = 5,528 (5%)

Sherman Mountains

total spruce-fir acres = 22,233

acres with no age data = 5,449 (25%)

Laramie Mountains

total spruce-fir acres = 2,482

acres with no age data = 270 (11%)

Sierra Madre

total spruce-fir acres = 85,195

acres with no age data = 38,366 (45%)

Figure 19b—Age distribution of spruce-fir forests on the four units of the MBNF. Bars represent 10-year age classes. 

Source: MBNF.
36 USDA Forest Service RMRS-GTR-139. 2005

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Snowy Range

total aspen acres = 30,636

acres with no age data = 9,346 (31%)

Sherman Mountains

total aspen acres = 36,138

acres with no age data = 17,160 (47%)

Laramie Mountains

total aspen acres = 2,538

acres with no age data = 1,589 (63%)

Sierra Madre

total aspen acres = 79,384

acres with no age data = 38,340 (48%)

Figure 19c—Age distribution of aspen woodlands on the four units of the MBNF. Bars represent 10-year age classes. 

Source: MBNF. 
USDA Forest Service RMRS-GTR-139. 2005 37

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Snowy Range

total ponderosa pine acres = 1,130

acres with no age data = 1,151 (87%)

Laramie Mountains

total ponderosa pine acres = 84,507

acres with no age data = 50,083 (59%)

Sherman Mountains

total ponderosa pine acres = 9,220

acres with no age data = 3,979 (43%)

Sierra Madre

total ponderosa pine acres = 28

acres with no age data = 10 (36%)

Figure 19d—Age distribution of ponderosa pine woodlands on the four units of the MBNF. Bars represent 10-

year age classes. Source: MBNF.
38 USDA Forest Service RMRS-GTR-139. 2005

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total Douglas-fir acres = 6,975

acres with no age data = 2,373 (34%)

Sherman Mountains

total Douglas-fir acres = 723

acres with no age data = 0

Laramie Mountains

total Douglas-fir acres = 1,604

acres with no age data = 96 (6%)

Sierra Madre

total Douglas-fir acres = 2,374

acres with no age data = 908 (38%)

Figure 19e—Age distribution of Douglas-fir woodlands on the four units of the MBNF. Bars represent 10-year age 

classes. Source: MBNF.
USDA Forest Service RMRS-GTR-139. 2005 39

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Snowy Range

total limber pine acres = 1,520

acres with no age data = 507 (33%)

Sherman Mountains

total limber pine acres = 393

acres with no age data = 10 (3%)

Laramie Mountains

total limber pine acres = 9,874

acres with no age data = 9,268 (94%)

Sierra Madre

total limber pine acres = 198

acres with no age data = 182 (92%)

Figure 19f—Age distribution of limber pine woodlands on the four units of the MBNF. Bars represent 10-year 

age classes. Source: MBNF.
40 USDA Forest Service RMRS-GTR-139. 2005

1922-1936—At least 45% of trees larger than 10 inches dbh (25 cm) were cut, with many larger sawlogs 

left on the ground; some but not all slash was piled 

and burned.

1936-1940—Only trees that could be used for ties were 

removed; 65-85% of stand volume could be taken. 

The demand for railroad ties gradually diminished, 

with the last tie drive on the MBNF in 1940.

After about 1950, clearcutting became more common because of an increased emphasis on maximizing 

wood production for construction purposes, plus a desire by some silviculturists to restructure stands that had 

been tie-hacked or high-graded. In general on National 

Forests, old forests with comparatively little wood production commonly were clearcut to produce even-aged 

stands of rapidly growing trees that could be more easily subjected to thinning and other management activities (Langston 1995). The clearcuts were thought to be 

similar to large fires, one of the kinds of disturbances 

with which the forest species had evolved. Indeed, both 

fires and clearcuts do create a patchy landscape mosaic 

(though with different edge characteristics), and lodgepole pine has proven to be well adapted for establishment 

in the patches created by both types of disturbances. In 

some areas in the Rocky Mountains, Engelmann spruce 

and subalpine fir can be important pioneer species, e.g., 

in Grand Teton National Park (Doyle and others 1998). 

Tree regeneration, without planting, has been very good 

on most sites below 2888 m (<9,500 ft) elevation. In 

fact, lodgepole pine commonly regenerates so abundantly on the MBNF that precommercial thinning is common about 20 years after a clearcut (Grasso 1981; Geils 

and Jacobi 1991).

In contrast, new tree establishment is much slower 

near treeline. The last natural disturbance in such areas was probably a fire or windstorm that occurred several centuries previously. Climatic conditions there are 

marginal for tree growth, especially when high winds 

blow the snow into downwind forests, leaving insufficient snow to protect tree seedlings from winter desiccation and high-intensity solar radiation. Some seedlings 

do survive, but natural reforestation occurs more slowly 

than expected.

The 1985 management plan for the MBNF (page III45) specifies that the rotation age for clearcutting in subalpine/montane landscapes is 90-140 years for stands of 

lodgepole pine, 80-120 years for stands of aspen, and 

100+ years for stands of other cover types. For two- and 

three-step shelterwood cuts, the rotation ages are 100-

180 for stands dominated by Engelmann spruce and 

subalpine fir, 90-140 years for lodgepole pine, and 100+ 

for other cover types. If a mean fire-return interval of 

150-200 years is assumed for natural disturbances such 

as fires (Romme and Knight 1981, Romme and Despain 

1989), then the silvicultural rotation ages in general tend 

to be shorter by possibly 10-100 years, depending on 

the mean fire-return interval at specific sites. Other approaches have been recommended for characterizing fire 

interval (Kipfmueller and Baker 2000; Baker and Ehle 

2000; Howe and Baker 2003), and would give varying 

results, but fire rotations generally appear to be longer 

than silvicultural rotations.

The stand age at which harvesting usually occurs also 

is different from that for natural disturbances. For example, both mountain pine beetle and spruce beetle typically affect stands that are at least 70-80 years old. In 

contrast, precommercial thinning is often prescribed for 

20- or 30-year-old stands (Geils and Jacobi 1991; Koch 

1996). Further variation from natural disturbances also 

has resulted from machine yarding, where whole trees 

are brought to the landing before delimbing. This results 

in a concentration of slash and residual coarse woody 

debris on the edges of the clearcuts. Following natural 

disturbances, this material would be distributed more 

evenly across the affected area. The machinery used for 

this kind of yarding, as well as for slash treatments such 

as piling and burning, roller chopping, and tractor walking, can scarify the soil in ways that favor the establishment of lodgepole pine and Englemann spruce seedlings. 

However, such scarification may also slow the recovery 

of some understory plants, such as grouse whortleberry (Vaccinium scoparium) (Ahlefeldt and Speas 1996; 

Selmants 2000).

4.3. Stand structure of high-elevation forests

Silvicultural treatments are designed partially, if not 

entirely, to manipulate stand structure in a way that favors the growth of certain species. Selective harvesting, 

for example, creates small gaps that are similar to natural gaps caused by the death of a single tree or several 

neighboring trees. Light penetration to the forest floor 

is increased and competition for water and nutrients is 

reduced, thereby enhancing the growth of the remaining trees as well as some understory plants. Selective 

harvests create or maintain forests with an uneven-aged 

structure. Timber harvests also have been prescribed to 

improve wildlife habitat, such as by increasing the abundance of aspen, or to increase streamflow by catching 

more snow in small patch cuts where less of the snow 

water is used by trees.

Stand structure can be described in terms of variables 

or attributes such as the following: tree density, percent 

canopy cover, the density of canopy gaps, the density 

and cover of understory plants, the ratio of above-ground 
USDA Forest Service RMRS-GTR-139. 2005 41

biomass and below-ground biomass, tree species diversity, total plant diversity, the number of trees in a series of 

age classes and size classes, forest floor depth, the density of snags, and the amount and distribution of coarse 

woody debris. Small-scale patches or gaps of one to several hundred square meters, such as those created by the 

death of one or a few trees, are an aspect of stand structure. When discussing landscape structure, larger gaps 

of one to many hectares are considered. Both stand and 

landscape structure change through time as well as space 

because of natural and human-caused influences.

Regrettably, there are essentially no data on stand 

structure during the reference period (1600-1850). HRV 

analyses are still possible, but they must be tentative and, 

as noted, they are based on what is known about how 

current stand structure changes with time during succession and spatially from one place to another, along with 

impressions of what the forests must have been like prior 

to the arrival of European-Americans (based on the journals of early explorers and information about the biology of the dominant species). In the synopsis that follows, 

we present our conclusions on whether or not past management practices are causing key ecosystem variables 

to exceed our best understanding of their HRV. For each 

conclusion, we indicate our confidence level. Our analysis could not be site specific, but the information that we 

present can be used when developing site-specific management plans. The focus in this section is on forests 

dominated by lodgepole pine, Engelmann spruce, and 

subalpine fir. Aspen groves are considered briefly but 

are discussed more fully in section 6. Ponderosa pine, 

Douglas-fir, and limber pine woodlands are considered 

in section 5.

As noted previously, HRV analyses should consider 

the spatial and temporal variability of specific variables. 

Based on the available information reviewed in previous 

sections, the nature of silvicultural practices, and our 

current understanding of ecosystem dynamics, the summary statements that follow seem reasonable for variables pertaining to stand structure in the high-elevation 

forests of the MBNF.

4.3.1. Tree density

The naturally occurring number of trees per hectare is highly variable through space and time, ranging 

from a few hundred mature trees in open stands to over 

15,000 in 100-year-old doghair stands of lodgepole pine. 

Various studies have shown that high lodgepole pine 

seedling densities are partially a function of the number 

of serotinous trees in a stand and fire intensity (Franklin 

and Laven 1989; Anderson and Romme 1991; Turner 

and others 1997). Ten years following the 1988 fires in 

YNP forests, tree seedling density ranged from essentially no seedlings to extremely high densities (sometimes exceeding 50 seedlings per square meter). With 

succession, densely stocked stands are slowly thinned 

by natural mortality, but many very dense “doghair” 

stands persist for well over 100 years—probably until 

the next stand-replacing disturbance. The same occurs 

in the MBNF.

When thinning is done for enhancing tree growth, 

the goal is to have approximately 1500 trees per hectare (860-1240 trees/ha [2124-3063 trees/acre] at age 20 

on the Laramie District [Geils and Jacobi 1991]). It is 

doubtful that timber harvesting creates tree densities in 

stands that are higher or lower than the HRV (high confidence), but there may be more stands of low tree density in the landscape than during the reference period 

because of silvicultural thinning (moderate confidence), 

as discussed in section 4.4.7. Where thinning has occurred, the trees probably are more evenly distributed 

than would occur otherwise, simply because even spacing is considered to be advantageous for tree growth.

4.3.2. Seedling and sapling density

Engelmann spruce, subalpine fir, aspen, and lodgepole pine can all be pioneer species in some situations, 

though lodgepole pine usually is most common in the 

early years of succession on the MBNF. The establishment of new seedlings for all of these species can 

be quite slow or surprisingly rapid (Turner and others 

1997). Seedling establishment on high-elevation sites 

with exposure to west and northwest winds is typically slower than elsewhere, regardless of the type of disturbance. Aspen persists in the stands primarily through 

root sprouting, though seedlings were abundant in some 

areas after the 1988 fires in YNP (Romme and others 

1995). Spruce and lodgepole pine are thought to prefer 

mineral soil for seedling establishment, but they often 

become established where competition for water and nutrients is reduced even if the forest floor remains intact, 

such as after a wind storm, insect epidemic, a timber 

harvest with minimal disturbance to the forest floor, or 

less intense fires. There is no evidence to suggest that 

seedling and sapling density is beyond the HRV (high 

confidence). The density of lodgepole pine seedlings 

following both fires and clearcutting is often very high 

(Grasso 1981; Geils and Jacobi 1991).

4.3.3. Percent canopy cover

There is no basis for arguing that processes involved 

with natural forest development and succession lead to 

maximizing the rate of tree growth, or even total forest 

productivity (annual tree growth plus annual understory 
42 USDA Forest Service RMRS-GTR-139. 2005

plant growth). Biologically, plant adaptations facilitate 

persistence and the production of seed more than growth 

rates. By thinning forests to maximize tree growth, silviculturists create a canopy that is more open than would 

have occurred otherwise, thereby most likely pushing 

percent canopy cover below the range of means for this 

variable during the HRV period in forests of comparable 

age and site conditions (moderate confidence). This effect may last no more than about 10 years, as the canopies of the remaining trees normally expand to fill the 

openings that were created. Considerations of canopy 

cover should include the cover provided by standing 

dead trees (snags), such as following a fire, as this kind 

of cover provides habitat for some animals and probably 

is important for the establishment of some plant seedlings.

4.3.4. Density and size of canopy gaps

Canopy gaps occur episodically in all forests. If a 

sufficiently large group of neighboring trees die, perhaps 15-30, a root gap is created as well, where competition for water and nutrients is reduced significantly 

(Parsons and others 1994). Gaps, large or small, have a 

great influence on the growth of understory plants (including saplings), and indirectly they affect age class 

structure, size class structure, species diversity, and the 

abundance of some kinds of animals and other organisms. Silvicultural systems that include group selection 

and thinning could create canopy and root gaps that are 

similar to those found in unmanaged forests if not too 

many trees are harvested. However, as with percent canopy cover, the number and size of gaps in a managed forest is likely to be higher than the range of means for this 

variable during the HRV period (moderate confidence), 

simply because the management goal is often to foster 

rapid tree growth.

4.3.5. Density and cover of understory plants

If managed forests have lower canopy cover or a high 

density of canopy gaps, the density and cover of understory plants probably will be higher. However, timber 

harvesting can leave an abundance of slash that could 

suppress some understory plants, as could the soil compaction, scarification, or other changes caused by the 

machinery that might be used. Species composition can 

be affected by slash treatments that disrupt the soil profile, which reduces the sprouting capacity of some understory plants and can be favorable for the invasion of 

new species with wind-borne propagules. Some exotic 

plants can invade both managed and burned forests, but 

thus far they apparently do not persist in 30- to 50-yrold post-harvest stands at high elevations in the MBNF 

(Selmants 2000). Exotic plants often are common in road 

ditches, but they have not yet invaded the adjacent forests in great numbers on the MBNF (Dillon 1998). Given 

that the density and cover of understory plants is highly 

variable in natural stands as well as in managed stands, 

there is no indication, to date, that timber harvesting 

would push variables associated with the understory vegetation beyond the HRV (moderate confidence). 

Some plant species may be more sensitive to timber 

harvesting than to fire, wind storms, or other natural 

disturbances, and, of course, the road building associated with harvesting decimates the native vegetation and 

often provides habitat for introduced, weedy species.

4.3.6. Ratio of aboveground biomass to 

belowground biomass

The root/shoot ratio can vary from about 0.2 to 0.3 

in natural lodgepole pine forests, with high values more 

likely to occur in high tree density stands (Pearson and 

others 1984). Silvicultural practices have an effect on 

tree density, but shifts in root/shoot ratios attributable to 

harvesting probably are not large enough to exceed the 

HRV (high confidence).

4.3.7. Tree species and genetic diversity

Apparently all of the tree species present in the 1700s 

and 1800s are still very common on the MBNF. Initially, 

high grading through selective cutting may have left an 

abundance of trees with less desirable wood characteristics, such as subalpine fir, or a large number of malformed 

or slow growing trees; and more recently the practice of 

clearcutting may have increased the abundance of lodgepole pine. However, fires may have caused some of the 

same shifts. No new tree species have been introduced. 

Research is insufficient to indicate that harvesting during the last century has changed the genetic diversity of 

stands, though if fire suppression has been effective, or 

becomes more effective in the future, the abundance of 

serotinous lodgepole pine trees in managed forests may 

decline. Currently, the abundance of stands with at least 

some serotinous lodgepole pine is high (USFS data; Phil 

Krueger, personal communication). Therefore, tree species and genetic diversity probably are within the HRV 

(high confidence).

4.3.8. Total plant species diversity

The species diversity for any group that exists in any 

ecosystem is determined in part by disturbances. Timber 

harvesting, with the concomitant removal of CWD and 

the effects of machinery and slash disposal practices, is 

a new kind of disturbance to which some species may 

not be adapted. However, there are no data to indicate 
USDA Forest Service RMRS-GTR-139. 2005 43

that total plant diversity is lower or higher than the HRV 

(moderate confidence). Some species may be rare and 

easily extirpated from an area, such as various species 

of orchids, but evidence for this happening over a large 

area is not now available (Selmants 2000). Some plants 

are thought to be interior forest species that could be adversely affected because of depth-of-edge effects associated with road building and patch cutting (Dillon 1998; 

Baker and Dillon 2000), but the depth-of-edge effects 

on such species, beyond road ditches, are difficult to determine and are not well known. The cover of grouse 

whortleberry declines in proportion to the intensity of 

slash treatment, but this low shrub persists in the stand 

and gradually increases in abundance (Selmants 2000). 

Post-harvest stands dominated by lodgepole pine at lower elevations appear to recover more rapidly than stands 

of spruce and fir at higher elevations (Selmants 2000). 

Many stands have been subjected to at least two entries 

for the harvesting of large trees, with the first one occurring around the turn of the century. In such stands, the 

species that are present today may be the species that can 

tolerate timber harvesting practices. Additional study is 

required to determine which species cannot.

