ponderosa pine

SCIENCE

FROM

Analysis

A Holistic Approach

Ecological

Restoration of Low-Elevation Dry Forests of the Northern Rocky Mountains

A bo u t t he W i l d e r n e s s S o c i e t y The Wilderness Society’s mission is to protect wilderness and inspire Americans to care for wild places. Founded by prominent naturalists and biologists, including Robert Marshall and Aldo Leopold, the organization played an important role in helping pioneer science-based conservation advocacy and policy making. The Society remains dedicated to the concept that careful, credible science, combined with bold advocacy, and unswerving vision is the key to conservation success. Headquartered in Washington, D.C., The Wilderness Society maintains twelve regional offices where our staff address on-the-ground conservation issues linked to local communities. Since spearheading passage of the seminal Wilderness Act in 1964, we have been a leading advocate for every major piece of Wilderness legislation enacted by Congress. Our effectiveness stems not only from our passion for protecting America’s most special places, but also from the sound scientific research that underpins every aspect of our work.

Abo u t t he Ec o l o gy a nd Economics Resear ch Depar tment The Wilderness Society’s Ecology and Economics Research Department (EERD) consists of experts in economics, ecology, and landscape analysis, including 12 Ph.D.-level scientists. This outstanding team provides the science to answer pressing questions about mineral exploration and development, forest and fire management, climate change, and many other issues affecting public lands. This information is key to understanding often complicated environmental issues, and ultimately making the right choices toward achieving lasting protection for the resources and places that sustain us and our ways of life. EERD provides science to inform not only The Wilderness Society’s own conservation campaigns, but also the decisions being made by communities, land managers, legislators, and others about the future of America’s wild places. The Wilderness Society is a national non-profit organization and was founded in 1935.

Restoration of Low-Elevation Dry Forests of the Northern Rocky Mountains A Holistic Approach

by

Michele R. Crist

Thomas H. DeLuca, Ph.D. Bo Wilmer and

Gregory H. Aplet, Ph.D.

January 2009

RESTORATION OF LOW-ELEVATION DRY FORESTS OF THE NORTHERN ROCKY MOUNTAINS PAGE i

Acknowledgments

The authors wish to extend their sincere appreciation to several colleagues within The Wilderness Society for their valuable comments on earlier drafts of this manuscript. These include, but are not limited to, John McCarthy, Scott Brennan, Dr. Joe Kerkvliet, and Tom Fry. The authors thank Dr. Charlie Luce with the Rocky Mountain Research Station of the US Forest Service for his valuable time and input. Thanks also go to Jill Grenon for helping with stewardship contracting research, to Sarah DeWeerdt for her tireless editing skills, to Mitchelle Stephenson for graphic design and preparation of the final document, and to Christine Soliva for her patience and logistical input on the production of this report. Special thanks to district rangers Tim Love and Neil Bosworth for helping us understand the on-the-ground application of forest restoration in the Northern Rockies and to Tom Tidwell, Regional Forester, for his support of regional collaborative forest restoration efforts and for his broad vision for forest restoration throughout the region. We also extend our deepest gratitude to the Aspenwood Foundation for their generous support of the forest restoration program and their devotion to the preservation of wild landscapes in the Northern Rockies.

Citation

Crist, M.R., T.H. DeLuca, B. Wilmer, and G.H. Aplet. 2009 Restoration of LowElevation Dry Forests of the Northern Rocky Mountains: A Holistic Approach. Washington, D.C.: The Wilderness Society.

Editor: Sarah DeWeerdt

Design/format: Mitchelle Stephenson

Printed in the United States of America by Todd Allan Printing on recycled paper.

© The Wilderness Society January 2009 1615 M Street, NW Washington, DC 20036 Tel: 202-833-2300 Fax: 202-454-4337

Web site: www.wilderness.org

This science report is one of a series that stems from conservation research studies conducted by The Wilderness Society’s Ecology and Economics Research Department. Other reports in the series that focus on similar issues include: Targeting the Community Fire Planning Zone: Mapping Matters; Ecological Analysis, May 2005, Wilmer and Aplet.

Managing the Landscape for Fire: A Three-Zone, Landscape-Scale Fire Management Strategy; Science and Policy Brief, September 2006, Aplet. Environmental Benefits and Consequences of Biofuel Development in the United States; Science and Policy Brief, May 2007, DeLuca.

To get copies of these reports, visit wilderness.org or contact the Ecology and Economics Research Department at 202-833-2300.

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Foreword

Low elevation, dry coniferous forests composed mainly of ponderosa pine are the most widely distributed and well known forests in the West, so much so that many consider the distribution of this forest an iconic “symbol of the American West.” Sadly, decades of timber extraction, fire suppression, and land development have degraded these forests in the Northern Rockies to a greater degree than any other forest type in that region—victims of accessibility, the common occurrence of fire, and the high value of the standing timber.