4.3.9. Age-class structure of stands

Our current understanding of high-elevation forests 

suggests that a full range of age-class structures must 

have existed in the 1700s and 1800s, from even-aged 

stands of lodgepole pine that had developed after fires 

in the 1600s and 1700s to older uneven-aged stands that 

had not burned for centuries and were dominated primarily by Engelmann spruce and subalpine fir (possibly mixed with lodgepole pine and aspen). As noted 

previously, a common approach for management since 

the mid-1900s has been to convert uneven-aged stands 

of older forests to even-aged stands of trees that grow 

comparatively rapidly. However, the variety of age class 

structures on the MBNF today, at the stand level, probably is within the HRV in most areas (moderate confidence). The proportion of the landscape in forests of 

different ages is discussed in section 4.4.

4.3.10. Size-class structure of stands

As with age structure, there must have been a great 

diversity of size-class structures in the high-elevation 

forests during the reference period, ranging from stands 

with uniformly small trees to stands with very large trees 

coexisting with small and mid-sized trees. The size-class 

structure of stands would change through time and would 

vary across the landscape. Fires at high elevations prior to the mid-1800s would have killed many of the big 

trees from time to time, and probably they would have 

burned over large areas (possibly 25-50% or more of the 

Snowy Range or Sierra Madre units, though there is still 

no evidence for such fires on the MBNF; ~36% of YNP 

burned in 1988 by either surface or crown fires). Both 

fire and timber harvest have continued to convert areas 

with large, old trees to stands of younger and smaller 

trees, with the important exception that timber harvesting can extend into areas that had not burned for centuries (such as on leeward slopes or in ravines, both of 

which are thought by some to burn less frequently). In 

such areas, the size-class structure is probably outside 

the HRV because insufficient time has elapsed to allow 

the growth of new large trees to replace those that were 

harvested (moderate confidence).

Across the rest of the landscape, where disturbances 

were frequent during the reference period, the variety of 

size-class structures, at the stand level, is probably within the HRV because both clearcutting and intense fires 

would create even-aged stands (moderate confidence). 

Similarly, insect epidemics and partial cutting, where 

they occurred, could create a structure characteristic of 

uneven-aged stands. The unnatural size-class structure 

created by seed tree and shelterwood silvicultural systems is a short-lived variation because the residual trees 

typically are cut during the second entry, creating an 

even-aged stand.

With regard to temporal variation in the size-class 

structure, it is known that lodgepole pine sapling density 

after both fire and timber harvesting can vary from extremely high in some areas to very low elsewhere (Turner 

and others 1997); and that natural thinning of the pine is 

slow to occur. This is illustrated by the commonly observed “doghair” stands that have great uniformity and 

can persist for 130 years or more. With time, either a 

disturbance occurs that changes the size-class structure 

dramatically or some natural thinning surely must occur. Because this thinning is very slow to occur in stands 

of lodgepole pine, however, the kind of thinning recommended by forest managers to promote growth of the 

residual trees could cause changes that are beyond the 

HRV (moderate confidence). Notably, size-class structure appears to be more variable spatially than through 

time, although this surely depends on the initial density 

of the post-fire or post-harvest seedlings. Some stands 

that have an initial low density become more dense as 

new trees become established, while high density stands 

either change hardly at all or become less dense—at 

least after a century or more.

4.3.11. Forest floor depth

High-elevation forests in the Rocky Mountains 

are characterized by an abrupt boundary between the 
44 USDA Forest Service RMRS-GTR-139. 2005

mineral soil and the forest floor, which is comprised 

mostly of fine and coarse litter in varying stages of decomposition. The abrupt boundary is caused by periodic fires that burn the forest floor and expose much of the 

mineral soil, after which a new forest floor gradually accumulates because of the slow rate of decomposition (Fahey 1983; Fahey and Knight 1986). Because of 

the absence of earthworms in coniferous forests, the litter does not become well incorporated into the mineral 

soil. Consequently, just as there is considerable variation 

in the ages of the forests on the Snowy Range and Sierra 

Madre units, there is considerable variation in the forest 

floor depths of stands.

In addition to time since disturbance, variation in forest floor depth is caused by downslope litter movement 

due to gravity and flowing water, which creates deeper 

forest floors on lower slopes than on or near ridgetops. 

Moreover, fallen logs often accumulate detritus on their 

upper side, and squirrels create middens of cone scales 

under some trees. Usually the forest floor is 2-6 cm thick, 

but for a variety of reasons, including tree fall patterns 

and middens created by squirrels, the HRV is broad, 

ranging from essentially no forest floor to a depth of 

several decimeters. Forest managers often recommend 

that much of the forest floor be scarified or removed, 

thereby creating a better seed bed for new tree establishment (Alexander 1966, 1987a); or they tend to suppress 

fires and minimize disturbance to the forest floor, thereby allowing for a continuation of forest floor thickening—at least until the next fire that cannot be controlled. 

In either case, the range of forest floor depths is within 

the HRV in stands over most of the landscape, although 

small areas may be above or below the HRV due to management activities (moderate confidence).

4.3.12. Mineral soil affected

In general, the forest floor is important for minimizing soil erosion because it increases the rate of infiltration 

during periods of snowmelt and heavy rainfall. While a 

natural fire can expose the mineral soil just as a timber 

harvesting operation can, a notable difference between 

the two kinds of disturbances is the depth to which the 

mineral soil under the forest floor is disrupted (Bockheim 

and others 1975; Halpern and Franklin 1990; Dion 1998). 

Scarification by a bulldozer blade or the blades of a roller-chopper disrupts soil structure to a greater depth than 

does a fire, which normally burns the litter without breaking up the mineral soil (though some localized erosion 

and soil heating may occur). In the process of mechanical 

scarification, the surface roots of surviving trees and understory plants are commonly broken, which could lead 

to higher mortality than would be expected for plants that 

often survive fires (Selmants 2000). Fires typically burn 

plant shoots severely, but except in patches where fuels 

are exceptionally heavy, the soil temperature typically is not hot enough to kill the roots. Thus, most plants, 

other than the conifers, survive because they are capable 

of sprouting from underground tissues that are not affected, especially in montane forest (Selmants 2000).

Therefore, mechanized scarification, where it occurs, creates a new kind of disturbance and the HRV for 

surface soil disruption is exceeded (high confidence). 

Lodgepole pine and Engelmann spruce seedling densities can be higher if the mineral soil is exposed, whether 

by fire or mechanical scarification (Alexander 1987a).

4.3.13. Snag density

The importance of standing dead trees (snags) as a 

component of habitat is widely accepted (Hutto and others 1992; Hutto 1995; Bull and others 1997, Parks and 

others 1997; Hagan and Grove 1999). Management prescriptions now specify the number of snags that should 

be left per hectare, as well as the number of live trees 

that will become the snags of the future (green tree retention). In the absence of such prescriptions, and in 

stands with a history of timber harvesting, there are 

fewer snags now than prior to the arrival of European 

Americans (Harris 1999) because every natural disturbance, with the possible exception of extraordinary wind 

storms, creates many dead standing trees. Early photographs often show an abundance of snags, giving forests 

a “salt and pepper” appearance. Native Americans probably used very few standing dead trees, at least those of 

large size. However, with the advent of thinning, partial cutting, stand improvement, firewood gathering, and 

wood harvesting in general, the rate at which fallen or 

removed snags are replaced with new snags must have 

been reduced. Past harvesting on the MBNF surely has 

lowered snag density to below the range of means for this 

variable in stands where harvesting has occurred (high 

confidence). Since about 1950 on the Snowy Range and 

Sierra Madre units, wood harvesting has occurred on approximately 35% and 23% of the forested land, respectively, which was judged suitable for harvesting. Before 

1950, harvesting was done in other areas as well.

4.3.14. Coarse woody debris

Natural disturbances do not remove large pieces of wood from forests. Even after an intensive fire, 

most of the wood remains in the form of dead standing trees (snags), which becomes coarse woody debris 

(CWD) within two or three decades (Lotan and others 1985; Tinker 1999; Laudenslayer and others 2002). 

The leaves, branches, and smaller wood are consumed 
USDA Forest Service RMRS-GTR-139. 2005 45

by fires, but Tinker and Knight (2000) found that only 

about 8% of the wood >7.5 cm was burned after an 

intense fire in Yellowstone National Park. Other estimates for the proportion of CWD consumed by canopy 

fires range from 12 to 65% (Tinker 1999). With time, 

the downed wood becomes incorporated into the forest floor (Graham and others 1994), creating the impression that many 100+-yr-old stands have very little 

CWD. However, raking the forest floor reveals the remains of many decomposing logs. New CWD debris is 

added as the larger trees die and fall, one by one during 

stand development or in large numbers during a windstorm or after the next fire.

Tinker and Knight (2000), using YNP and MBNF 

data, concluded that the amount of CWD remaining after several clearcuts is less than after the same number 

of fires on comparable sites. Moreover, and as might be 

expected, they found that the time required for CWD to 

cover 100% of the ground surface area in a stand after a 

fire was much less than after a clearcut, and that this time 

interval became similar for a clearcut and fire only when 

slash amounts left on the ground were doubled from average amounts left using current management practices 

(in other words, 200% of current CWD amounts in slash 

could keep practices within the HRV on some sites).

Thus, harvesting is fundamentally different from any 

kind of natural disturbance and produces a forest that 

is beyond the range of means during the HRV period 

for both snags and CWD (high confidence). Ecosystem 

processes associated with dead trees and large pieces of 

wood also could be different than during the reference 

period. USFS guidelines include recommendations for 

the amount of CWD that should be left during a harvest 

(Graham and others 1994), but still, the recommended 

amounts will be lower after several rotations of timber 

harvesting than after the same number of fires on comparable sites (Tinker 1999). CWD can be found elsewhere 

on the landscape, such as in forests withdrawn from harvesting, but the CWD dynamics in specific stands is 

changed by timber harvesting.

The consequences of such deviations from the HRV 

are poorly understood, but CWD provides important habitat for some species of fungi and a variety of 

small mammals and insects, all of which can be important sources of food for vertebrates (Harmon and others 

1986; Hagan and Grove 1999; Laudenslayer and others 

2002); and the organic compounds derived from decomposing wood must have influenced the development of 

soils on which wood production in the future depends. 

CWD is also known to be important for diversifying the 

structure of streams (Laudenslayer and others 2002).

4.4. Landscape structure at high elevations

The spatial and temporal variability of ecosystem variables can also be considered at the landscape 

scale, where, as noted previously, individual stands are 

the patches in a mosaic that covers square kilometers. 

Stand-replacing fires, wind storms, some insect epidemics, road construction, and the silvicultural practices of 

patch-cutting and clearcutting are the primary causes of 

landscape changes. Such changes are obvious from aerial photographs and satellite images, and they are known 

to influence the movement of animals, water, nutrients, 

and sediments across the landscape. The landscape mosaic also can affect the spread of disturbances and subsequent succession (Turner 1980; Knight 1987; Turner 

and others 2001), such as after the fires in YNP (Turner 

and others 1997).

The mosaic of stands included in a landscape can be 

described in terms of patchiness, the number and proportions of different kinds of land cover types (landscape 

diversity), the amount of land in edge and interior forest 

habitat, the proportion of the landscape in different land 

cover types and successional stages (or stand ages), the 

ratio of forest to non-forest land, the rate at which new 

patch types are formed on a landscape, and various spatial statistics. As with stand structure, the characteristics 

of landscapes during our reference period (1600-1850) 

must be surmised from information on current forests 

and the dominant species (summarized in section 3.2).

All available evidence suggests that high-elevation, 

forested landscapes encountered by early explorers in 

southern Wyoming must have included large areas of 

even-aged cohorts that originated during crown fires that 

occurred every 100 to 200 years or more. Uneven-aged 

stands of older forest would be found in ravines and other 

places where fires were less likely to burn or in other areas 

where, by chance, there had not been a fire for several centuries (Romme and Knight 1981). The landscapes probably were patchy due to uneven burning (Foster and others 

1998; Howe and Baker 2003), as large fires would burn 

for several weeks or months, with greater intensity some 

days than others. Winds would shift from time to time, 

causing the flames to move in different directions. Also, 

spotting would occur and the flames often would move 

more rapidly uphill than downhill. New edges would be 

created between burned and unburned stands, creating 

environments favorable for the windthrow of both dead 

and live trees as well as the spread of certain insects and 

pathogens. Moreover, different parts of a burn would experience different successional trajectories, with doghair 

stands of lodgepole pine in some areas, stands of average or below-average tree density elsewhere, and various 
46 USDA Forest Service RMRS-GTR-139. 2005

combinations of aspen, Engelmann spruce, subalpine fir 

and lodgepole pine across the landscape (Stahelin 1943; 

Turner and others 1997; Reed and others 1999). The HRV 

for succession and successional processes surely would 

have been quite broad.

The only spatially explicit study of pre-EuropeanAmerican landscape patterns on the MBNF thus far 

was done by Kipfmueller and Baker (2000) on a small, 

3241 ha (8005 acre) watershed in the northern part of 

the Snowy Range unit (portions of the Rock Creek and 

Cooper Creek drainages). Using a detailed analysis of 

fire history, they reconstructed the landscape pattern as 

it existed in 1868. They identified 14 stand replacing 

fires in their study area between 1569 and 1996. Some 

patches of old forest did not burn during this time. Most 

of the fires were small, burning less than 200 ha (494 

acres). They concluded that climatic conditions played 

a dominant role in controlling the occurrence of standreplacing fires, but that fire suppression had been effective in their study area. In a companion report that used 

the same data set, Baker and Kipfmueller (2001) concluded that the natural fire regime left relatively large, 

connected patches of interior forest. Ninety percent of 

their study area burned over the time period that they 

were able to examine with the tree-ring record, and, as 

Romme and Despain (1989) found in YNP, there was 

no evidence for a shifting mosaic steady state on the 

landscape.

The initial effect of timber harvesting on landscapes, 

especially clearcutting and the requisite roads, was to 

“perforate” areas of interior forest. Although edges have 

always been created between burned and unburned areas, human-created edges around harvested units and 

along roads are more abrupt. Moreover, the effects of 

roads are permanent unless the roads are allowed to revegetate after their utilization for a management activity 

is completed. The effects of any edge on plants and animals that require interior forest conditions can be detrimental (Murcia 1995; Beauvais 1997; Baker and Dillon 

2000); and when depth-of-edge effects become widespread, a landscape is said to be fragmented. In fragmented forests, tracts of interior forest are hard to find. 

Some people restrict use of the term “fragmentation” 

to disturbances caused by timber harvesting and road 

building, primarily because the kinds of edges or transitions created by humans are different. For this report, 

we define fragmentation as an increase in the number 

of abrupt transitions between forest and non-forest or 

young-forest that reduce the area of interior forest and 

increase the area affected by edges. Any disturbance to 

the forest canopy–whether initiated by humans, fire, or 

wind—can cause fragmentation. Still, there seems little 

doubt that the forests of the MBNF are more fragmented 

now than prior to the advent of large scale timber harvesting and road building (Baker 1994; Baker and Dillon 

2000; Reed and others 1996a, 1996b; Tinker and others 

1998; R. Knight and others 2000).

Tinker and Baker (2000) examined the characteristics 

of clearcut harvesting as it was practiced from 1985 to 

1993 on the MBNF. During this short time period, over 

95% of scheduled clearcuts under the 1984 forest plan 

were in lodgepole pine. Typically, multiple clearcuts averaging about 6 ha (15 acres) were placed within the 

boundaries of large timber sale areas, with an average of 

about 550 ha (1359 acres) that were cut each year. Using 

these observed patterns, and current MBNF guidelines for clearcut placement (i.e., a minimum age of 90 

years and with no new clearcuts placed next to existing 

clearcuts or stands less than 21 years old), Tinker and 

Baker used GIS models to project the effects of this harvest regime for 100 years into the future. Results from 

this simulation showed that the forest would become increasingly more fragmented (i.e., with more edges between old and young forests). Stands >50 ha (124 acres) 

would be virtually eliminated, as would interior forest; 

and patch sizes and core area sizes would become much 

less variable.

Essentially, Tinker and Baker’s results suggest that, 

if practices observed from 1985-1993 continue, the spatially variable landscape patterns created by fire would 

cease to exist. In their place would be a patchwork of 

more or less equally sized patches with staggered age 

distributions. If some old stands are left in tact, all ageclasses would continue to be represented on the landscape, but patches of the same age-class would become 

small and isolated rather than large and contiguous. 

Tinker and Baker concluded that future landscape structure will have little resemblance to conditions in the 

1700s and 1800s if the practices observed from 1985-

1993 continue (which included the rather unusual patchcutting associated with the Coon Creek Project). Other 

studies of landscape patterns created by clearcut harvesting and road building have reached the same conclusion 

(Reed and others 1996a, 1996b; Miller and others 1996; 

Tinker and others 1998; Baker 2000).

The utilization of clearcut harvesting as the primary 

mechanism for stand-replacing disturbance, rather than 

fire, leads to several interesting comparisons (fig. 20). 