While restoration practices being applied throughout the Rocky Mountain West are aimed at reestablishing historical stand structure and reinvigorating natural forest processes, most of these efforts depend on examples from the southwestern United States where pure ponderosa pine forests experience repeated occurrences of short-duration, medium- to low-severity fires. In contrast, a mixed-severity fire regime is more characteristic of the forests found on low- to mid-elevations in the Northern Rockies. We must design restoration practices with these differences in mind. The effects of past management activities and fire suppression in the Northern Rockies are most evident at low elevations and on gentle slopes. Here, restoring a fire-resilient stand structure of large, widely spaced trees with a grassy understory requires thinning and prescribed burning. In higher and less accessible elevations, however, past management activities have had less effect. These forests need a different approach to restoration, one that focuses on the reintroduction of fire as a natural disturbance as well as the prevention of future catastrophic losses of existing old growth to stand-replacing wildfire. The Wilderness Society’s report, Restoration of Low-Elevation Dry Forests of the Northern Rocky Mountains: A Holistic Approach, stresses the important role of a holistic or whole-systems approach to forest restoration to ensure the long term ecological integrity of the region’s low elevation forests. Authors Michele Crist, Dr. Thomas DeLuca, Bo Wilmer, and Dr. Greg Aplet argue that successful restoration will require an adaptive approach, which accounts for historical land management activities, ecological cycles, biodiversity, and climatic cycles—and is appropriately responsive to effectiveness monitoring results. The report’s landscape analysis highlights two different restoration approaches in Montana and Idaho. Prescribed fire and thinning are appropriate for areas that have been logged and kept free of fire. They also have the potential to provide raw material for the wood products industry. However, thinning would not be appropriate for forests in roadless areas and wilderness where there is little or no history of past logging. These forests need only the return of fire to sustain their ecological integrity.

The Wilderness Society is working with citizens, policy makers, land managers, and timber industry representatives to help ensure that forest and fire policy ultimately support successful restoration activities that maintain a robust regional timber industry, improve the condition of degraded forests, and enhance the resilience of these low elevation landscapes to fire and climate change.

William H. Meadows President

Spencer Phillips, Ph.D. Vice President

Ecology & Economics Research Department

RESTORATION OF LOW-ELEVATION DRY FORESTS OF THE NORTHERN ROCKY MOUNTAINS PAGE iii

Table of Contents

Report Highlights ......................................................................PAGE iv

Introduction ..............................................................................PAGE 1 Ecology of the Northern Rocky Mountain Dry Montane Forest......................................................PAGE 4

Past Land Management Activities and Their Effects in Rocky Mountain Dry Montane Forests ...................PAGE 12 The Restoration Imperative: A Call for Holistic Landscape Management...................................PAGE 15 Opportunities for Dry Montane Forest Restoration in the Northern Rocky Mountains...............................PAGE 17 Elements of a Successful Restoration Program .............................PAGE 22

Conclusions .............................................................................PAGE 28 Literature Cited........................................................................PAGE 31

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Report Highlights

The dry montane forests of the northern Rocky Mountains form a majestic and archetypal landscape of the American West. Stands of pure ponderosa pine (Pinus ponderosa), or pine intermixed with Douglas-fir (Pseudotsuga menziesii) and western larch (Larix occidentalis), cover the lower slopes of these mountains and provide important habitat for a number of wildlife species.

Since European settlement of the region began in the nineteenth century, these forests have been affected by logging, grazing, road-building, and fire suppression, activities that have altered the structure of the forests (particularly by increasing stand density), reduced their ecological integrity, and increased their susceptibility to fires. Recent and ongoing climate change has further increased the danger of wildfires in these forests. Although the dry montane forest is naturally a fire-sculpted ecosystem, these changes have created a risk of more frequent and more severe fires that may even threaten the long-term persistence of the forest ecosystem.

As a result, the dry montane forest is now the main target for forest restoration and fuel reduction treatments throughout the Northern Rockies region. However, the model of restoration that has been applied in these forests originates in the ponderosa pine ecosystems of the southwestern United States, where broad, park-like stands of nearly pure ponderosa pine historically experienced nonlethal surface fire on a short return interval. By contrast, in the Northern Rockies, a patchier ecosystem of mixed stands typically experienced mixedseverity fires that burned hot in places and hardly at all in others, on a variable fire return interval.

In this report, we argue that successful forest restoration strategies for the Northern Rockies must take into account the specific ecology of forests in this region, as well as the history of land management activities in a particular place. The effects of past management activities have been greatest in the Northern Rockies’ more accessible ponderosa pine forests—those at low elevations and on gentle slopes. Here, thinning and prescribed burning may be needed to return to a historical stand structure of large, widely spaced trees with a grassy understory. However, the picture is different at higher elevations and on steeper slopes. In these less accessible locations, past management activities have not had as significant an effect on forest structure and function. These forests require a different approach to restoration, focusing on the reintroduction of natural fire and the prevention of future catastrophic loss of existing old growth. We also assess restoration opportunities in the context of surrounding human settlements. From this analysis it is clear that the majority of dry montane forest in Idaho and Montana (the bulk of the Northern Rockies region) is relatively accessible, occurring within approximately five miles of communities. In addition, most of these forests are likely to have suffered effects from past land management activities that make them vulnerable to fire, and may require thinning prior to the reintroduction of fire. Restoration of these forests has the potential both to improve community protection from wildfire and provide critical wildlife habitat.

Successful forest restoration strategies must take into account the specific ecology of forests as well as the history of land management activities in a particular place.

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Finally, we suggest that successful forest restoration requires attention to several interrelated key elements, including: • Use adaptive management, considering each management action as an experiment to inform future activities, and reorienting the management approach if the ecosystem’s trajectory is not meeting objectives • Set priorities, choosing restoration projects with an awareness of their potential for impact in a larger ecosystem context, and with the goal of maintaining all the characteristic elements of the forest mosaic • Take the long view, planning restoration strategies with an understanding of short- and long-term temporal variables such as seasonal cycles and centennial and millennial climatic cycles • Consider the whole ecosystem, with an awareness of the effects of restoration activities on wildlife species, non-native species, soil and soil processes, and insect and disease risks. Without careful planning, restoration treatments may lead to degradation of wildlife habitat, increased weed invasion, and degraded soil conditions • Be mindful of socioeconomic context, planning restoration projects with the existing infrastructure and economic base of surrounding communities in mind, and allowing ample opportunity for public comment and support from all stakeholders • Monitor restoration success with effectiveness monitoring of ecosystem health, aquatic integrity, fire hazard, and biodiversity variables Following these principles will increase the success of forest restoration efforts throughout the Rocky Mountain West. In turn, restoring natural conditions and ecosystem processes to dry montane forests will increase their ability to adapt to, and survive, a changing climate regime.

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Introduction

Although dry montane forest is a fire-maintained ecosystem that historically was sculpted by low- and mixed-severity burns, the history of management actions has made these forests more susceptible to stand-replacing wildfires. Changing climatic conditions are also increasing the frequency and severity of fire throughout the West, and further increasing the vulnerability of dry montane forests to stand-replacing fires (Boisvenue and Running 2006; Running 2006; Westerling et al. 2006).

As a result, the Rocky Mountain dry montane forest today is considered a threatened ecosystem (Mehl and Haufler 2001), and Idaho Partners-in-Flight (Ritters 2000) identified late-seral ponderosa pine as one of two “highest priority” habitats for restoration in Idaho. The loss and degradation of the dry montane forest has had corresponding consequences for the wildlife species that depend on this environment. Loss of ponderosa pine habitat has resulted in decreased range and population sizes for several species, and has also caused some species to be listed as threatened under the federal Endangered Species Act. Nineteen species on the Idaho Department of Fish and Game’s Priority Species List are either closely or generally associated with the late-seral ponderosa pine ecosystem (Mehl and Haufler 2001). Ecological restoration is defined as the process of assisting the recovery of an ecosystem that has been degraded, damaged, or destroyed (Society for Ecological Research 2004). Forest restoration has been broadly promoted as a means of recreating historical forest stand structure and ecological function, while reducing fuel loading and increasing resilience to fire (Arno et al. 1995; Mast et al. 1999; Covington 2000; Fiedler 2000; Covington 2003; Fiedler et al. 2004). The Healthy Forests Restoration Act (http://www.whitehouse.gov/infocus/healthyforests/), signed into law in 2003, proposes fuel reduction treat-

PHOTO BY GREG APLET

The Rocky Mountain dry montane forest—characterized by stands of ponderosa pine (Pinus ponderosa) either alone or in combination with Douglas-fir (Pseudotsuga menziesii) and western larch (Larix occidentalis)—is an important ecosystem of the American West, providing crucial habitat for a variety of wildlife species. However, exploitation of these forests for timber, grazing, and road-building since European settlement, as well as fire suppression over the past 100 years, has altered and degraded forest structure and function (Arno 1980; Barrett and Arno 1982; Arno et al. 1995; Fule et al. 1997; Mast et al. 1999; Moore et al. 1999; ICBEMP 2000; Baker et al. 2006), with an estimated 89 to 95 percent of the low-elevation forests having been altered by past human activities (Mehl and Haufler 2001).

An open, post-logging stand in Idaho’s Boise National Forest composed of large, old trees, abundant brush, and ponderosa pine and Douglas-fir regeneration.