First, the amount of area disturbed each year by clearcutting (not including possible edge effects) is actually within the HRV disturbed by fire in YNP, which we believe 

is a useful proxy for the MBNF because of its similar 

vegetation. Secondly, however, fire-caused disturbances 

in most years cover only a small area (0-500 ha [0-1235 
USDA Forest Service RMRS-GTR-139. 2005 47

MBNF Clearcuts 1950-1994

0

1

2

3

4

5

6

7

8

9

0-499

500-999

1000-1499

1500-1999

2000-24992500-2999

3000-3499

3500-3999

4000-44994500-49995000-5499

5500-59996000-6499

6500-6999

7000-7499

7500-7999

8000-8499

8500-8999

9000-9499

9500-9999

10000+

Area/year (ha / million ha)

Number of years

range = 0 - 4392 ha 

mean = 2273

st. dev. = 1170 

median = 2242 

MBNF Clearcuts 1972-1988

0

1

2

3

4

5

0-499

500-999

1000-14991500-1999

2000-2499

2500-29993000-34993500-3999

4000-44994500-4999

5000-5499

5500-59996000-6499

6500-6999

7000-74997500-79998000-84998500-89999000-9499

9500-9999

10000+

Area/year (ha / million ha)

Number of years

range = 922 - 3509 

mean = 1981

st. dev. = 808 

median = 1875

Yellowstone Fires 1972-1988

0

2

4

6

8

10

12

14

0-499

1000-1499

2000-2499

3000-34994000-4499

5000-5499

6000-6499

7000-7499

8000-84999000-9499

10000+

Area/year (ha / million ha)

Number of years

range = 0 - 10,423 

mean = 1,138

st. dev. = 2,847

median = 12

Statistics for 1972-1987:

(a)

(c)

(b)

500-999

1500-19992500-29993500-3999

4500-49995500-5999

6500-6999

7500-7999

8500-8999

9500-9999

Figure 20—Frequency distributions of area disturbed per year for two different types of stand-replacing disturbance—clearcuts in the MBNF for 1950-1994 and 1972-1988, and fires in Yellowstone National Park from 

1972-1988. Data are standardized to hectares per million hectares of forested area. Statistics for YNP were 

calculated for 1972-1987. If the 321,000 ha burned in 1988 are included, the statistics (standardized) are: 

range = 0 - 401,250; mean = 24,674; standard deviation = 97,081; median = 16. Sources: MBNF data from 

von Ahlefeldt and Speas (1996); YNP data from Despain and Romme (1991).
48 USDA Forest Service RMRS-GTR-139. 2005

acres]), with very large areas (>5000 ha [12,350 acres]) 

affected in occasional years conducive to fire. The fire 

regime on the MBNF has not changed much in the last 

50 years, but fire sizes probably have been smaller than 

during the reference period (see fig. 16) because of successful suppression (Kipfmueller and Baker 2000).

In contrast to fire, clearcutting on the MBNF has tended to disturb a larger and relatively consistent amount 

of land area each year (mostly 500-3500 ha [1235-8645 

acres]), plus there is the additional landscape effect of 

the required road network for harvesting. Therefore, 

the landscape mosaic created by clearcut harvesting, as 

practiced from 1950 to 1994, is not likely to resemble 

the variable mosaic created by a “natural” fire regime. 

As noted, that has not been a management objective. 

Significantly, on the two units with most of the high-elevation forests (Snowy Range and Sierra Madre), changes attributable to harvesting will occur on only about half 

of the forest land, as that was the amount classified as 

commercial and suitable for harvesting in 1985 (MBNF 

Management Plan 1985; the current plan indicates the 

proportion of the forested land classified as suitable and 

appropriate for timber harvesting is 37%). If the effects 

of roads are considered, then the affected portion of the 

forested land is much larger (Reed and others 1996a, 

Baker and Dillon 2000).

The vegetation mosaic of Rocky Mountain landscapes 

is known to vary greatly through time, primarily because 

of large-scale fires and other natural disturbances. For 

example, Romme and Despain (1989) found in YNP 

that the amount of land area in mature forests occupied 

about 20% of their study area in the early 1700s, but 

about 40% in the late 1900s. When the land area occupied by old forests was high, the land area in young forests was low. Such shifts in the landscape mosaic surely 

have ecological effects on species composition, biodiversity, nutrient cycling, and other ecological phenomena. A similar shifting mosaic must have occurred over 

much of the high-elevation forests of the MBNF, as suggested by the results of Baker and Kipfmueller (2001).

In the following sections, we summarize our conclusions about specific landscape variables and the probability that they are within or beyond their HRV.

4.4.1. Landscape diversity (number and 

proportion of land-cover types)

All of the major vegetation types present during the 

reference period probably are still found today, though 

in different abundances and with subtle variations that 

could be caused by climate change as well as by land 

management practices. Old-growth forests (generally 

>200 years old with considerable CWD on the ground 

and a relatively large number of snags) probably cover 

less area and mid-successional forests (50-150 years old) 

cover more area, largely due to the impacts of EuropeanAmericans since the late 1800s. Humans have caused 

shifts in the abundance of different cover types and age 

classes across the landscape, but the number of different kinds of major land cover types probably has not 

changed (high confidence).

Many of the silvicultural prescriptions during the last 

50 years have been for the conversion of slow-growing 

old forests into faster-growing young forests. The plan 

in recent years has been to harvest previously harvested 

forests within a rotation time of about 90-140 years—an 

interval that is generally shorter than the natural rotations 

for the kinds of significant disturbances that occurred in 

individual, high-elevation stands prior to dominance by 

European-Americans. Spruce-fir forests may become 

less common if they are harvested to create pine stands 

that produce wood more quickly, though some remaining stands of this forest type at high elevations have 

been classified as unsuitable for harvest. Harvest rotation times of 140 years or less will stop the succession of 

pine forests to spruce-fir forests; and with fire suppression, aspen may become less common in some areas—

though silvicultural prescriptions can favor aspen as well 

as the different species of conifers. In general, we have 

the impression that, while changes have occurred, the 

proportion of high-elevation landscapes in different cover types has not yet exceeded the HRV (moderate confidence). This could occur in the future depending on the 

rate at which spruce-fir forests are harvested and the success of fire suppression. The proportion of the MBNF in 

different cover types at the present time is summarized 

in table 3 and section 3.2.

4.4.2. Proportion of landscape in different 

stand-age classes

Studies on high-elevation coniferous forests in YNP 

suggest that the proportion of land area in young, middle-aged, and old forests has varied considerably during 

the last several centuries (Romme and Despain 1989). 

Similar patterns must have occurred in the MBNF. 

However, a tendency to harvest the larger, slow-growing 

trees of old forests, and an ability to harvest such forests 

on north slopes and ravines where fires might not have 

burned as often, suggests that the amount of old forest 

today is lower on the MBNF than at any time during 

the reference period. Approximately 2% of the lodgepole pine and 9% of the spruce-fir forests have been estimated at 200 years old or older on the Snowy Range 

and Sierra Madre units combined (see figs. 19a and b). 

Further study is required to know with confidence if 
USDA Forest Service RMRS-GTR-139. 2005 49

the abundance of old forests is below the HRV for each 

forest type, but the trend is probably in that direction 

(moderate confidence). Considering that the proportion 

of the forests in different age classes fluctuates through 

time, keeping the proportion of the older forests at a low 

level would also be beyond the HRV. In YNP, Romme 

and Despain (1989) found that, since about 1740, at 

least 10% of the lodgepole pine forests of their study 

area were >300 years old (LP3 age class), and at least 

another 15% were 150-300 years old (LP2 age class). 

Such results further suggest that the MBNF land area in 

old forests could be below the HRV.

MBNF data suggest that about 25-50% of the forests on the Snowy Range and Sierra Madre units were 

burned and harvested in the late 1800s and early 1900s 

(see figs. 6 and 19), creating the large number of 100- 

to 130-yr-old stands that exist today. This is reflected 

in the land area with middle-aged forest (60-150 years 

old), which appears to be above the HRV (fig. 21; moderate confidence). Unfortunately, determining the age 

of older, uneven-aged stands is difficult, and commonly 

the results are under-estimates (Kipfmueller and Baker 

1998a). Determining if the MBNF data are adequate for 

estimating land area in different age classes during the 

reference period would require additional fieldwork and 

is beyond the scope of our study. Therefore, we cannot 

have a high level of confidence in our conclusion about 

age-class structure on the landscape. It would be useful 

to determine the ages of the forests that were harvested 

or burned during the late 1800s and early 1900s.

Figure 21—Abundance of early, middle, and late successional stages of subalpine 

forests on the Snowy Range 

and Sierra Madre units of 

the MBNF, compared to 

historical absolute ranges for the various stages in 

YNP. MBNF data represent 

lodgepole pine and spruce/

fir forests for which age 

data are available (147,640 

ha of the Snowy Range, including Sheep Mountain, 

and 66,000 ha of the Sierra 

Madre; source: MBNF). YNP 

data are from similar forests 

within the 129,600 ha study 

area of Despain and Romme 

(1991). EC = existing condition. See table 4 for definitions of LP0, LP1, LP2, and 

LP3.
50 USDA Forest Service RMRS-GTR-139. 2005

4.4.3. Edge, interior forest habitat, and 

patchiness

Because of increased rates of timber harvesting and 

road building during the last 55 years, it seems clear, 

as discussed previously, that the amount of edge has increased and the amount of interior forest has decreased 

over much of the MBNF (Baker 1994; Baker and Dillon 

2000; Reed and others 1996a, 1996b; Tinker and others 

1998; Tinker and Baker, 2000; Knight and others 2000, 

Baker and Kipfmueller 2000). Both characteristics probably have exceeded the HRV for the reference period 

(high confidence).

A considerable amount of edge during the reference 

period surely was caused, as it is today, by abrupt environmental changes (Knight and Reiners 2000) and natural disturbances (Veblen 2000), but there are differences 

in patterns created by human activities. Generally, roads 

and harvested plots have edges that are more abrupt, and 

the disturbed areas differ in shape and location from natural patterns. Roads are long and narrow, but they influence a large area in the MBNF (fig. 10), as do clearcuts 

(fig. 8), and both are significant causes of fragmented 

landscapes. Most likely, there are now more mid-size 

patches (created by clearcut harvests) that are more regularly distributed across the landscape than during our 

reference period (high confidence), when there probably 

were larger patches of burned forest due to larger fires.

4.4.4. Ratio of forest to non-forest land

Successful regeneration of new stands usually follows timber harvesting on the MBNF, though the rate 

of tree establishment and growth has been slower on 

wind-swept sites at high elevations (such as the French 

Creek clearcuts visible from the Snowy Range road). In 

most areas the stocking rate is very good, if not too high, 

from a silvicultural perspective. Our observations suggest that harvesting has not enlarged subalpine meadows, and trees have not invaded many of the meadows. 

Consequently, the ratio of forest to non-forest land (see 

table 3) is probably within the HRV on high-elevation 

landscapes (high confidence).

4.4.5. Rate at which new patches are formed

The 1985 MBNF management plan suggests that the 

silvicultural rotations for lodgepole pine, mixed conifer/spruce-fir, and aspen forests would be 90-140, 100-

180, and 80-120 years, respectively. Considering that 

the mean crown-fire interval (point scale) probably was 

150-200 years in these forest types at high elevations, 

and that disturbances other than fire would have affected 

a smaller area than harvesting, it is likely that the rotation is shorter than the HRV for new patch formation by 

20-120 years, depending on forest type, forest location, 

and the manner in which MFI is estimated (high confidence).

4.4.6 Proportion of land with high snag and 

CWD densities

Timber harvesting removes significant numbers of 

large trees, as discussed in sections 4.3.13 and 4.3.14, 

and large trees are the source for snags and CWD. 

Considering the large amount of lodgepole pine and 

Engelmann spruce that have been harvested on the 

MBNF from the late 1800s to the present, it seems probable that the proportion of land with high snag and CWD 

densities is lower than the HRV (high confidence).

4.4.7 Proportion of land with high tree density

Silvicultural prescriptions often call for lowering tree 

density to increase the growth of those that remain. The 

thinning of trees has affected a substantial area on the 

MBNF, but patch cutting and clearcutting often have 

been preferred, to minimize wind damage and provide 

more effective control of dwarf mistletoe. Forest thinning in some areas reduces the proportion of the forested land with trees of high density, possibly to the point 

where this variable is below the HRV (low confidence). 

Natural thinning certainly does occur, but optimal spacing for maximum tree growth was not characteristic of 

forest ecosystems during the reference period. Most 

thinning has occurred on sites where there is the potential for high tree density, as that is where tree growth is 

most likely to make the effort worthwhile.

4.5 Summary for high-elevation forests

In summary for high-elevation forests, human-caused 

disturbances that occurred during the initial period of 

European-American settlement in the late 1800s left 

a significant imprint on the MBNF. Still, the HRV for 

some variables probably was quite broad during the reference period (1600-1850). At the stand scale, the apparent deviations from the HRV are the number of snags, 

the amount of coarse woody debris, density and size of 

canopy gaps, age- and size-class structure in some areas, forest floor depth, and the degree to which mineral soil is affected by disturbances. At the landscape 

scale, deviations from the HRV also appear to be driven by the introduction of timber harvest as the primary 

disturbance agent, combined with road construction and 

the successful suppression of some fires. Since the late 

1800s, an estimated 70-80% of the forests have been cut 

or burned on the MBNF. Also, due to a comparatively 

large amount of accessible terrain and a relatively high 
USDA Forest Service RMRS-GTR-139. 2005 51

road density (miles of road/square mile), a high level 

of human-caused edge forest has been created. Such 

changes cause a concomitant reduction in the area of 

interior forest.

Considering that large areas of forest were disturbed 

by European-Americans in the late 1800s, the species 

that are present today, over a century later, are those that 

have been able to survive. There are insufficient data to 

identify which species, if any, have become extinct due 

to timber harvesting and road construction, but scientists 

have identified lists of sensitive species. Further research 

may identify other sensitive species as well as species 

that are less sensitive than currently thought.

5. The HRV of Low-Elevation Forests 

and Woodlands

Forests, woodlands and savannas composed of ponderosa pine, limber pine, aspen, and/or Douglas-fir are 

common throughout the Rocky Mountains at low-elevations and are often intermingled with foothill shrublands 

and grasslands. Lodgepole pine occurs in the foothills as 

well. When the trees grow densely enough to be considered forests, the stands are small and typically located in 

ravines or on north slopes. 

In the MBNF, the most common types of low-elevation forests are those dominated by lodgepole pine and 

ponderosa pine. Lodgepole pine is not usually thought 

of as a foothill species, but it occurs at the edge of the 

mountains, such as on the east side of the Snowy Range 

unit where the surrounding terrain is too high or dry for 

other tree species. The dynamics of lodgepole pine forests at low-elevations are very similar to those at high elevations, with the possible exception that fire may occur 

more frequently in the foothills. As lodgepole pine forests have been discussed previously, in this section we 

focus primarily on ponderosa pine.

Ponderosa pine in the MBNF is found primarily in 

the Laramie Mountains (figs. 22 and 5d). Generally, 

it occurs in areas that receive relatively high summer 

precipitation (280-760 mm/year) and where the growing season is relatively warm (Knight 1994a, 1999). 

Ponderosa pine seedlings appear to be more sensitive to cold temperatures than those of lodgepole pine 

(Cochran and Berntsen 1973), which may account for 

why ponderosa pine is rare in the Snowy Range and 

Sierra Madre (where it is found only in small areas on 

warm, south-facing slopes at low elevations). Ponderosa 

pine is able to form extensive stands in the Laramie 

Mountains, where it comprises 51% of the forested area 

and is most common at elevations of 1875 to 2590 m 

(6150 to 8500 ft). Ponderosa pine is less dense on the 

Sherman Mountains unit, but it makes up 47% of the 

relatively small area with trees (38% forested; see table 

3 and fig. 5c).

Douglas-fir occurs at similar elevations as ponderosa pine, though it grows more densely on mesic sites at 

somewhat higher elevations or on north-facing slopes. 

Douglas-fir can be found in small stands in the MBNF 

(fig. 23), such as on Sheep Mountain and along the 

North Platte River, but it is more common to the west 

and northwest in Wyoming, in Colorado to the south, 

and in Montana to the north. On the MBNF, Douglasfir occupies 2%, 2%, 4%, and 1% of the forested land 

on the Sierra Madre, Snowy Range, Sherman Mountains 

and Laramie Mountains units, respectively (von 

Ahlefeldt and Speas 1996). Despite its low abundance in 

the MBNF at the present time, some investigators have 

speculated that Douglas-fir was once more widespread 

in portions of the Snowy Range, Sherman Mountains, 

and Laramie Peak units (Douglas 1913, Alexander and 

others 1986, von Ahlefeldt and Speas 1996).

Limber pine may occur with ponderosa pine or 

Douglas-fir, but it is ecologically distinct in various 

ways. First, it tolerates a wide range of environmental 

conditions (fig. 24), as is evident by the fact that it occurs on escarpments at 1528 m (5042 ft) elevation near 

Pine Bluffs, Wyoming (on the western Great Plains near 

the Nebraska border) and at alpine treeline above 3000 

m (9900 ft) (McNaughton 1984; Schuster and others 

1994). Perhaps due to an inability of seedlings to compete with other species, it is often found on rocky soils 

and ridges (Lepper 1974; Alexander 1986) where other 

species are not able to survive. Also, the distribution of 

limber pine appears to be affected by the behavior of 

the Clark’s nutcracker (Tomback 1983), with the birds 

planting the seeds on exposed ridges and slopes where 

snow is less likely to accumulate. Based on the MBNF 

data, limber pine occurs on <1% of land area on the 

Snowy Range and Sierra Madre units, but 2% and 9% of 

the woodlands on the Sherman Mountains and Laramie 

Mountains units, respectively (von Ahlefeldt and Speas 

1996).