RESTORATION OF LOW-ELEVATION DRY FORESTS OF THE NORTHERN ROCKY MOUNTAINS PAGE 2

ments geared toward reducing fire hazard, but also restoring historical forest structure and function.

In the Rocky Mountain West, the term “forest restoration” has typically been applied to forest management treatments that use silvicultural thinning and prescribed fire to reduce fuel loading and the potential for stand-replacing crown fires. This approach derives from a large body of research primarily conducted in the southwestern United States, where pure or nearly pure stands of ponderosa pine once grew on broad, gently sloping plateaus in a very dry climate, with light grass fires occurring every 3 to 7 years. This environment and fire regime produced vast, “park-like” stands of open forest, with large, widely spaced trees towering over a thick understory of grasses.

In southwestern forests, heavy grazing by livestock in the late nineteenth century consumed the grassy understory, subsequent logging removed the large, old-growth trees, and effective fire suppression disrupted the pattern of frequent fire. This resulted in the establishment of dense stands of small-diameter trees and a forest structure that now supports large, severe wildfires. In response, scientists and managers have advocated thinning of dense forests down to very low “presettlement” densities to restore conditions under which crown fire is unlikely. This “southwestern model” has become the standard for ecological restoration in ponderosa pine-dominated forests throughout the West. However, the dry forests of the Northern Rockies differ from those of the Southwest in several ecologically significant ways, raising questions about the appropriateness of applying the southwestern model to these forests. In the Northern Rockies, a wetter, cooler climate leads to less frequent fires, even in pure stands of ponderosa pine. The admixture of Douglas-fir and western larch creates different conditions for burning, and the complex terrain of steep mountain slopes results in fires that burn fiercely in some places and hardly at all in others. The resulting “mixed-severity fire regime” historically yielded a complex mosaic of patches of large, old trees and young, recovering forest.

Inappropriate application of restoration treatments on a landscape may lead to failed restoration efforts (DellaSala et al. 2003). Effective forest restoration, sustainable forest management, and re-establishment of natural wild landscapes can only be realized when land management is conducted in the context of the historical range of conditions and intended future land use objectives (DellaSala et al. 2003; Brown et al. 2004; Noss et al. 2006). Thus, the ecological uniqueness of the ponderosa pine ecosystem in the Northern Rockies suggests the need for a different model of forest restoration in this region (Baker et al. 2006; Hessburg et al. 2007). We recommend two primary approaches to forest restoration in the Northern Rockies, depending on the ecology and land use history of specific sites within this ecosystem. At the lowest elevations and on gentle slopes in the Northern Rockies, historical forest characteristics were similar to those of the Southwest, with frequent surface fire and widely spaced trees. Also like the Southwest, these forests have been heavily roaded and logged, resulting in the establishment of dense stands of small trees. The appropriate restoration model here may indeed look much like the southwestern “thin and burn” approach. On steeper slopes, fire burned hotter and patchier and resulted in a more complex forest mosaic. These steep slopes were also less favored for logging (and

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less accessible for grazing and road-building), and still retain many large trees. Their structure has been largely unaffected by logging, and only somewhat affected by fire exclusion in the twentieth century. Forest restoration there requires only the return of fire.

PHOTO COURTESY US FOREST SERVICE PHOTO COURTESY US FOREST SERVICE

In this report, we describe the ecology and ecological values of the Northern Rocky Mountain dry montane forest and describe an appropriate restoration approach that adequately addresses the history, diversity, complexity, and stochasticity of this important forest ecosystem. We begin by providing an overview of the ecology and natural disturbance regime of the dry montane forest in the Northern Rockies, then detail the history of land management activities and how they have affected different parts of the ecosystem. This leads us to a holistic vision for restoration of the dry montane forest ecosystem that takes into account its broad variability on the landscape. We also describe opportunities across the region where different restoration activities may be warranted, and address the elements of a successful restoration program, including the importance of monitoring the success of restoration efforts.

Two historical images from Idaho’s Boise National Forest: The top photo, taken in 1913, shows evidence of a recent fire in the natural stand structure of a dry montane forest. The bottom photo, taken 18 years later in 1931, shows the same forest with advanced regeneration that has changed the stand structure.


Ecology of the Northern Rocky Mountain Dry Montane Forest

Using past and recent research, historical photographs, and written accounts, we have pieced together a characterization of the historical stand structure, pattern, composition, and function of dry montane forests in the Northern Rocky Mountains, the distinct floristic zone comprising northwestern Wyoming, Montana, Idaho, northeastern Oregon and Washington, and southern Alberta and British Columbia (Peet 1988). These forests have a patchy and variable structure, a result of variations in elevation, precipitation, topography, overstory/understory dynamics, and natural disturbance regimes, especially fire.