Depending on various physical and biological factors, 

ponderosa pine, Douglas-fir, and limber pine can occur 

individually or in mixed stands. Due to the underlying 

soils and geology, or variations in topography and microclimate, these species often form open woodlands or 

patchy mosaics, with groups of trees interspersed with 

shrubs, grasses and forbs. Common species associated 

with such woodlands in Wyoming include skunkbush 

sumac (Rhus trilobata), common chokecherry (Prunus 
52 USDA Forest Service RMRS-GTR-139. 2005

virginiana), mountain big sagebrush (Artemisia tridentata var. vaseyana), Oregon grape (Mahonia repens), western snowberry (Symphoricarpos occidentalis), Rocky 

Mountain juniper (Juniperus scopulorum), ground juniper (Juniperus communis), Idaho fescue (Festuca idahoensis), king spikefescue (Leucopoa kingii), sideoats 

gramma (Bouteloua curtipendula), and little bluestem 

(Andropogon scoparius) (Knight 1994a). 

Maximum ages for trees in low-elevation forests on 

the MBNF usually have been estimated at 250-350 years 

for ponderosa pine, 200-300 years for Douglas-fir, and 

500+ years for limber pine. Brown and others (2000) 

found ponderosa pines that are 500 years old in the 

Laramie Peak area. Stand ages might have been considerably younger or older than the ages of individual trees, 

but ponderosa pine and Douglas-fir have thick bark, enabling them to survive many surface fires, and often they 

Figure 22—Distribution of ponderosa pine forests and woodlands on the MBNF. Source: MBNF.

occur on sites that are not highly flammable, such as 

rocky escarpments (States 1968). Both insect epidemics and fire would have killed some of the trees (usually 

not all of them), thereby contributing to the open nature 

of the stands.

5.1. Disturbances at low elevations

Fire is thought to be the predominant factor in the 

maintenance of overall stand and landscape structure in 

both low- and high-elevation forests, but other disturbances have also played important roles. Mountain pine 

beetles, for example, have at times caused large standreplacing disturbances, such as in the Laramie Peak unit 

(1988-94). Other agents seem to be individually less important— including root rots, wind storms, and heavy 

frosts as well as other insects—but the interactive, often 
USDA Forest Service RMRS-GTR-139. 2005 53

Figure 23—Distribution of Douglas-fir woodlands on the MBNF. Source: MBNF.

synergistic relationships between the various disturbance 

agents present in low-elevation forests have enhanced the 

natural diversity (fig. 25; Lundquist 1995a, 1995b).

5.1.1. Fire

The dynamics of ponderosa pine woodlands prior to 

the time of fire suppression has been the subject of considerable debate (fig. 26) (e.g., Covington and Moore 

1994; Shinneman and Baker 1997; Veblen and others 

2000). The most common description assumes relatively frequent, low-intensity surface fires that kill most of 

the small trees but few of the older, thick-barked trees—

thus maintaining low-density ponderosa pine woodlands that only rarely would have been burned by crown 

fires. Historic accounts describing “open and park-like” 

stands with understories rich in grasses were common in the American Southwest (Cooper 1960); and 

on the eastern slope of the Laramie Mountains, many 

early explorers in the 1800s described ponderosa pine 

as “thinly scattered,” “stunted,” “scraggy,” and growing 

“most sparsely” (Dorn 1986, pp. 3-8). Photographs taken in the 1800s documented open stands of ponderosa 

pine and Douglas-fir throughout the Rocky Mountains 

(Progulske 1974; Gruell 1983b; Veblen and Lorenz 

1986, 1991; Johnson 1987; Parrish and others 1996).

Fire history studies for portions of the Black Hills 

and Front Range, both areas with stands similar to those 

on the MBNF, have generally estimated mean fire intervals (MFI) of 15-45 years (Laven and others 1980; 

Fisher and others 1987; Goldblum and Veblen 1992; 

Brown 1994; Brown and Sieg 1996, 1999), compared 

to MFIs of less than 10 years in the Southwest (Weaver 

1951; Dieterich 1980; Fule and others 1997). Ponderosa 

pine fire scars from a natural area near Laramie Peak 

(the Ashenfelder Basin) indicate a median fire interval of 26 to 34 years (Brown and others 2000; fig. 27). 
54 USDA Forest Service RMRS-GTR-139. 2005

Where Douglas-fir and ponderosa pine stands occur 

near each other in Colorado, fire intervals for stands of 

Douglas-fir appear to be slightly longer (Goldblum and 

Veblen 1992), possibly because Douglas-fir usually occurs on more mesic sites where fires may have been less 

frequent. In a thorough review of the literature for the 

Colorado Front Range, Veblen (2000) found a range in 

MFI of 5-40 years. Baker and Ehle (2001) concluded, 

however, that such ranges are too narrow and are biased 

toward shorter intervals.

A second explanation of ponderosa pine dynamics suggests that infrequent crown fires are more characteristic 

of the ponderosa pine forests than frequent surface fires, 

such as Shinneman and Baker (1997) proposed for the 

Black Hills of South Dakota and northeastern Wyoming. 

According to Shinneman and Baker, disturbances in the 

ponderosa pine forests of this area were infrequent and 

unpredictable stand-replacing fires that created large areas of dense, closed-canopy, even-aged forest similar to 

stands of lodgepole pine. They review the journals and 

photographs of early explorers such as Dodge (1876), 

Newton and Jenny (1880), and Graves (1899), which 

provide evidence that broad expanses of dense ponderosa pine forests and extensive, stand-replacing fires were 

common in the Black Hills prior to the arrival of EuropeanAmericans. Early observations on Casper Mountain, on 

the north end of the Laramie Mountains (though outside 

the MBNF boundary), also described “extensive tracts 

of Pitch Pine [ponderosa]” and “Gigantic pines in great 

density” (Dorn 1986, pp. 10-12).

In reality, both ponderosa pine savannas and dense 

forests could have coexisted at low elevations in the 

MBNF during the reference period, especially in the 

Laramie Mountains unit. Variability in site conditions 

Figure 24—Distribution of limber pine woodlands on the MBNF. Source: MBNF.
USDA Forest Service RMRS-GTR-139. 2005 55

Figure 25—Diagram illustrating 

the interactions among predis-

posing agents, killing agents, 

and tree responses observed 

at three sites in the Black Hills 

dominated by ponderosa pine. 

Dashed arrows with large ar-

rowheads indicate pathways 

enhanced by timber harvest-

ing. Dashed arrows with open 

arrowheads indicate path-

ways diminished by timber 

harvesting. Solid arrows sug-

gest pathways largely unaffect-

ed by timber harvesting. After 

Lundquist 1995a.

Figure 26—Diagram illustrat-

ing various pathways of stand 

development in ponderosa 

pine ecosystems. Rectangles 

represent vegetation states, 

ovals represent processes, 

and arrows represent chang-

es through time. Interactions 

among the components en-

closed with dashed lines rep-

resent the following models 

or trajectories described in 

the literature: (1) a predomi-

nant model for the southwest-

ern United States and thought 

to be common on relative-

ly dry sites in Colorado and 

Wyoming; (2) the trajectory of 

many ponderosa pine forests 

during this century; and (3) 

the model described for the 

Black Hills and probably ap-

plicable to more mesic sites in 

Colorado and Wyoming. See 

text for discussion and litera-

ture citations.
56 USDA Forest Service RMRS-GTR-139. 2005

throughout the montane and foothill zone likely resulted in variability in fire regimes as well (Thomas Veblen, 

personal communication). Studies in the Colorado Front 

Range have identified mixed-severity fire regimes, which 

have a combination of surface and stand-replacing fires 

over a period of time in the same area (Shinneman and 

Baker 1997; Kaufmann and others 2000; Veblen and 

others 2000; Ehle 2001; Ehle and Baker 2003).

Because there is evidence for regimes involving surface, crown, and mixed-severity fires, fire frequency 

in ponderosa pine and Douglas-fir probably would not 

have been the same in all areas. This is reflected in 

part by the significant variation in MFIs among different studies and locations within the region (table 6). 

MFI estimates vary widely (e.g., 3-161 years, Laven 

and others 1980; 8-82 years for Ashenfelder Basin, 

Brown and others 2000), partly because of differences 

in methodology (Veblen 2000) and partly due to differences in climate, fuel conditions, topographic position, 

and various uncertainties and biases (Baker and Ehle 

2001). Nevertheless, warmer and drier sites (or years) 

probably experienced more frequent fires, while fires 

may have been less frequent when or where conditions 

were more mesic. Veblen and others (2000) conducted 

a detailed study of tree rings in ponderosa pine forests 

of the Colorado Front Range, finding that the alternating pattern of wet and dry periods within a 5-year period is conducive to fire spread. Wet periods apparently 

associated with El Niño conditions allowed the production of fine fuels that increased flammability during 

dry periods associated with La Niña conditions. The effect of El Niño, however, may not be as strong in areas 

north of the Colorado Front Range, such as in Wyoming 

(Veblen 2000).

A factor complicating our ability to interpret the HRV 

for fire in low-elevation forests and woodlands is the 

role that humans have had in influencing the fire regime. 

Prior to the arrival of European-Americans, fires set by 

Native Americans probably were most common at lower 

elevations (e.g., Barrett and Arno 1982; Gruell 1983a) 

and were inseparable from “natural” fires (Martin and 

Sapsis 1991). However, in a recent review of the effects 

of Native Americans on fires in the Rocky Mountains, 

Baker (2002a) concluded that—away from heavy use areas—there is no evidence to suggest that fire regimes 

were strongly influenced by people until the arrival of 

European-Americans.

During the past century, however, human influences 

appear to have lengthened the fire return interval in many 

ponderosa pine forests throughout the Rocky Mountains 

(Covington and Moore 1994; Parrish and others 1996; 

Veblen 2000; Brown and others 2000; Keane and others 2002). On the MBNF, Brown and others (2000) estimated that fire had not occurred since 1911 and 1909 

at two sites within the Ashenfelder Basin (north of 

Laramie Peak) (fig. 27). The resultant 86- and 88-year 

intervals since the last fire are 12 and 6 years longer than 

any recorded in the 1800s, and they are over three times 

longer than the median before the arrival of EuropeanAmericans. Notably, Baker and Ehle (2003) argue that 

additional data are required before such conclusions can 

be accepted with confidence.

In many areas, it seems clear that fire suppression during the 20th century has contributed to the establishment 

Figure 27—Return intervals for surface fires in stands of ponderosa pine in the Ashenfelder Basin, north of Laramie Peak on 

the Laramie Mountains unit of the MBNF. The HRV line represents the HRV for fire return intervals prior to ca 1900; the 

box on the HRV line is an estimate of the median interval during the same period. HRV = historic range of variability; EC 

= existing condition, which is essentially the number of years since the last fire. Data from Brown and others 2000
USDA Forest Service RMRS-GTR-139. 2005 57

of dense stands of smaller trees that are now very susceptible to stand-replacing fires (Covington and Moore 

1994; Parrish and others 1996; Fule and others 1997; 

Keane and others 2002). Livestock grazing also may 

have played a role by reducing competition for young 

seedlings from grasses and forbs and reducing the 

amount of fine fuels, as observed in Utah (Madany 

and West 1983; Keane and others 2002). Moreover, 

favorable climatic conditions during the early 1900s 

(especially around 1920) could have contributed to 

higher ponderosa pine densities (Savage and others 

1996), and if adequate moisture was available, warming during the last century could promote ponderosa 

pine seedling establishment. As a result, intense standreplacing fires in ponderosa pine forests, in the West 

as a whole, may be more common presently than during the 18th and 19th centuries (Agee 1997; Keane and 

others 2002). Fire severity, in general, must be directly 

proportional to the length of the fire free interval, with 

more severe fires occurring where or when fire intervals are longer.

At Devil’s Tower National Monument in northern 

Wyoming, small fires caused by individual lightning 

strikes were common historically (Fisher and others 

1987). Such fires would typically damage only one or 

two trees before subsiding due to weather conditions or 

fuel discontinuity. Larger fires were less frequent, occurring only at times when weather and fuel conditions 

facilitated fire spread. In parts of the Black Hills, where 

fuel accumulated more rapidly in ponderosa pine forests because of moderate climatic conditions, fires were 

apparently larger and more intense, resembling those 

in lodgepole pine forests (Shinneman and Baker 1997). 

Fires as large as 20,000 to 60,000 ha. (49,400 to 148,200 

acres) have been reported in the Black Hills (Dodge 1876; 

Shinneman and Baker 1997, based on Graves 1899). 

Since 1890 most fires in the Black Hills have been comparatively small, except for the Jasper Fire in 2002 that 

burned 33,850 ha (83,580 acres). Significantly, many 

ponderosa pine woodlands and savannas in the MBNF 

are patchy, which could have limited the spread of fire 

to smaller areas than observed in the Black Hills (unless 

the trees were separated by flammable grasslands).

Since 1945 small fires have been much more common in the Laramie Mountains than more intense, widespread fires (fig. 28). From 1945-1993, 95% of all fires 

in the Laramie and Sherman Mountains units were less 

than 4 ha (9.9 acres), while only 11 fires (ca. 1.5%) were 

larger than 125 ha (309 acres) and only two (0.2%) were 

larger than 400 ha (988 acres) (von Ahlefeldt and Speas 

Table 6—Examples of fire return intervals reported for stands (point-scale) of ponderosa pine and Douglas-fir in the Rocky Mountains. 

Baker and Ehle (2001, 2003) concluded that differences in such estimates are not significant.

Fire return interval Type of fire Basis for

Source Forest type / location estimate (years) event estimates

Brown et al. (2000) Ponderosa pine in the Laramie Pre-settlement median Surface Fire-scarred trees

Mountains unit of the MBNF intervals = 26 to 33.5;

(Ashenfelder Basin) Range = 8 to 82

Fisher and others 1987 Ponderosa pine in Devils Tower 1600-1770: 27 Surface Fire-scarred trees

National Monument, Wyoming 1770-1900: 14

Pre-settlement: 19

Since 1900: 42

Houston 1973 Douglas-fir in Yellowstone Pre-settlement: 20-25 Surface Fire-scarred trees

National Park, Wyoming

Laven et al. 1980 Ponderosa pine in Roosevelt 3-161 (mean = 45.8) Surface Fire-scarred trees

National Forest, Colorado

Goldblum and Veblen 1992 Ponderosa pine/Douglas-fir 3-49 

near Boulder, Colorado Mean pre-1859: 32 Surface Fire-scarred trees

Mean 1859-1920: 8

Mean after 1920: 28

Brown 1994; Brown and Ponderosa pine in Jewel Cave 9-79 Surface Fire-scarred trees

Sieg 1996 National Monument, South

Dakota

Swetnam and Baisan 1996 Ponderosa pine in Northern Pre-settlement: 2-12 Surface Fire-scarred

Arizona

McCune 1983 Douglas-fir in Montana Pre-settlement: 60

Arno and Gruell 1986 Douglas-fir in Montana Pre-settlement: 35-40

Shinneman and Baker 1997 Ponderosa pine in Black Hills Pre-settlement: 250-300+ Crown Historical observaNational Forest, Wyoming and tions; stand ages

South Dakota
58 USDA Forest Service RMRS-GTR-139. 2005

1996). The total area burned on both units from 1945 

to 1993 was only 6,190 ha (15,289 acres), or 3% of the 

land area (von Ahlefeldt and Speas 1996). Notably, despite the relatively small extent of fires during the last 

50 years, lightning ignitions have been higher on the 

Laramie and Sherman Mountains than elsewhere on the 

MBNF (fig. 29), and two very large crown fires occurred 

in 2002 in the Laramie Peak area—the Hentzel Fire that 

burned 6016 ha (14,855 acres) and the Reese Fire that 

burned 7633 ha (18,848 acres).

An examination of MBNF data for stand ages suggests 

that most of the stands of ponderosa pine in the Laramie 

Peak area originated about 80-130 years ago (see fig. 

19d), though there are some forests that are considerably older. As with high-elevation forests, these young 

ages could be attributable to timber harvesting and burning in the late 1800s and early 1900s. However, because 

of more frequent surface fires and the patchy, multipleaged nature of many ponderosa pine stands, it is more 

difficult to determine stand age than for lodgepole pine 

stands. Such data must be interpreted cautiously, as the 

stands could be considerably older than fig. 19d indicates. Overall, due to effective fire suppression, we conclude that fire size and frequency are below the HRV and 

fire intensity is above the HRV for low-elevation forests 

(high confidence).

5.1.2. Insects

The primary insect affecting low-elevation forests on 

the MBNF, particularly those dominated by ponderosa 

19

244

525

9 9 2 0

0

100

200

300

400

500

600

A B C D E F G

Size class

Number of fires

0

500

1000

1500

2000

2500

Area burned (ha)

number of fires area burned

Figure 28—Number of fires and area burned on the Laramie Mountains and Sherman Mountains, 

1945-1993, on the Medicine Bow National Forest. Based on MBNF data summarized by von 

Ahlefeldt and Speas (1996). Size classes are: A < 0.1 ha.; B 0.1-3.9 ha.; C 4-39 ha.; D 40-122 ha.; 

E 123-409 ha.; F 410-2049 ha.; G >2049 ha.

pine, is the native mountain pine beetle (Dendroctonus 

ponderosae). The native Pandora moth, Douglas-fir beetle, and western spruce budworm may be present also, 

but their effects are relatively minor (Schmid and Mata 

1996). The mountain pine beetle survives within the 

phloem underneath the bark throughout nearly all of 

its life cycle (McCambridge and others 1979). Healthy 

trees can often defend themselves by producing resin to 

expel the beetle, but stressed or damaged trees may lack 

this ability.

The effects of beetle epidemics in ponderosa pine 

are similar to those for lodgepole pine. Epidemics often start when the beetles kill small groups of the larger 

trees (Schmid and Mata 1996). If adjacent trees are of 

old age, water stressed, or infected with diseases, the epidemic may spread across the landscape, killing groups 

of 100 or more trees (Schmid and Mata 1996). The results of such outbreaks include a reduction in the average 

tree diameter in a stand and the creation of small canopy 

gaps as well as large openings (Schmid and Mata 1996). 

As beetle-killed trees become snags and downed logs, 

they contribute to structural diversity (and therefore 

habitat diversity), at the same time providing fuel for the 

next fire (Lundquist 1995a, 1995b; Schmid and Mata 

1996). Other species not susceptible to mountain pine 

beetle, such as Douglas-fir, may become more abundant 

following an epidemic (Schmid and Mata 1996).

Historically, mountain pine beetle epidemics in the 

southern Rocky Mountains, including the MBNF, are 

thought to have been more severe and more frequent in 
USDA Forest Service RMRS-GTR-139. 2005 59

ponderosa pine than lodgepole pine, perhaps because 

of warmer winters at lower elevations. On the scale of 

landscapes, an epidemic may occur in some areas every 11 to 20 years (Schmid and Mata 1996), often coinciding with periods of dry climatic conditions (Beal 

1943). Return intervals for individual stands must have 

been longer, with epidemics occurring once every 50 

to 100 years (Johnson 1995, Schmid and Mata 1996). 

Such frequent epidemics may have contributed to maintaining an open stand structure with low tree density, as 

even the most severe epidemics probably did not kill all 

the trees. The duration of an epidemic apparently varied 

from 2 to 14 years (Schmid and Mata 1996), and the severity probably varied as well (McCambridge and others 

1979). In the Black Hills, historical reports indicate that 

most trees on a 50,000 ha (123,500 acres) tract of forest were killed by a beetle epidemic in 1895 (Shinneman 

and Baker 1997).

Ponderosa pine has likely evolved with the influence 

of the mountain pine beetle, as it has with the influence 

of fire. However, the increase in tree densities that have 

been reported can contribute to both competition and 

water stress, thereby reducing tree vigor and potentially 

shortening the time between epidemics. The most recent 

epidemic on the MBNF occurred from 1988 to 1994 and 

caused severe mortality at lower elevations near Laramie 

Peak, from north of Esterbrook to Palmer Canyon 

(Howard Pickerd, personal communication). Because of 

the apparently unusual tree densities, it is possible that 

the severity of this epidemic was outside the HRV. It is 

Figure 29—Comparison of fire starts in the four units of the MBNF, standardized to fire starts per 

million hectares per year. The Laramie Peak and Pole Mountain units have mostly low elevation forests (predominantly ponderosa pine). Note that lightning-caused fires are much more frequent in the Laramie and Sherman Mountains than elsewhere. The horizontal line indicates the 

mean and bar height indicates range. Source: MBNF fire records for 1945-1993 summarized by 

von Ahlefeldt and Speas 1996.

also possible, however, that the warm and dry conditions 

of the 1980s were the primary cause. If that is true, then 

this most recent epidemic may have been a natural event, 

well within the HRV (low confidence).

Overall, Schmid and Mata (1996) characterize insect populations and their effects in the Southern Rocky 

Mountains (including the MBNF) as potentially within 

their HRV, but with some important caveats. Specifically, 

they stress that measurements of “epidemics” are imprecise, which makes comparison of past and present epidemics difficult. It is clear, however, that the insects 

affecting Rocky Mountain forests are native and have 

most likely experienced periodic population increases to 

epidemic levels for thousands of years. Also, while human influences such as silvicultural treatments and fire 

suppression have the potential to move epidemic frequency and severity either below or above historic levels, insects will always be part of low-elevation forests 

and woodlands in the Rocky Mountains (Schmid and 

Mata 1996).

5.1.3. Disease

The most common diseases in low-elevation woodlands are Armillaria root disease (Armillaria mellea), 

red rot (Dichomitus squalens), Schweinitzii root rot 

(Phaeolus schweinitzii) and limb rust (Peridermium filamentosum) (Alexander 1986; Lundquist 1995a, 1995b). 

While such diseases can cause localized mortality and 

tree damage, none of them, individually, has had a 

great impact on the forest as a whole. Rather, it is the 
60 USDA Forest Service RMRS-GTR-139. 2005

combined effects of diseases and other disturbance 

agents that influence stand structure and composition 

(see fig. 25). In some cases, timber harvest and bark beetle epidemics may enhance the effects of Armillaria root 

disease and some other parasites (Lundquist 1995a), but 

these agents have not been reported as serious problems 

on the MBNF. Also, dwarf mistletoe (Arceuthobium 

vaginatum) is not common on ponderosa pine in the 

MBNF, though it is common to the south in Colorado. 

The rarity of dwarf mistletoe on the ponderosa pine of 

Wyoming may be due to geographic isolation or factors 

that either make the trees more resistant to mistletoe infection or provide an environment less suitable for mistletoe (Hawksworth and Wiens 1972).

With one known exception, the diseases found on 

trees at low elevations are native, are not thought to have 

been greatly affected by human influences, and are within the HRV (moderate confidence). The exception is the 

white pine blister rust (Cronartium ribicola), which infects five-needle pines such as limber pine. This exotic disease appears to be migrating southward from the 

Northern Rockies, where it is a serious problem for white 

bark pine (Pinus albicaulis). The rust has been found on 

limber pine in the Sherman Mountains unit and in northern Colorado (P. Krueger, personal communication). It 

probably occurs in the Laramie Peak area as well. The 

presence of an exotic pathogen could push some ecosystem variables beyond their HRV if its effects and periodicity are different than native pathogens. Unfortunately, 

such differences are not usually known at the time of invasion, and after that, the elimination of the pathogen is 

exceedingly difficult, if it’s possible at all.

5.1.4. Wind

Depending on the presence of predisposing factors, 

such as disease, insects, fire, and a shallow root system, 

the effects of wind will vary from localized tree damage 

to blowdowns over large areas (see fig. 25; Alexander 

1986; Lundquist 1995a; Shinneman and Baker 1997). 

Limber pine often occurs on very windy sites, but its 

“limber” branches are surely an adaptation to reduce 

breakage. In the Black Hills, Graves (1899) found evidence for one severe wind storm in the 19th century. Another occurred in 1940 (Shinneman and Baker 

1997), which leveled 1.5 million board feet of ponderosa pine. Despite the severe effects of such events 

(Lindemann and Baker 2001, 2002; Baker and others 

2002), wind is a natural phenomenon that will always 

be a potential disturbance agent in low-elevation forests 

and woodlands. Blowdown events probably are within 

the HRV (high confidence).

5.1.5. Timber harvest

During the 1860s, the low-elevation forests of the 

Sherman Mountains east of Laramie were harvested extensively for railroad ties, construction materials and 

fuelwood. In the Laramie Mountains unit to the north, 

many trees were harvested for telegraph poles (Grasso 

and others 1981; Boglioli and others 2002). More recent 

harvesting at low elevations has been limited primarily to ponderosa pine, especially since about 1980 when 

salvage logging following mountain pine beetle damage 

was done in the Albany Peak, Black Mountain, Laramie 

Peak, and Friend Park areas (von Ahlefeldt and Speas 

1996). The primary method employed has been thinning 

(i.e., selective harvest or “beetle harvests”), which was 

intended to slow the spread of the beetles while providing a supply of wood. Because of the limited extent of 

such harvests, the land area harvested per year has not 

been calculated and specific comparisons to other disturbance regimes are not now possible.

In general, one of the primary effects of past harvesting in ponderosa pine has been the loss of large, 

old trees—sometimes thought of as “overmature” trees. 

Both selective harvest methods and salvage logging have 

tended to remove the old trees that were important components of ponderosa pine landscapes (Covington and 

Moore 1994; Shinneman and Baker 1997). Such trees 

were a valuable source of wood, but as noted for highelevation landscapes, they also provided important elements of wildlife habitat, both as live trees and as snags 

or downed woody material (Cunningham and others 

1980; Hutto and others 1992; Crompton 1994; Hutto 

1995; Dykstra 1996; Harris 1999). At the same time, the 

decaying wood provided a mechanism for nutrient storage and a potentially valuable source of organic compounds for the soil.

Another effect of harvesting has been to reduce the 

natural variability in stand structure and age distribution 

caused by historic disturbances. In the Front Range of 

Colorado, harvesting done for the purpose of improving 

tree growth also produces a greater uniformity in density 

and age structure (Myers 1974; Alexander 1986). The effects of such an approach can be seen on the Black Hills 

National Forest, where harvesting by industry and firewood cutters has left fewer snags and less downed wood 

than in unmanaged stands (Lundquist 1995a, 1995b).

Although past timber harvesting methods may have 

created conditions that are outside the HRV, new approaches for harvesting can be a means to restore some 

stand characteristics (Covington and Moore 1994; 

Romme and others 2000). Where open stands of large 
USDA Forest Service RMRS-GTR-139. 2005 61

ponderosa pine were common prior to the arrival of 

European-Americans, the thinning of young trees that 

have invaded in recent decades, combined with the retention of larger trees and the application of prescribed 

fire, has become a common prescription (Harrington 

1981; Romme and others 2000; Kaufmann and others 

2000; Arno and Allison-Bunnell 2002).

5.1.6. Livestock grazing

Domestic livestock were first introduced in southeast Wyoming in the late-1800s, but the animals were 

not regulated on federal lands until the establishment 

of Forest Reserves around 1900 (Boglioli and others 

2002). Although records are not available before 1925 

for the Sherman Mountains unit, and before 1935 for the 

Laramie Mountains unit, grazing intensity in these areas likely peaked during the early 1900s (von Ahlefeldt 

and Speas 1996; Robert Mountain, personal communication). For the MBNF as a whole, early reports of 

the range condition in the Snowy Range and Sierra 

Madre Mountains indicated overuse and severe erosion 

(Blackhall 1909; Mullison 1909). A thorough review of 

the effects of grazing is beyond the scope of this report, 

but the number of livestock on the MBNF has declined 

steadily in recent decades (see section 3.3). The composition of herbaceous species in ponderosa pine savannas 

of the early to middle 1800s is difficult to determine. 

Grazing and fire suppression have been widespread in 

ponderosa pine savannas, and unaffected reference areas 

are rare. Likewise, although some historic descriptions 

of range conditions exist, their usefulness is limited. For 

example, Douglas (1912) provided detailed information 

on the various “grazing types” found on the MBNF, but 

his observations were made after some of the most intensive livestock grazing had already occurred. Johnson 

(1987) compared historic and modern photographs of 

rangeland in the lowlands of southeast and south-central Wyoming, revealing that very little visual change 

has occurred in many areas. However, changes in grass 

and forb species are difficult to detect from photographs. 

Moreover, some of the photographs themselves, taken 

during the 1870s, may have been taken after a period of 

heavy grazing.

Regionally, some effects of domestic livestock grazing in ponderosa pine ecosystems include the following: (1) The creation of patches caused by feeding, 

trampling, and nutrient input (urine and feces); (2) increases in soil erosion where more than 50% of herbaceous cover is removed (Currie 1975); (3) declines in 

some native species of bunchgrass that are sensitive to 

grazing (e.g., Johnson 1956; Madany and West 1983); 

(4) increases in the abundance of grazing-resistant species; (5) the introduction of exotic species such as crested wheatgrass and smooth brome; and (6) increases in 

tree and shrub density (Madany and West 1983).

Aside from direct effects on vegetation, any discussion of grazing in the context of HRV should consider 

the populations and behaviors of native herbivores before and after the changes brought about by EuropeanAmericans. In the ponderosa pine forests and woodlands 

of the MBNF, at low elevations, we know from historical 

accounts that native herbivores such as bison, elk, deer, 

and pronghorn antelope were common in the early 1800s 

(Dorn 1986). It is reasonable to assume that native grasses and forbs should be more resistant to grazing effects 

in such areas than in areas where native herbivores were 

less common (e.g., the Great Basin deserts to the west; 

Mack and Thompson 1982). However, because population estimates of native herbivores are not available for 

the reference period, direct comparisons to modern levels are impossible. Significantly, domestic and native 

herbivores do have different foraging habits. For example, bison typically roam widely, thus distributing their 

effects over large areas, while cattle often stay close to 

water, concentrating their effects in a smaller area (van 

Vuren 1982).

Overall, we don’t know if current grazing in ponderosa pine ecosystems on the MBNF is within the HRV. 

Of course, the presence of domesticated, non-native herbivores is inherently outside the HRV, but where plants 

have evolved with grazing by large native herbivores, 

the unusual effects of livestock could be minor if their 

numbers are not excessive. In 1994, grazing on the entire MBNF was reported at 46,619 AUMs for sheep and 

35,223 AUMs for cattle (von Ahlefeldt and Speas 1996). 

Such numbers are much lower than 80 or 100 years ago, 

when the heaviest grazing by domestic animals occurred 

(see section 3.3).

5.2 Stand structure at low-elevations

Based on written and photographic records from 

the 1800s, as well as reconstructions from recent studies, qualitative comparisons can be made to the conditions that have developed after more than a century of 

European-American influence. In this section we review 

the available information and draw conclusions pertinent to the HRV. In addition to timber harvesting, the effects of fire suppression, livestock grazing, and climate 

change are considered.
62 USDA Forest Service RMRS-GTR-139. 2005

5.2.1. Tree density, sapling density, canopy 

cover, age-class structure, and size-class 

structure

Qualitative information suggests that low-density ponderosa pine savannas occurred in Wyoming and 

Colorado during the 1800s (e.g., Progulske 1974; Dorn 

1986; Veblen and Lorenz 1991). Moreover, reconstructions from fire scars and other sources indicate that some 

stands in the Southwest had been maintained for hundreds of years by frequent surface fires (White 1985; 

Covington and Moore 1994; and others). The same probably occurred at low elevations in the MBNF. However, 

“dense” stands were found in the Black Hills in the 

1800s (Dodge 1876; Newton and Jenny 1880; Graves 

1899), probably on more mesic sites, and such stands 

probably were characterized by infrequent, stand-replacing fires (Shinneman and Baker 1997). The pre-1900 

canopy cover might have been 50 to 80% in such areas 

(Graves 1899). Working in the Black Hills, McAdams 

(1995) estimated that pre-European tree densities ranged 

widely, from 15 to 1600 trees/ha (37 to 3952 trees/acre). 

Similarly, Ehle and Baker (2003) reported ponderosa 

pine densities of 70 to 3000 trees/ha (173-7410 trees/

acre) for Rocky Mountain National Park to the south of 

the MBNF. Notably, their research in Rocky Mountain 

National Park suggests that the general trend in tree density with stand age may be downward rather than upward, due to mortality caused by factors other than fire 

(Ehle 2001; Ehle and Baker 2003). This would occur by 

natural thinning in areas where a dense, even-aged stand 

of ponderosa pine develops after a stand-replacing fire—

similar to those sometimes formed by lodgepole pine at 

higher elevations.

Where frequent surface fires occurred historically, 

large, older trees >14 inches (35 cm) dbh typically survived and dominated most stands. In contrast, many 

ponderosa pine forests today are dominated by young 

saplings and mid-sized trees (Covington and Moore 

1994; Keane and others 2002). This is the case for many 

ponderosa pine stands on the MBNF (von Ahlefeldt and 

Speas 1996). We found no evidence, however, to suggest that shifts in tree species composition have occurred on the MBNF, such as in the Southwest where 

Fule and others (1997) found that trees less adapted to 

fire than ponderosa pine have become more important 

in many stands (e.g., Douglas-fir and Gambel oak). One 

result of increased conifer density is increased leaf area 

index (LAI, the amount of leaf area per unit of ground 

area), which can lead to higher evapotranspiration and 

a reduction in streamflow (Orr 1975; Covington and 

Moore 1994; Keane and others 2002). Effects of this 

nature might occur in the Laramie Mountains unit, but 

the proportion of the landscape with this condition is 

not known.

In general, it seems reasonable to conclude that on relatively dry sites at low elevations where fire suppression 

has been effective, the high tree density, sapling density, 

and canopy cover observed today are outside their respective range of means during the HRV reference period (moderate confidence). The same is not necessarily 

true for more mesic sites, where high density and cover 

may be within the HRV. Also, in stands where harvest 

has occurred or fire suppression has been effective, the 

altered age-class structure and size-class structure today 

are likely outside the HRV (moderate confidence).