Forest Composition and Structure

The dry montane forests of the Northern Rocky Mountains are mainly composed of stands of pure ponderosa pine, or ponderosa pine intermixed with Douglas-fir and western larch. These tree species thrive on sites that are relatively warm and dry, and tend to dominate at the lower elevations in the Rocky Mountains, approximately 3,200-7,000 feet. Average temperature is relatively consistent throughout the dry montane forest ecosystem, and typical of the conditions throughout the range of ponderosa pine (i.e., average highs of ~85° F in summer and average lows between 10 and 20° F in winter). Annual precipitation ranges from less than 15 inches per year in eastern Montana to more than 25 inches per year in northern Idaho and mostly comes as snow during the winter months. This situation differs greatly from that of the ponderosa pine forests of the Southwest, where consistently dry springs result in early-summer fires followed by a late-summer monsoon (Oliver and Ryker 1990). Compared to the forests of the Southwest, the ponderosa pine forests of the Northern Rocky Mountains generally experience more variable weather conditions from year to year, and during drought years the likelihood of fire typically increases throughout the summer. Species composition of the dry montane forest varies across the region, primarily depending on elevation, temperature, and precipitation conditions. Briefly, ponderosa pine dominates at low, warm, and dry locations, with Douglas-fir becoming increasingly prominent at higher, colder, and wetter locations. Forests also become denser with increasing elevation. At about 3,200-5,500 feet, ponderosa pine begins to appear at the fringes of the arid sagebrush steppe characteristic of the foothills and large valleys in the Rocky Mountains. At these lowest elevations, ponderosa pine trees are sparse and randomly dispersed, but as elevation increases the trees become more numerous, eventually forming open forests on the warmest, driest sites (annual precipitation around 11-17 inches) (Oliver and Ryker 1990). As elevation and annual precipitation increase, the open stands of ponderosa pine become denser, and Douglas-fir and western larch begin to appear. At higher elevations (approximately 5,000-7,000 feet), tree density varies depending on the natural disturbance regime and/or other abiotic factors. At these elevations, Douglas-fir may dominate on sites that receive 20-25 inches annual precipitation, while on drier sites ponderosa pine and Douglas-fir may form codominant stands. Above 7,000 feet, ponderosa pine, Douglas-fir, and western larch become minor associates in the subalpine forests.

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The highly dissected topography of Idaho and western Montana also affects forest ecology. There, mixed ponderosa pine/Douglas-fir forests may be found on more mesic sites at low elevations, especially on north-facing slopes. Similarly, pure ponderosa pine stands may be found at higher elevations on relatively warm, xeric sites such as south-facing slopes and exposed ridges.

Soil and geologic conditions at depth can also influence forest composition, making elevation and aspect imperfect guides. For example, limestone residuum may have finely textured surface soils, but it is highly fractured and well drained at depth. This creates extremely dry soil conditions in summer, thus favoring the presence of very open stands of Douglas-fir at elevations above 7,000 feet (Nimlos and Tomer 1982). Similarly, coarse-textured, low-fertility parent materials such as quartzite colluvium are particularly low-productivity soils (Mandzak and Moore 1994), and may yield dry forest conditions at higher elevations and on north-facing slopes. Volcanic ash capped soils are highly productive (with high water-holding capacity and high fertility), and tend to support more mesic vegetation at lower elevations.

The composition of the understory plant community in dry montane forests also varies depending on a combination of abiotic factors (e.g., topography, elevation, moisture, soils) and the natural disturTABLE 1. bance regime. Typically, at lower elevations and on south-facing slopes, the understory may be rather sparse due to the lack of moisture and prevalence of coarse-textured soils. Species that predominate here are bunchgrasses, forbs, and small shrubs, with a high density of shrubs seen throughout Common Name some stands (Leiberg 1897, 1899a, 1899b, 1899c, 1900a, Saskatoon serviceberry 1900b; U.S. Forest Service 1937). On more mesic sites, Bearberry understories can range from a patchy distribution of shrubs Heartleaf arnica interspersed with grasses and forbs, to a denser distribution Sagebrush of tall shrubs, forbs, and grasses (Leiberg 1897, 1899a, Arrowleaf balsamroot 1899b, 1899c, 1900a, 1900b; U.S. Forest Service 1937). Pinegrass While the landscape distribution of understory grasses, forbs, Elk sedge and shrubs varies, the species composition of understories Sun sedge remains relatively similar throughout the dry montane zone Tobacco bush (Table 1). Mountain mahogany

Common Understory Species of the Northern Rocky Mountain Dry Montane Forest

The dense, patchy, or sparse distribution of tree species also depends on overstory-understory dynamics, which refers to the effects of tree canopies on ground-layer plants including shrubs, herbaceous vegetation, and cryptogams (mostly mosses and lichens on the soil surface), as well as tree seedlings. Most research has focused on competition between species in the herbaceous layer, such as grasses, shrubs, and seedlings. Competition for light (Moir 1966), nutrients (Elliott and White 1987), water, and a combination of these understory influences can affect tree seedling growth and survival, and in turn determine the character of the stand in the next generation. In return, continual growth of the overstory pines and other tree species reduces cover, density, and biomass of many understory species, especially those that grow best in full sunlight such as bunchgrasses and shrubs.