5.2.2. Spatial distribution of trees

An important pre-1900 characteristic of ponderosa 

pine forests appears to be the clumped distribution of 

tree cohorts in some areas at low elevations. Historical 

observations and reconstructions of open-grown ponderosa pine in the Southwest indicate that trees grew in 

clumps or patches of less than 0.4 ha (1 acre) and three 

to 44 trees (Cooper 1960; White 1985; Moore and others 1999). Covington and Moore (1994) found that age 

structure within and between these clumps varied depending upon scale. Small, even-aged groups of two to 

five trees occurred in a stand that was uneven-aged as a 

whole (White 1985). Such patterns could occur in the 

Laramie Peak area, though Ehle and Baker (2003) found 

that dense, even-aged stands were common in Rocky 

Mountain National Park. Further research is necessary 

to determine if ponderosa pine distribution patterns observed in the Southwest are also found to the north.

5.2.3. Snags and coarse woody debris

Working in the ponderosa pine forests of the Black 

Hills, which has a climate similar to the Laramie Peak 

area, Lundquist (1995a) identified 10 classes of snags 

and five classes of logs (which are part of the coarse 

woody debris) on the forest floor, finding that different 

types of disturbance create different types of snags and 

logs. For example, the stems of beetle-killed trees tend to 

break, leaving a relatively short snag, while trees killed 

by root rot are commonly uprooted and become CWD on 

the forest floor. Timber harvest, in the small area where 

it occurs at low elevations on the MBNF, tends to (1) 

reduce the number and diversity of snag types and (2) 

distribute log types among the five classes more evenly 

on the forest floor (Lundquist 1995a, 1995b). This evidence, combined with old photographs in Wyoming that 

show a much larger number of dead trees in low-elevation forests than occurs today, suggests that snags and 

downed wood are now lower than the HRV (moderate 
USDA Forest Service RMRS-GTR-139. 2005 63

confidence) where harvesting has occurred (including 

casual firewood cutting). Notably, fires in low-elevation 

forests burn some CWD, possibly more than observed 

by Tinker (1999) in high-elevation forests, because of 

frequent surface fires. However, with less frequent fires 

this wood could persist for a longer time, and if CWD inputs remain constant, then CWD may increase. We think 

that fire suppression has not been sufficiently successful to raise the amount of CWD to above the HRV (low 

confidence). This opinion is based partly on the considerable longevity of the trees that dominate the low-elevation landscapes (ponderosa pine, Douglas-fir, and 

limber pine) and the fact that pathogenic problems are 

still rather localized where these species occur.

5.2.4. Forest floor depth

Surface fires in low-elevation forests would undoubtedly burn much of the forest floor, and because they appear to be more frequent than stand-replacing crown 

fires, it is plausible that the forest floor (litter and duff) 

in such forests did not become as deep as in high-elevation forests during the reference period (1600-1850). 

Fire suppression probably is causing the development of 

deeper forest floors than normally would be expected, 

pushing this variable beyond its range of means during 

the HRV period (high confidence). A deeper forest floor 

in such forests probably contributes to the flammability 

of the forest, as does the increased tree density and more 

continuous fuel loads, which increase the potential for 

hotter surface fires that may kill more of the larger trees 

than previously.

5.2.5. Understory plant cover, species 

composition, and diversity

Where trees and saplings have become more dense, 

tree canopy cover increases and there is a concomitant reduction in the amount of understory cover (Moir 

1966; McPherson 1992). Also, shade-tolerant species 

become more abundant as canopy cover increases. In areas where tree density has increased or where harvesting has not thinned the trees, it seems probable that the 

percent cover of understory vegetation has been pushed 

below the HRV (low confidence). In some areas, exotic plants such as cheatgrass have invaded (such as near 

Woods Landing), but currently this is not thought to be 

a significant problem on the MBNF as a whole (Robert 

Mountain, personal communication). No data are available to suggest that plant species diversity is beyond the 

HRV, whether for trees or understory plants. Considering 

that the low elevation forests may have been quite dense 

at times, we think that species diversity probably is within the HRV (low confidence).

In summary, a combination of human influences appear to have caused several structural characteristics of 

ponderosa pine stands to exceed their HRV. Great attention has been given to this situation, often in the context 

of discussing “forest health.” Low-elevation forests on 

the MBNF cover a small proportion of the terrain and 

provide a rather small amount of the wood that is harvested. However, they are important for wildlife habitat 

and recreation and they occupy a large proportion of the 

Sherman and Laramie Mountains units.

5.3. Landscape structure at low-elevations

Very little information is available regarding the landscape structure of low-elevation forests and woodlands 

on the MBNF, both historically and at present. Romme 

and others (2000) stress that information about landscape patterns in ponderosa pine forests is very limited 

throughout the Southern Rocky Mountains (which includes the MBNF).

Clearly, however, changes in landscape pattern are 

occurring in low-elevation forests dominated by ponderosa pine, especially in some parts of the Laramie Peak 

area. Due to fire suppression, logging, livestock grazing, and possibly climate change, not only are some of 

the forests becoming more dense and probably beyond 

the HRV (moderate confidence), but ponderosa pine is 

invading some patches of grassland as well. Thus, forests and woodlands probably are occupying a larger area 

than previously in such habitats, while grasslands and 

savannas now occur on less of the landscape. Research 

by Shinneman and Baker (1997) in the Black Hills suggests that this trend cannot be generalized wherever ponderosa pine occurs in the region. It should also be noted 

that ponderosa pine is commonly restricted to rocky outcrops on the fringes of the MBNF, where fires may be 

less frequent (States 1968; Knight 1994a). When trees 

are clumped for edaphic reasons, they are not likely to 

become established in nearby grasslands.

Another factor contributing to landscape change is the 

harvesting of large trees, which promotes the growth of 

saplings. As a result, there probably are now more small 

trees than before in some areas, and, as small trees invade, tree cover probably has become greater and more 

uniform in parts of the Laramie Mountains unit. The net 

effect of fire suppression and harvesting has been a landscape that is less patchy and more homogeneous, with 

the result that snag density, CWD volume, and proportion of landscape in different cover types are beyond 

their range of means during the HRV period (moderate 

confidence).

This situation probably would pertain to Douglas-fir 

forests and woodlands as well, though this species is not 
64 USDA Forest Service RMRS-GTR-139. 2005

common on the MBNF. Limber pine is not usually harvested for wood and fire in limber pine communities is 

thought to be less frequent. Thus far, there is no evidence 

that limber pine distribution and density have changed 

because of human activity, though the invasion of the 

white pine blister rust is a concern in the Laramie Peak 

and Sherman Mountains units. The extensive outbreak 

of the mountain pine beetle on ponderosa pine from 

1988 to 1994 could be within the HRV, as Shinneman 

and Baker (1997) found historical records for a beetle 

outbreak over much of the Black Hills in 1895 (with 

90% of the trees killed).

5.4. Summary for low-elevation forests and 

woodlands

Overall, maintaining low-elevation forests within 

their HRV will require attention to the diversity of disturbance agents and processes with which these ecosystems appear to have developed. Under management 

strategies that include fire suppression, livestock grazing, and the removal of large trees and snags, conditions 

will quickly depart from the HRV if they have not already. On the other hand, strategies that recognize the 

importance of fire, insects, and other natural disturbances, and aim to maintain large trees and snags, will 

most likely allow low-elevation forests to be restored to 

their HRV. Romme and others (2000) suggest that dense 

stands of ponderosa pine are not necessarily bad, as they 

undoubtedly have always occurred on sites where, for 

whatever reason, surface fires may have been excluded for perhaps 40-50 years. They note, however, that 

such stands were most likely interspersed on the landscape with grasslands and savannas where surface fires 

occurred more frequently. The use of surface fires as a 

management tool has become increasingly difficult as 

more people construct cabins and homes at low elevations. It seems clear that widespread fire suppression at 

both high and low elevations has the potential of affecting a larger area than timber harvesting.

6. The HRV of Aspen Forests and 

Woodlands

Aspen usually is found in small, clonal groves 

throughout the MBNF, primarily at middle and low elevations (fig. 30). An exception to this pattern is in the 

Sierra Madre unit, where aspen dominates 20% of the 

forested area and occurs in very large stands west of the 

Continental Divide. By comparison, aspen dominates 

16% of the forests on the Sherman Mountains unit, 

4% on the Snowy Range unit, and 1% on the Laramie 

Mountains unit (von Ahlefeldt and Speas 1996)—always 

in small groves. Usually aspen is found from 2130-2590 

m (7000 to 8500 ft) in ravines or other concave surfaces 

where the soil is fine textured and moisture is available 

for a longer time during the summer. Interestingly, aspen is the only broadleaf, deciduous tree that occurs on 

the uplands of the MBNF. Though deciduous, aspen is 

evergreen in the sense that its bark has chlorophyll and 

is capable of contributing to photosynthesis in the spring 

before leaves develop and in the autumn after the leaves 

have fallen.

Aspen stands typically have relatively open canopies, 

allowing light to reach the forest floor. Lush herbaceous 

understories develop, primarily because of the light and 

the more abundant moisture and nutrient availability on 

sites where aspen becomes established. Alexander and 

others (1986) described three aspen habitat types on the 

MBNF with different understory characteristics—the 

aspen/Fendler’s meadow rue (Thalictrum fendleri) and 

aspen/pine reedgrass (Calamagrostis rubescens) habitat 

types on moist sites with relatively deep soils, primarily 

in the Sierra Madre and Snowy Range units, and the aspen/elk sedge (Carex geyeri) habitat type on somewhat 

drier sites throughout the MBNF (though most common 

in the Sierra Madre unit). 

The extensive stands of large aspen on the west slopes 

of the Sierra Madre occur where winter and summer precipitation is apparently higher because of the Arizona 

monsoon (Knight 1994a). Aspen dominates approximately 20,308 ha (50,161 acres) in the Sierra Madre, 

the largest area of contiguous aspen forest in Wyoming. 

The present stands appear to have originated following 

extensive fires in the late 1800s near Battle Mountain, 

though the forests may have existed before the fires as 

well. A large proportion of the aspen stands in all four 

units of the MBNF are approximately 100 years old (see 

19c), because of extensive disturbances during the late 

1800s and early 1900s. Alternatively, the abundance of 

100-year-old stands may be attributable to the fact that 

aspen trees (the shoots or ramets of the aspen plant) usually live only about 100 years and there may not have 

been a stand replacing disturbance in the last century.

Aspen reproduction typically is asexual, with new 

shoots produced from root sprouts—a process sometimes 

referred to as suckering (Barnes 1966; Bartos and others 

1991). This, combined with the way aspen is able to persist in the understory of some mature forests, explains 

why aspen tends to develop where it occurred previously. Sexual reproduction is quite rare, though seedlings 

do occur when severe disturbances are followed by the 
USDA Forest Service RMRS-GTR-139. 2005 65

extended, moist conditions required for seedling establishment (McDonough 1985; Romme and others 1997). 

Because of this, sexual reproduction is thought to be 

rare and episodic (DeByle and Winokur 1985; Romme 

and others 1997). There is considerable genetic diversity between clones, with some clones adapted for higher 

elevations and some responding differently than others to weather conditions (Jelinkski and Cheliak 1992). 

For example, it is common to see two adjacent aspen 

stands (clones) in the fall, one with yellow leaves and 

the other with green leaves. The clones are thought to 

be very long-lived, perhaps thousands of years (Veblen 

and Lorenz 1991). However, the oldest trees (ramets) 

in climax aspen communities on the MBNF were about 

180 years old in the mid-1980s (see fig. 19c). Some aspen stands on the MBNF appear to be self-perpetuating 

(Alexander and others 1986); others have an understory 

of conifer saplings—especially subalpine fir and lodgepole pine.

Aspen forests and woodlands support a diversity of 

birds and small mammals, as good forage and ground 

cover is provided. Loose and Anderson (1995) found 

woodpecker nests more commonly in aspen groves than 

in other forest types, and Struempf (1999) found higher 

nest survival in aspen than in conifer forests. Large aspen >18 cm dbh (average 26.7 cm; range 18-45 cm) with 

heart rot or dead tops are frequently occupied by nesting woodpeckers (Loose and Anderson 1995). The largest aspen recorded by Severson (1963) on the MBNF 

was 66 cm dbh. The mean dbh of aspen in Loose and 

Anderson’s study area in the Snowy Range was 11.8 

cm. The plant species composition of most aspen forests is much different from coniferous forests, and consequently the animal species composition is different as 

Figure 30—Distribution of aspen forests and woodlands on the MBNF. Source: MBNF.
66 USDA Forest Service RMRS-GTR-139. 2005

well. While aspen does not occupy a large area on any 

of the units of the MBNF except the Sierra Madre, this 

species contributes significantly to the overall biological 

and aesthetic diversity of the landscape.

6.1. Disturbance history, human influences 

and succession

Aspen can be seral or climax depending on site conditions and historical factors (Mueggler 1985), but typically it is thought of as a pioneer species (Bartos and 

Campbell 1998). Usually a few aspen persist in the understory of coniferous forests and produce an abundance 

of new sprouts when the next stand-replacing disturbance occurs. Notably, while young aspen shoots within a clone share the same root system with the mature 

canopy dominants, few are able to survive in the understory.

Extensive fires during the settlement period appear 

to have created large stands of aspen, both in Colorado 

(Veblen and Lorenz 1986) and on the west slope of the 

Sierra Madre. It is unknown if these large stands developed largely from sprouts or seedlings. It is also unclear 

whether aspen or conifers were dominant prior to the 

last fires (Severson 1963). With attempts at fire suppression in recent decades, along with natural succession, 

conifers are now becoming co-dominants with aspen 

in some areas, causing some managers to recommend 

that the conifers be harvested or that whole stands be 

burned to restore the aspen grove. In other areas, stands 

of aspen are not being invaded by conifers and appear to 

be relatively stable communities (Alexander and others 

1986; Mueggler 1985; Shepperd and others 2001). Thus 

far we have found no evidence that the amount of land 

area dominated by aspen on the MBNF is either lower or 

higher than the HRV (moderate confidence).

As with other forest types, various kinds of disturbances play a role in aspen dynamics, including windthrow, snow breakage, insect epidemics, disease, fire, 

timber harvesting, and grazing and browsing by domestic and native ungulates. Aspen has relatively weak stems 

and brittle branches, and is therefore easily damaged by 

heavy snow accumulation, avalanches, and wind (Veblen 

and Lorenz 1991). Various insects defoliate aspen trees, 

but two of special concern are the forest tent caterpillar (Malacosoma disstria) and the large aspen tortrix 

(Choristoneura conflicta) (Allen and Harris 1999). Both 

can affect large areas over longer periods than most other insects. After the overwintering eggs hatch, the larvae of the tent caterpillar feed on the growing leaves. 

When population sizes are high, this caterpillar quickly 

defoliates entire trees and stands—usually above 2424 

m (8000 ft). The trees may produce new leaves, but infestation for several years in a row is lethal. When this 

happens, new aspen groves develop by sprouting or the 

conifer saplings in the understory grow into the canopy 

more rapidly. Epidemics occur every 10 to 20 years in 

Northern New Mexico and Southern Colorado, with a 

single epidemic lasting 10 years or longer (Hinds 1976). 

Unfortunately, we have not found comparable information for the MBNF.

The tortrix similarly has larvae that feed at the time of 

leaf expansion, but this insect also feeds on buds and can 

reach epidemic population sizes for two to three years 

(Allen and Harris 1999). No epidemics of these insects 

have been reported for the MBNF.

In general, there is no basis to conclude that the characteristics of insect outbreaks in aspen stands are beyond 

their HRV (moderate confidence).

With regard to diseases, aspen are affected by a variety of fungi and viruses, including leaf blights, stem and 

root rots, and bark cankers (Allen and Harris 1999). The 

known diseases are native and tend to spread when the 

thin bark of aspen is damaged by animals, falling trees, 

humans, or other agents (Hinds 1976). Some diseases 

appear to increase with wounding associated with timber harvesting and campers (Hinds 1976; Walters and 

others 1982), but conversion of older stands to younger 

stands may have reduced the occurrence of disease at a 

regional scale. There are numerous mechanisms whereby aspen diseases can be spread, both natural and human, and consequently it is still impossible to determine 

if they are within the HRV.

Mean fire return intervals for aspen groves probably 

are essentially the same as for the other forest types with 

which they occur. Veblen and others (1994) found fire 

return intervals in aspen forests of 160 and 240 years 

(mean = 202) in Colorado, which is similar to some 

nearby coniferous forests. While aspen is not considered 

to form a highly flammable forest, it often burns when 

the adjacent coniferous forests burn (DeByle and others 

1987, 1989). Lightning-caused ignitions in aspen stands 

are probably rare, but fire has definitely played an important role in the establishment of new stands (Brown 

and DeByle 1989; Romme and others 1997). If fires are 

of high intensity, some aspen root systems may be killed, 

thus favoring development of conifer stands (Parker and 

Parker 1983). However, the relatively mesophytic herbaceous understory in many aspen stands probably results 

more often in cooler fires than in adjacent stands of conifers. Such fires favor the development of large numbers 

of aspen root sprouts even though the aboveground part 

of the older trees (shoots) is killed (Veblen and Lorenz 

1991).
USDA Forest Service RMRS-GTR-139. 2005 67

As with coniferous forests, modern fire suppression 

combined with fine fuel reduction by livestock grazing 

probably has contributed to the lengthening of fire return 

intervals in some aspen forests. Moreover, fire suppression has almost certainly reduced the amount of aspen 

forest in some areas in Utah because of succession to conifers (Bartos and Campbell 1998; Bartos 2001). When 

large fires burn in the future, whether because of weather, 

fuel conditions, or changes in fire management policies, 

aspen regeneration and seedling establishment probably 

will occur (Romme and others 1997). In the absence of 

overbrowsing of aspen suckers by elk and/or livestock, 

such regeneration would likely lead to increases in the 

amount of aspen forest on the landscape, even if only 

temporarily (i.e., for at least 100 years in areas where conifers are likely to invade again through succession).

Fluctuations in aspen cover in some areas can be attributed to heavy browsing by large populations of ungulates, especially elk (DeByle 1979; Kay 1990, 1993; Kay 

and Wagner 1996; Baker and others 1997; Hessl 2002) 

and secondarily domestic livestock. This effect has been 

observed primarily in the vicinity of National Parks where 

ungulate populations have been high in recent years, 

such as in YNP and Rocky Mountain National Park. 

Some observations indicate that the effects of browsing 

and trampling on aspen regeneration are common in the 

MBNF, particularly in the Laramie Mountains unit (von 

Ahlefeldt and Speas 1996, page 6-71).

Very little harvesting of aspen has occurred in the 

MBNF, though there are discussions about creating 

young stands of aspen through clearcutting as well as 

with prescribed burns. Only about 283 ha (700 acres) 

of the aspen on the west slope of the Sierra Madre had 

been harvested by the year 2000 (<2% of the aspen forest in that area).

In summary, because of very little timber harvesting, the variables pertinent to the structure of modern 

day aspen stands are probably within their HRV on the 

MBNF (moderate confidence). At the landscape scale, 

there may be a smaller area in aspen at the present 

time, due to succession to conifers, but this too has undoubtedly fluctuated during the last several centuries. 

Thus, it appears that management guidelines for aspen 

on the MBNF can be driven more by what is desired 

than by concerns about conditions that have exceeded 

the HRV. The primary human effects on aspen groves 

have been fire suppression on the landscape as a whole, 

the concentration of large numbers of ungulates on winter ranges in the foothills, and the introduction of exotic plants. While possibly different than during much of 

the reference period, when fires may have burned over 

larger areas than they do today, and when native ungulates may have been abundant at times, the disturbance 

history and structure of aspen stands probably is within 

the HRV (moderate confidence).

In fact, fire suppression may not have had much effect, considering that large fires appear to be climatically driven phenomena (Bessie and Johnson 1995; Weir 

and others 1995; Kipfmueller and Baker 2000) which 

humans typically cannot control. Big game populations 

probably fluctuated as well (DeByle 1979), with high 

populations occurring after a series of mild winters, or 

when predation by wolves, mountain lions, and bears 

was low. Information pertinent to the debate about possible causes of changes in the abundance of aspen has 

been summarized in Shepperd and others (2001) and 

Hessl (2002). It seems clear that a trend and its causes 

in one portion of the range of this widespread species 

are not necessarily the same in other areas. Much has 

been learned since the review of aspen ecology edited by 

DeByle and Winokur (1985), but much is still unknown 

about the long-term dynamics of this species in specific 

areas such as the MBNF.

7. The HRV of Non-Forest Vegetation

Though not as abundant as forests, different kinds of 

grasslands, shrublands and forblands are found throughout the MBNF (table 3; distribution maps available from 

the MBNF). Just as in forests, changes in structure and 

composition may occur through time as a result of both 

natural and human factors. The non-forest vegetation 

types are treated in less detail than the forests, primarily because there is less information suitable for an HRV 

interpretation.

Grasslands (including grass-dominated meadows) are 

found on the MBNF from the foothills to the upper limits of alpine tundra. Bluebunch wheatgrass (Elymus spicatum) is a characteristic species at low elevations with 

25-50 cm of precipitation annually (Williams 1961, 

1963). At cooler, more mesic high-elevation sites, Idaho 

fescue (Festuca idahoensis) becomes more common, often associated with shrubs such as threetip sagebrush 

(Artemisia tripartita). Wet meadows at high elevations 

have an abundance of alpine hairgrass (Deschampsia 

caespitosa).

In general, grasslands add variety to the landscape 

mosaic, occurring in places where the soil is too shallow, 

too dry, too fine textured, or too wet to allow tree seedling establishment (Knight 1994a; Doering and Reider 

1992). The physical location of many mountain meadows and grasslands has likely been stable for centuries due to site-specific abiotic conditions (Doering and 

Reider 1992).
68 USDA Forest Service RMRS-GTR-139. 2005

In some instances, however, changing climatic or 

edaphic conditions may allow tree seedling establishment 

on areas that were historically occupied by meadows. 

Some examples include the establishment of lodgepole 

pine in subalpine meadows in Yellowstone National Park 

(Jakubos and Romme 1993), and the growth of subalpine fir and Engelmann spruce near treeline in Rocky 

Mountain National Park (Hessl and Baker 1997). In both 

cases, climate changes after ~1870 apparently created 

favorable conditions for tree seedling establishment. Of 

course, the converse of this also may be true where climatic conditions following forest disturbances lead to 

the development (at least temporarily) of a mountain 

meadow (Billings 1969). Thus, while abiotic conditions 

can maintain relatively stable grasslands for long periods of time, the distribution of mountain meadows and 

grasslands undoubtedly changes as the climate changes.

On the MBNF, grasslands are scattered throughout all 

four units. In the Snowy Range and Sierra Madre they 

occur primarily as small meadows interspersed among 

lodgepole pine and spruce/fir forests, with the exception 

of a few large parks (e.g., Cinnabar Park, Dry Park, and 

Headquarters Park). Alpine grasslands, a form of tundra, 

are found directly surrounding the highest peaks of the 

Snowy Range, blending in with patches of krummholz 

composed primarily of stunted, wind-swept Engelmann 

spruce, subalpine fir, and, less frequently, limber pine. In 

the Laramie Mountains, grasslands are interspersed with 

ponderosa pine, lodgepole pine, limber pine, Douglasfir, aspen, and upland shrubs. Much of the Sherman 

Mountains unit is often described as grassland, though the 

abundance of the short three-tip sagebrush suggests that 

this vegetation type should be classified as a shrubland.

Succession in mountain grasslands is difficult to 

evaluate, but Costello and Schwan (1946) concluded that an association of Idaho fescue and mountain 

muhly (Muhlenbergia montana) is a subclimax vegetation type where ponderosa pine occurs in the Rocky 

Mountains. Shifts in species composition can occur as 

a result of grazing by large mammals. The review by 

Johnson (1956) of the effects of grazing by big game 

and livestock on pine-bunchgrass rangelands in Central 

Colorado is pertinent to the MBNF.

Various kinds of shrublands are also common. 

Dominant species include serviceberry (Amelanchier 

alnifolia), sagebrush (mostly mountain big sagebrush, 

Artemisia tridentata var. vaseyana), threetip sagebrush 

(Artemisia tripartita), fringed sagewort (Artemisia frigida), mountain mahogany (Cercocarpus montanus), antelope bitterbrush (Purshia tridentata), shrubby cinquefoil 

(Pentaphylloides floribunda), greasewood (Sarcobatus 

vermiculatus), and snowberrry (Symphoricarpos albus, 

occidentalis, and oreophilus). Additional species include 

chokecherry (Prunus virginiana), snowbrush ceanothus 

(Ceanothus velutinus) and wild rose (Rosa spp.). In general, all of the shrubs are characteristic of the mixed 

foothill shrubland cover-type (Knight 1994a). Their distribution is controlled largely by elevation, edaphic, and 

topographic influences.

Mountain big sagebrush, for example, often occurs 

where moisture is available but the soils are well drained 

and temperatures are not too cold (Burke and others 

1989). Mountain mahogany is found primarily on rocky, 

sedimentary substrates in the Sherman Mountains unit 

(Brooks 1962; Knight 1994a; von Ahlefeldt and Speas 

1996). Bitterbrush and snowberry often are mixed with 

mountain big sagebrush in the foothills, especially at 

lower elevations where snow accumulates, such as in 

ravines. At higher elevations and in association with 

wet meadows, shrublands dominated by mountain silver sagebrush and shrubby potentilla are common. In 

the eastern portions of the Laramie Mountains unit, at 

lower elevations, skunkbush sumac (Rhus trilobata) 

sometimes occurs with bitterbrush. In riparian zones, 

shrublands dominated by willow (Salix spp.) are common throughout the MBNF.

Another type of shrubland (or woodland) on the 

MBNF is dominated by Gambel oak (Quercus gambelii). Perhaps because its seedlings are apparently not 

able to tolerate spring frosts and low summer precipitation (Neilson and Wullstein 1983), the distribution of 

Gambel oak on the MBNF is limited to relatively low elevations of 1675 to 1980 m (5500-6500 ft), on the southwest side of the Sierra Madre near the Colorado state 

line (Knight 1994a). Gambel oak survival in this area 

probably is the result of increased summer precipitation due to the influence of the Arizona monsoon combined with relatively warmer temperatures below 1980 

m (6500 ft). Farther south, Gambel oak commonly occurs with ponderosa pine. Such woodlands can become 

quite flammable, but the oak sprouts vigorously and may 

increase in density after fires (Harper and others 1985). 

With fire suppression, the oak woodlands may be invaded by Rocky Mountain juniper.

Other types of shrubland are found only at high elevations in the MBNF. Where the subalpine and alpine 

zones come together, subalpine fir, Engelmann spruce, 

and sometimes limber pine form wind-swept, shrubby 

patches referred to as krummholz, which is classified as 

a shrubland in the MBNF database. In the alpine zone, 

shrublands of alpine dryad and willow occur in areas 

that are relatively moist or protected from the wind.

The alpine area surrounding Medicine Bow Peak 

provides an important element of biological diversity. 
USDA Forest Service RMRS-GTR-139. 2005 69

Numerous studies have been conducted on the geology, soils, vegetation, and grazing of this relatively small 

alpine area (e.g., Ray 1940; Mears 1953, 1962; Bliss 

1956; Billings and Bliss 1959; Billings and Mooney 

1959; Johnson 1962; Billings 1969; Thilenius 1975), 

many of which are reviewed in Knight (1994a) and von 

Ahlefeldt and Speas (1996). In general, the alpine vegetation of the Medicine Bow Peak area is dominated by 

short grasses, shrubs, and forbs that are able to withstand 

the severe environment characterized by high winds, low 

humidity, cold soil temperatures, high ultraviolet radiation, short growing season, low soil moisture, and great 

daily temperature fluctuations (Bliss 1956).

7.1. The effects of disturbances on non-forest 

vegetation

As with forest vegetation, the composition and structure of non-forest cover types is affected by both natural 

and human-caused disturbances. Fire and herbivory have 

been the primary historical disturbance agents in non-forest vegetation, but the characteristics of both have changed 

in the past century, primarily due to the introduction of fire 

suppression and grazing by domestic livestock.

Fire suppression probably has been in place long 

enough to cause comparatively high levels of shrub cover 

and the invasion of trees into some shrublands and grasslands. Indeed, the USFS now uses prescribed burning to 

reduce the cover of woody plants (Cook and others 1994), 

which stimulates the growth of herbaceous plants and 

slows the invasion of woody species. Long periods without fires could have occurred during the reference period 

as well, at least in some areas. If true, shrub and tree cover 

are still within the HRV (moderate confidence).

Where big sagebrush is a dominant species, vegetation changes can be quite dramatic after fire because this 

species is not capable of sprouting. When a fire occurs, 

grasses and forbs become the dominants for periods of 

several decades. Within approximately 10 years, however, big sagebrush usually will reinvade as new seedlings 

become established. Some grasses, such as bluebunch 

wheatgrass, may also decline in abundance following 

hot fires (or heavy spring grazing), but they, too, usually recover. In contrast, the other shrub species often 

resprout after a fire, such as Gambel oak, serviceberry, bitterbrush, and snowberry (Cook and others 1994). 

Consequently, their cover is rapidly restored, and sometimes their density increases to above pre-fire levels 

(Harrington 1985, 1987). Sprouting is an important adaptation for surviving fires, but some species are adapted in other ways as well. For example, the heat of fire 

facilitates the breaking of snowbush ceanothus seed dormancy (Zavitkovski and Newton 1968).

As shrub cover increases with time since fire, the 

flammability of the vegetation also increases due to increased amounts of small woody fuels. Prescribed fires 

can help prevent wildfires, but the presence of ranches, summer homes, and resorts in aesthetically pleasing 

foothill environments often restrict the implementation 

of this tool.

Bison and other herbivores probably grazed montane 

grasslands prior to the arrival of domestic livestock. If 

true, moderate grazing probably would not have driven 

variables beyond their HRV. However, sheep grazing at 

high elevations in the early 1900s appears to have exceeded the intensity of grazing by native herbivores. A longterm manager of the Sierra Madre unit, James Blackhall, 

wrote in 1916: “The grazing industry of this section is 

nearly extinct, brought on by overstocking, in fact with 

the exception of that portion which is protected by the 

long period that the snow lies on the ground, the short 

summer season and the early advent of winter, it may be 

said that the country is destitute of all forage necessary to 

support animal life . . .” (Bruce 1959, p. 51). Other parts 

of the MBNF probably experienced heavy livestock grazing as well during this time, pushing at that time various 

ecosystem variables beyond their range of means during 

the HRV reference period (high confidence).

Such heavy grazing intensities no longer occur on the 

MBNF (see section 3.3) and there is evidence that the vegetation is recovering. Still, livestock grazing was so widespread for so long that the effects are difficult to identify, 

even for trained observers (Thilenius 1975; Turner and 

Paulsen 1976; Fleischner 1994). Some plant species now 

may be quite rare because of livestock grazing, and some 

soil characteristics may not have been restored. Domestic 

livestock are thought to have different grazing and browsing habits than native ungulates, which could have accentuated the adverse effects of very high numbers. Our 

observations suggest that rangelands appear to be in fair 

to good condition in most areas, both in the upland and in 

riparian zones, but further research is necessary to know 

if the ecosystems are again within the HRV.

Other disturbances that affect the dynamics of nonforest vegetation are pocket gophers and the invasion 

of exotic plants. The burrowing of pocket gophers can 

be conspicuous where the animals bring large amounts 

of soil to the surface. Some of the displaced soil is in 

elongated mounds (middens), which are created during the winter as soil is moved into the animals’ snow 

tunnels. This burrowing greatly softens and aerates the 

soil profile. We have no evidence to suggest that the 
70 USDA Forest Service RMRS-GTR-139. 2005

densities of pocket gophers are within or beyond their 

HRV in different localities.

In contrast, exotic plants now are common in the meadows and shrublands of the MBNF and in some stands 

of aspen. Disturbances that accompany forest restoration, timber harvesting, livestock grazing, high intensity 

recreation, and road construction may facilitate the establishment and spread of exotic plants. Rapid increases in such plants can cause a decline in the abundance 

of some native plants and can cause changes to other 

ecosystem processes, such as the spread of fires. Exotic 

plants are still a minor problem in the post-harvest forests of the MBNF (Selmants 2000), but dramatic changes in non-forest vegetation could occur if exotic plants 

such as cheatgrass become more abundant (Keane and 

others 2002). The introduced bluegrasses may have less 

obvious effects on ecosystem variables, but in general 

their effects are still poorly known. Recent research on 

Rocky Mountain grasslands suggests that exotic plants 

are most likely to invade ecosystems where native plant 

diversity is high (Stohlgren and others 1999a), because 

of the favorable soil conditions found in such places, and 

that grazing probably has little effect on the accelerated 

spread of most exotic plant species across the landscape 

(Stohlgren and others 1999b).

In summary, intense livestock grazing in the early 

part of the 1900s, and fire suppression and the introduction of exotic plants in the mid-1900s, have caused 

changes over the years to non-forest ecosystems in some 

parts of the MBNF, probably pushing some variables 

beyond their HRV at that time. Livestock grazing has 

been greatly reduced in recent decades, and many of the 

rangelands probably are in better condition than before. 

However, insufficient research has been done to determine if some present-day ecosystem characteristics are 

within their HRV. Fire suppression during the last 50 

years or more probably has caused changes in the nonforest vegetation as well, but there may have been comparable periods without fires during the 1600s and 1700s. 

The adverse effects of exotic plants could become more 

severe in the future. Compared to grazing intensity and 

fire suppression, the spread of exotic plants is more difficult to control. However, the adverse effects of exotic 

plants are in many cases more subtle.

8. Summary of Probable HRV 

Deviations

Based on the information that we reviewed and our 

understanding of the ecology of the MBNF, it seems 

clear that some ecosystem variables could be outside 

their HRV or trending in that direction in areas where 

timber harvesting, livestock grazing, fire suppression, 

and the introduction of exotic plants have occurred. The 

removal of large wood is a type of disturbance that is different from any other with which the forest ecosystems 

have developed, with the result that both stand structure 

and landscape structure have been affected in some areas. Largely as a consequence of timber harvesting and 

fire suppression, ecosystem development is now proceeding in such areas in ways that are different than prior to the advent of current forest management practices. 

Does this matter? An answer is beyond the scope of this 

report. However, deviations from the HRV, where they 

occur, should be evaluated to determine if they could 

lead to undesirable consequences.