Idaho fescue Rough fescue Common juniper Lupine Oregon-grape Mountain sweetroot Ninebark Chokecherry Bluebunch wheatgrass Antelope bitterbrush Russet buffaloberry White spirea Snowberry Western meadowrue

Scientific Name Amelanchier alnifolia Arctostaphylos uva-ursi Arnica cordifolia Artemisia spp. Balsamorhiza sagittata Calamagrostis rubescens Carex geyeri Carex inops Ceanothus velutinus Cercocarpus montanus Festuca idahoensis Festuca scabrella Juniperus communis Lupinus spp. Mahonia repens Osmorhiza chilensis Physocarpus malvaceus Prunus virginiana Pseudoroegneria spicata Purshia tridentata Shepherdia canadensis Spiraea betulifolia Symphoricarpos spp. Thalictrum fendleri

Source: Pfister et al. 1977; Cooper et al. 1991.


Ponderosa pine forests of the Southwest are characterized by contiguous open stands of large-diameter trees that once covered thousands of acres, but in the Northern Rockies, tree-ring reconstructions and forest reserve reports reveal that tree density was historically highly variable from place to place (U.S. Forest Service 1937; Baker et al. 2006). Some stands contained a dense floor of saplings (ponderosa pine, Douglas-fir, and larch) amongst the large-diameter trees; other stands were very open, with little large woody debris and few seedlings and saplings (Brown et al. 2004; Baker et al. 2006). In Montana, reconstructions indicate tree densities in mature ponderosa pine forests around the year 1900 of 116 to 249 trees per hectare, though data are lacking for young stands but forest reserve reports (Leiberg 1897, 1899a, 1899b, 1899c, 1900a, 1900b; U.S. Forest Service 1937) for the Idaho and Montana Rockies document an even wider variability, reporting tree densities (for all ages) ranging from 17 to 19,760 trees per hectare in ponderosa pine forests and 39 to 7,410 trees per hectare in Douglas-fir forests around 1900 (Baker et al. 2006).

Disturbance

As in other forests of western North America, the landscape mosaic of tree density in dry montane forests of the Northern Rockies is greatly influenced by disturbance events—in the case of these forests primarily fire, but also insects, disease, and windthrow (Figure 1). Generally, dry montane forests experience a prolonged dry season every year (Agee 1993), which creates the conditions that support fire. The soil surface typically dries to the wilting point sometime between late June and the end of July; in mid-August the wilting point depth can exceed 20 inches (Pfister et al. 1977). By late June or July plants with shallow roots, including grasses, forbs, and some shrubs and young trees, become highly flammable (Peet 1988), and fires are ignited by lightning strikes. After a fire it takes approximately three years for fuels to reaccumulate to a level capable of carrying another fire (Heyerdahl 1997). Leaf litter and woody debris decompose relatively slowly in these dry forests, and therefore accumulate readily during fire-free periods (Fischer and Bradley 1987).

Overall, the dry montane forest ecosystem in the Northern Rocky Mountains is characterized by a mixed-severity fire regime, with the extent and severity of fires varying from place to place and depending on the amount and configuration of the flammable understory vegetation and abiotic factors (Baker et al. 2006; Hessburg et al. 2007). At lower-elevation, drier sites in the ponderosa pine zone and under moderate weather conditions, low-severity surface fires are promoted by low-density forest with a grassy understory and the ability of mature ponderosa pine to resist damage by fire (Barrett 1988; Wu 1999; Veblen et al. 2000; Arno and Allison-Bunnell 2002; Ehle and Baker 2003; Baker et al. 2006). Such fires are apt to creep through the forest understory, with occasional flare-ups in unusually dry areas, where there are dense fuels, and on steep slopes (Arno 1980). This fire regime produces forest stands characterized by groups of widely spaced trees with sparse understories. These fires also prune low branches from older trees, decreasing chances of a stand-replacing fire. Data from fire history studies, forest reserve reports, and atlases indicate that dry montane forests at low-to-mid elevations historically had a mean fire return interval of approximately 25 to 50 years (Arno 1976, 1981; Arno and Gruell 1983; Barrett 1981, 1991, 1993, 1997, 1999; Habeck 1992, 1994;

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FIGURE 1.