As our report shows, disturbances to the forests of 

the MBNF are not new. Indeed, the forest species have 

evolved with regular disturbances and the historical record indicates that there was a significant wave of human-caused disturbance in the late 1800s and early 

1900s. Current management practices often are occurring on forests that have already been subjected to at 

least one episode of timber harvesting and numerous 

episodes of burning. For this and other reasons that we 

have reviewed, there is great spatial and temporal variation in most ecosystem properties. It is reasonable to argue that the species that exist on the MBNF in the year 

2005 are those that survived rather intensive harvesting 

and burning caused by European-Americans during the 

late 1800s and early to mid-1900s. A century, however, 

is a short time in the development of forest ecosystems, 

and some of the early effects of humans could be augmented by the additional road construction and harvesting that have occurred since that time.

With the information available, we have attempted to 

draw conclusions about whether or not certain ecosystem variables are now beyond their HRV during the period 1600-1850 or trending in that direction (table 7). 

To summarize, we believe there is evidence to suggest 

that the following forest variables are within the HRV at 

high-elevations:

• Extensive, stand-replacing fires

• Insect population sizes

• Abundance of diseases (including dwarf mistletoe 

in many areas)

• Density of trees in all size classes (though the mean 

density may have been lowered in some areas)

• Tree and understory plant cover and diversity (number of species)

• Root/shoot ratios

• Forest floor depth
USDA Forest Service RMRS-GTR-139. 2005 71

Table 7—Summary of variables describing upland vegetation characteristics on the Medicine Bow National Forest relative to their estimated 

historic range of variability (HRV) from about 1600 to 1850, with an indication of our confidence in making each comparison. Stand 

characteristics are evaluated for areas modified by human activities only. Landscape characteristics are evaluated across the landscape for 

parts or all of the National Forest. For further discussion of each variable, refer to section 8 and the pages listed in the column on the right.

Within Outside Pages with discussion

Variable HRV HRV Confidence of variable

Disturbances

Stand-replacing fires at high elevations x Low 22-28, 52-58, 62, 66-70, 

 73-74

Surface fires at low elevations on dry sites x High 53-58

Insect outbreaks at high elevations x High 28-33, 52-59, 73-74

Disease outbreaks, except for dwarf x x Moderate 30-33, 59-60, 73-74

 mistletoe in some areas if white pine 

 blister rust 

 present

Wind x High 32-33, 60-61

Stand structure at high elevations

Density of trees in all age and size classes x High 18-22, 40-43

Tree cover and the density and size of x Moderate 18-22, 40-42

 canopy gaps

Understory plant density and cover x Moderate 42

Root/shoot ratios x High 42

Tree species and genetic diversity x High 42

Diversity of all plants x Moderate 42-43

Age and size-class structure of trees x Moderate 43

Forest floor depth x Moderate 43-44

Mineral soil affected x High 44

Snag density x High 44

Coarse woody debris x High 44-45

Landscape features at high elevations

Number and proportion of different land x High 8-12, 45-51

 cover types

Forest/non-forest land area ratio x High 50

Rate at which new, uniformly distributed, x High 18-22, 50

 mid-sized patches are formed

Proportion of forest land in different x Moderate 48-49

 stand-age classes

Proportion of land with high snag and x High 44-45, 50-51

 CWD densities (young and old 

 successional stages)

Proportion of land in edge, interior forest, x High 45-48, 50

 and patchy forest

Proportion of land with stands of high x Low 50-51

 tree density

Stand structure at low elevations

Species diversity x Low 61-63

Size- and age-class structure, canopy x Moderate 61-63

 cover, and tree and sapling density

Understory plant composition and cover x Low 61-63

Forest floor depth x High 63

Snag density x Moderate 62-63

Coarse woody debris x Moderate 62-63

Landscape features at low elevations

Proportion of landscape in high tree x Moderate 63-64

 density stands

Proportion of landscape with old trees and x Moderate 63-64

 high snag and CWD densities

Interspersion and proportion of landscape in x Moderate 63-64

 different cover types

Aspen

Representation in landscape x Moderate 64-67

Insect abundance x Moderate 66

Proportion of stands with conifers in understory x Moderate 64-67

Exotic plant species cover x High 67, 70

Non-forest vegetation Inadequate information to assess HRV, but 67-70

recovering from earlier periods of heavy grazing;

exotic plants are common in some areas
72 USDA Forest Service RMRS-GTR-139. 2005

• Proportion of the landscape in different cover types

• Ratio of forest to non-forest land

Similarly, the following forest variables appear to be 

within the HRV at lower-elevations:

• Aspen stand abundance and stand structure

• Abundance of native insects and diseases

• Understory plant species diversity (number of species)

In contrast, we believe other ecosystem variables are 

either beyond their HRV now or could be in the future if 

current management practices continue, as summarized 

below and in table 7.

Within stands at high elevations affected by 

timber harvest

1. Tree cover probably is lower and the number and size of 

canopy gaps probably is higher in harvested stands 

than the range of means for these variables during 

the HRV period in unmanaged forests of comparable age and site conditions (moderate confidence), 

due primarily to selective and shelterwood cutting 

but also due to fewer standing-dead trees.

2. Where timber harvesting has occurred on the MBNF, 

snag density and the amount of coarse woody debris is lower than the range of means for these variables during the HRV reference period for unmanaged stands of comparable age and site conditions 

(high confidence). Whole-tree yarding can accentuate this variation by concentrating coarse woody 

debris and slash at landings in a way that has not 

occurred previously.

3. Due to some methods of slash treatment, the intensity of soil scarification is beyond the HRV in some 

areas (high confidence). Burning can remove the 

forest floor, but the depth of soil disturbance after 

fire typically is less than disturbance following mechanical harvesting.

4. Age- and size-structure of managed stands within ravines or on north or leeward slopes is skewed toward smaller and younger trees than would have 

occurred during the HRV period, with less old forest, because of harvesting where fires would have 

been less likely to burn (moderate confidence).

5. Dwarf mistletoe abundance could be above its HRV 

in stands subjected to fire suppression and partial 

timber harvests that did not remove infected trees 

(moderate confidence).

Across the landscape at high elevations

6. Due to the timber harvests and human-caused fires 

in the late 1800s and early 1900s, the proportion 

of land in middle-aged forest (60-150 years old) is 

above the HRV (moderate confidence). Similarly, 

old-growth forest is less common on the landscape, especially in areas that historically had long 

fire-free intervals, such as in valley bottoms and 

on some north or leeward slopes (moderate confidence). Thus, ecosystem characteristics such as 

snags and coarse woody debris that depend on long 

periods of forest development may be less common across the landscape than during the reference 

period (high confidence).

7. The estimated rotation times for timber harvest in forests judged suitable for harvesting are generally 

shorter, on average, than the estimated site-specific mean intervals between natural disturbances in 

high-elevation forests. As a result, the rate at which 

new are patches are formed across the landscape 

probably exceeds the HRV (high confidence).

8. Characteristics of the size and spatial pattern of disturbances and forest patches across high elevation 

landscapes exceed their HRVs. Timber harvest typically produces mid-sized patches (i.e., 1-20 ha) that 

have less variability in size and shape than patches created by natural fires. The result is a decline 

in average patch size, the development of a more 

simple patch shape, less connectivity among patches of old forest, and a higher, more uniform level 

of landscape patchiness than would have occurred 

during the reference period (high confidence).

9. Due to roads and timber harvesting, the amount of 

edge is now higher and the land area of interior 

forests is lower, on average, than prior to the 1900s 

(high confidence).

10. The proportion of forested land with high tree density is somewhat lower because of partial thinning 

designed to maximize rates of tree growth (low 

confidence).

Within stands at low elevations affected 

by harvest, fire suppression, and livestock 

grazing

11. Fire suppression, livestock grazing, timber harvesting, and possibly climate change have led to the 

development of low-elevation forest stands with 

high densities of young trees, greater canopy cover, and less understory than usually occurred on 

many sites before 1850 (moderate confidence). 

Understory composition has also changed (low 

confidence) and forest floor depth has increased 

(high confidence).

12. By removing large trees and downed wood from the 

system, wood harvesting has lowered the abun-
USDA Forest Service RMRS-GTR-139. 2005 73

dance of snags and coarse woody debris in managed stands of low-elevation forests (moderate 

confidence). Harvesting in recent decades, where 

it has occurred, also has led to an increase in sapling density and a reduction in the average diameter of trees, thus creating younger, more uniformsized stands than existed before harvest (moderate 

confidence). Much of the harvesting that affected 

low-elevation forests on the MBNF occurred in the 

late 1800s and early 1900s.

13. Mean fire return intervals have become longer than 

the HRV at low elevations on drier sites, but fire 

intensity is probably higher due to the amount and 

continuity of fuels that have developed in some areas, and which could lead to more stand-replacing 

fires in the future (high confidence).

14. Forest floor depth may be deeper than the HRV at 

low elevations where fire suppression has been successful and prescribed burning has not been done 

(high confidence).

15. White pine blister rust outbreaks exceed the HRV in 

affected stands, as this disease is not native (high 

confidence).

Across the landscape at low elevations

16. Fire suppression, livestock grazing, and the logging 

of older, more fire resistant trees, where this has 

occurred at low elevations, probably has increased 

the amount and continuity of fuels to a point that is 

above the HRV, leading to potentially more standreplacing fires than occurred prior to 1850 (moderate confidence). Climate warming may contribute 

to this trend.

17. Fire suppression has reduced the level of interspersion of tree stands with grasslands (moderate confidence), and the average tree and sapling density has 

increased above the HRV (moderate confidence).

Aspen forests

18. In some aspen stands, grazing by livestock and native 

ungulates has reduced aspen densities and there has 

been an increase in the abundance of non-native 

species in the understory, to the point where such 

variables are beyond their HRVs (moderate confidence). Some aspen stands now have an understory 

of conifers, but this successional pattern probably 

occurred also during the reference period after long 

periods without fire (moderate confidence).

Non-forest vegetation

19. Some meadows, grasslands, and shrublands are 

probably grazed more, or have been in the early 

1900s, than during the HRV period. Reductions in 

livestock numbers since the early 1900s have been 

dramatic and rangeland condition is improving. 

Where heavy grazing occurred, some species probably declined in abundance, some probably became more abundant, and soil erosion probably occurred more rapidly (high confidence). Deviations 

from the HRV also were caused by the invasion of 

exotic plants that displace native species and cause 

changes in other ecosystem properties, but thus far 

their effects are not well known on the MBNF. In 

general, insufficient information is available for 

drawing conclusions about current HRV deviations 

in non-forest vegetation.

Frequently there has been a lack of pertinent information for the ecosystem variables we addressed in this 

report. Such variables can be identified as priorities for 

research, especially if they appear to have exceeded 

their HRV for the reference period. Research will also 

be required if decisions about the severity of a problem 

cannot be made with existing information or if ways of 

restoring the HRV for a specific variable cannot be identified.

Where deviations from the HRV do occur, various approaches can be used to work toward the goal of bringing 

variables within the HRV, if that is desired. In some cases, this may take many years to achieve. The challenge 

of bringing landscape variables within the HRV is likely 

to be greater on lands classified as suitable and appropriate for timber harvesting (about 37% of the MBNF at the 

present time). Such lands are typically the most productive and accessible, and therefore they have the highest 

potential for both timber production and for supporting 

the elements of landscape structure that are in decline, 

namely, large blocks of contiguous forest with relatively 

large trees.

As noted, it is beyond the scope of this report to evaluate the possible ecological consequences of exceeding 

the HRV. However, HRV analyses can guide discussions 

in the future as managers continue to develop ways of 

using Rocky Mountain forests in a sustainable way. 

Specific site conditions and management practices must 

be considered. Much has been learned about forest ecosystems, but the kinds of influences to which National 

Forests are now subjected are different than before—

and are occurring in a cumulative way. Human experience with modern forest management has been too short 

to speak with confidence about sustainability, but HRV 

analyses can help achieve that goal. Our report, which 

focuses primarily on upland vegetation, should be supplemented with HRV reports on (1) riparian and other 

wetland ecosystems, and (2) the population dynamics of 
74 USDA Forest Service RMRS-GTR-139. 2005

plant and animal species thought to be potentially threatened by management activities.

The kinds of deviations that we have identified are 

not unique to the MBNF. However, the MBNF and the 

Black Hills National Forest are the major sources of 

wood in Wyoming. The Black Hills have high rates of 

wood production because of the moderate climate and 

long growing season at low elevations. The MBNF is 

higher in elevation and less favorable for tree growth, 

but its terrain is comparatively flat and easily accessible, 

compared to, for example, the Shoshone National Forest. 

Also, much of the land area is forested, in contrast to the 

Bighorn National Forest, and only a small area of the 

MBNF is roadless, in contrast to the Shoshone National 

Forest and Bridger-Teton National Forest.

Another contributing factor for the comparatively 

high rate of wood harvesting in the MBNF may be the 

proximity of major markets for wood on the Colorado 

Front Range. However, those markets also have created 

demands for aesthetic and recreational amenities, with 

the result that it is no longer an easy decision to convert 

old forests with slow tree growth to young forests with 

rapid growth. Also, because of the construction of second homes adjacent to National Forests, no longer is it 

an easy decision to use surface fires to simulate natural 

processes, with the goal of maintaining or restoring the 

HRV for various attributes of stand structure.

Timber harvesting is sometimes promoted as a way of 

solving problems thought to be associated with fire suppression and the development of what appear to be unhealthy forests. Indeed, human values 100 years ago set 

in motion changes that have pushed some ecosystem variables outside their HRV. The high density of ponderosa 

pine in some areas is an example. However, sick and dying trees surely have always been a part of the MBNF—

sometimes over large areas. Such trees create habitat for 

native organisms that cannot survive on healthy trees. 

The biological diversity of the forest, which assures 

the long-term survival of forest ecosystems without human intervention, is higher because of periodic waves of 

mortality. Disturbances caused by European-Americans, 

however, can be different from those with which the native species have evolved. For some observers of management practices, forests labeled as unhealthy should 

be treated through harvesting the diseased and slow 

growing trees, if not all of the trees. That may be desirable in some situations, but such judgments should be 

preceded by analyses to determine whether or not fire 

suppression really has been effective and whether or not 

the “treatment” creates more problems than it solves. A 

stand of healthy trees surely is the goal of a tree farm on 

private land, but widespread stands of fast growing trees 

over much of a National Forest, where the pathogens are 

native, would not be within the HRV. Similarly, to harvest a forest to the point where it could not burn would 

exceed the HRV for some ecosystem variables. The continuing challenge of forest management is to determine 

if there are options for reducing the undesirable effects 

that can be associated with extracting wood, suppressing 

fires, and raising livestock while providing the other resources desired by modern society.

9. Literature Cited

Agee, J.K. 1997. Fire management for the 21st century. In: 

K.A. Kohm, and J.F., Franklin, eds. Creating a forestry for 

the 21st century: the science of ecosystem management. 

Washington, DC: Island Press: 191-201

Alexander, R.R. 1964. Minimizing windfall around clear 

cuttings in spruce-fir forests. Forest Science 10(2): 

130-142.

Alexander, R.R. 1966. Establishment of lodgepole pine 

reproduction after different slash disposal treatments. 

Res. Note RM-62. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Forest and 

Range Experiment Station.

Alexander, R.R. 1975. Partial cutting in old-growth lodgepole 

pine. Res. Pap. RM-136. Fort Collins, CO: U.S. Department 

of Agriculture, Forest Service, Rocky Mountain Forest and 

Range Experiment Station.

Alexander, R.R. 1986. Silvicultural systems and cutting 

methods for ponderosa pine forests in the Front Range of 

the Central Rocky Mountains. Gen. Tech. Rep. RM-128. 

Fort Collins, CO: U.S. Department of Agriculture, Forest 

Service, Rocky Mountain Forest and Range Experiment 

Station.

Alexander, R.R. 1987a. Ecology, silviculture, and management 

of the Engelmann spruce-subalpine fir type in the Central 

and Southern Rocky Mountains. Agric. Handb. 659. 

Washington, DC: U.S. Department of Agriculture, Forest 

Service.

Alexander, R.R. 1987b. Silvicultural systems, cutting methods, 

and cultural practices for Black Hills ponderosa pine. Gen. 

Tech. Rep. RM-139. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Forest and 

Range Experiment Station.

Alexander, R.R., and C.B. Edminster. 1980. Lodgepole pine 

management in the Central Rocky Mountains. Journal of 

Forestry 78: 198-201.

Alexander, R.R., G.R. Hoffman, and J.M. Wirsing. 1986. 

Forest vegetation of the Medicine Bow National Forest 

in southeastern Wyoming: a habitat type classification. 

Res. Pap. RM-271. Fort Collins, CO: U.S. Department of 

Agriculture, Forest Service, Rocky Mountain Forest and 

Range Experiment Station.

Allen, K.K., and J.L. Harris. 1999. Insects and diseases of the 

Bighorn National Forest. Internal report, Bighorn National 

Forest, Sheridan, Wyoming.

Allen, T.F.H., and T.W. Hoekstra. 1994. Toward a definition of 

sustainability. In: W.W. Covington, and L.F. DeBano, tech. 
USDA Forest Service RMRS-GTR-139. 2005 75

coords. Sustainable ecological systems: implementing an 

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