Historical Photos Illustrating Natural Disturbance in Dry Montane Forests Crown Burn on a north-facing slope

Crown Burn

Severe Surface Burn Light Surface Burn

In 1932 a survey was made of the 1931 Boise Basin mixed severity fire that occurred on the Boise National Forest in Idaho. This photo series depicts the burn severity conditions assessed by Forest Service researchers during this survey. The researchers stated that "fire damage in ponderosa pine forests should no longer be considered as being caused largely by surface fires.”

Ground Surface Burn Moderate Surface Burn


TABLE 2.

Evidence for High-Severity Fires in the Northern Rocky Mountain Ponderosa Pine Forest • Some ponderosa pine/Douglas-fir forests burned severely in the late 1800s and early 1900s during regional drought years (Barrett 1997; Brown et al. 1999). • In Montana, tree-ring studies show that some dry montane forests had infrequent high-severity fires as well as more frequent low-severity fires (Barrett 1988; Arno et al. 1995, 1997). • Numerous forest reserve reports also indicate that mixed- and high-severity fire occurred in pure ponderosa pine forests across the Northern Rockies (Leiberg 1897, 1899a, 1899b, 1899c, 1900a, 1900b).

Lehman 1995; Losensky 1993a, 1993b, 1993c, 1993d; Pierce 1995a, 1995b, 1995c; Zack and Morgan 1994). Fires of such frequency favored regeneration of conifers like ponderosa pine by exposing mineral soils.

At mid-to-high elevations in the ponderosa pine zone and on steep slopes, by contrast, forests are predisposed to experience some degree of crown fire on a regular basis (Hessberg et al. 2008). However, as has been documented for the Central Rockies, finescale abiotic factors also account for smaller areas of predominantly low-severity fire at these elevations (Sherriff 2004). Fire history data and forest age structures show variation in the history of fire severity along elevation, topographic, and moisture gradients within ponderosa pine forests. These local variations in moisture availability and other factors determine fuels productivity and vegetation attributes (Peet 1988; Sherriff 2004). We refer to fires that burn hot in some places and hardly at all in others as mixed-severity fires.

Although ponderosa pine forests are usually thought to experience only surface fire, there is ample evidence (Table 2) that more • Ogle and DuMond (1997) found that high-severisevere fires also occurred in Rocky Mountain ponderosa pine ty fires were reported during early examinations of ponderosa pine stands in several Idaho national forests. For example, the U.S. Forest Service (1937) described five forests between 1900 and 1915. severity classes for the Big Basin Fire of 1931, an extensive fire • The 1931 Big Basin Fire on the Boise National that burned ponderosa pine stands in the Boise National Forest. Forest was described as a mixed-severity fire that The severity classes range from ground fire to a patchy distribucreated a mosaic of burned conditions across an tion of alternating severities to stand-replacing burns. Standexpansive area (U.S. Forest Service 1937). replacing events were expansive and mainly occurred on northfacing steep slopes. They covered entire hillsides, burning hundreds or even thousands of acres depending on weather and terrain. Standreplacing events in ponderosa pine forests, similar to the ones that occurred in the Big Basin fire, are likely to occur on steep terrain and during drought years (Arno 1980; Baker et al. 2006). In addition to fuels and topography, climate also has a strong influence on fire patterns in the Northern Rockies. Several climatic processes (e.g., El Niño Southern Oscillation, Pacific Decadal Oscillation) combine and cause fluctuations in Pacific Ocean and Atlantic Ocean surface temperatures, which alters precipitation and temperature patterns on the scale of several years to a few decades and causes corresponding fluctuations in fire activity throughout the West (Cayan 1996; Mantua et al. 1997; Dettinger et al. 1998; Western Regional Climate Center 1998; McCabe and Dettinger 1999; Brown 2006; Sibold and Veblen 2006; Kitzberger et al. 2007). Periodic climatic fluctuations that result in warmer and drier winters tend to result in more fire activity during the late spring and summer season compared to periods of wet and cool winters (Cayan 1996; Mantua et al. 1997; Dettinger et al. 1998; Western Regional Climate Center 1998; McCabe and Dettinger 1998). However, cool, wet winters may result in a large growth of understory fuels which, if a warm, dry spring and summer follow, may increase the chance of fire.

Knowledge of the variability in climate over large time-scales yields the understanding that Rocky Mountain ponderosa pine forests have survived, adapted, and even thrived during dramatic changes in climate over the past thousands of

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years. Looking back over millennia reveals 132 Years that decades-long colder periods may promote frequent, low-severity fires, which fuel and in turn are fueled by increased understory growth; in contrast, warmer periods create the conditions for severe droughts and thus promote the occurrence of mixed-severity and stand-replacing fires. Meyer and Pierce (2003) and Pierce et al. (2004) found that during the Little Ice Age of AD 1500-1900, moderate annual to multi-annual droughts produced frequent fires in xeric ponderosa pine forests, primarily because of the growth and desiccation of grasses and other fine surface fuels. Relatively severe droughts during the Little Ice Age may have resulted in some stand-replacing fires, as suggested by other climate studies and tree-ring studies in other ponderosa pine forests (Shinneman and Baker 1997; Brown et al. 120 Years 1999; Veblen et al. 2000; Pierce et al. 2004) and noted by early explorers and scientists. But overall, fires in the Little Ice Age were less frequent and less severe compared to other time periods. By contrast, periods of warmer climate with intervals of severe drought, such as during the Medieval Warm Period (approximately AD 9001200) and from approximately 350 BC-AD 400, appear to promote high-severity fires in ponderosa pine forests. During these periods, several decades of unusually wet summers that occurred between severe droughts may have contributed to high tree densities that in turn increased the likelihood of stand-replacing fire (Meyer and Pierce 2003; Pierce et al. 2004).

An understanding of how forests have reacted to historical climate variability becomes even more important in light of recent and ongoing patterns of climate change. Over the last century, the Northern Hemisphere has experienced a rapid rise in temperatures (Grove 1988; Pollace et al. 1998; Mann et al. 1999). This warming has affected much of the western United States and has been accompanied by a 5 to 20 percent decrease in precipitation in Northern Rocky Mountain ponderosa pine environments (Karl et al. 1996). The fire-season Palmer Drought Severity Index, which correlates strongly with burn area in the Northern Rockies (specifically central Idaho and Yellowstone) (Meyer et al. 1995; Westerling et al. 2006), shows a significant increase in drought severity between 1895 and 2002. The late-twentieth-century increase in wildfires in the Northern Rockies has been attributed to the cumulative effects of fire suppression and other management practices, but recent studies have shown that it is likely also driven by climate variability and/or climate change (Meyer and Pierce 2003; Whitlock et al. 2003; Westerling et al. 2006; Heyerdahl et al. 2008; Whitlock et al. 2008). For example, Westerling et al. (2006) demonstrated that the increase in regionally synchronous fires in the late twentieth century was likely a result of warmer springs and longer fire seasons due to climate change. The current trend

45 Years

65 Years

This photo series, taken in 1934 depicts even-aged dry montane forest structures likely resulting from mixed-severity fires. Forest stand ages range from 45-132 years. The first photo is of an evenaged ponderosa pine stand 45 years old near New Meadows, Idaho. The second photo is of an even-aged ponderosa pine stand 65 years old in Steamboat Gulch near Idaho City, Boise N. F. The third photo is of an even-aged ponderosa pine stand 120 years old on the Boise N. F. The fourth photo is of an even-aged ponderosa pine stand 132 years old on Riordan Creek, Payette National Forest, Idaho.


of continued warming is thought to portend an increase in synchronous fires throughout the West (Westerling et al. 2006).

The density of a dry montane forest stand is influenced by its history of fires of varying severity. Following a low-severity fire, seedling establishment often increases, because increased soil moisture and nutrients from exposed mineral soils and scorched needles enable the seedlings to compete successfully with bunchgrasses (Sackett 1984; Boyce 1985; Bonnet et al. 2005). The regenerating young trees may be killed in the next surface fire; thus, long-term persistence of ponderosa pine in a low-severity fire regime likely requires a fire-free period of several decades or more. Shorter fire-free intervals have been found in other studies to cause periodic cohorts of ponderosa pine and other shade-tolerant trees (Agee 1993); however, this is not always the case. Where a seed source is lacking, similar to what has been found in the Central Rockies, it may take up to 100 years for trees to re-establish after a fire (Kaufman et al. 2000; Romme et al. 2006).

In areas where mixed- or high-severity fire occurs, regeneration has also been found to result in high-density stands. According to tree-ring studies in Montana and forest reserve reports, a dense understory of ponderosa pine and Douglasfir seedlings occurred after high-severity fires in several national forests; this high density persisted for decades and resulted in a denser stand of mature trees (Leiberg 1899a, 1899b, 1899c; U.S. Forest Service 1937; Arno et al. 1995, 1997; Ogle and Dumond 1997; Baker et al. 2006). As higher-density stands age during long fire-free intervals, conditions develop that increase the likelihood of stand-replacing fire. Alternatively, regeneration after a high-severity fire could be delayed in openings and may take over a century or more to fill in (Graves 1899; Leiberg 1899a, 1899b, 1899c; Kaufmann et al. 2000). Similar to what has been found in the Central Rockies, this historical variability was likely due to site conditions (Peet 1988) and often resulted in a mosaic of patches of different sizes composed of differing tree ages and densities.

Another determinant of regeneration patterns is nutrient availability, which serves as a significant driver of forest productivity and stand dynamics in lowto mid-elevation ponderosa pine/Douglas-fir forest ecosystems (Fisher and Binkley 2000). Fire directly influences nitrogen availability and soil nitrogen content (Smithwick et al. 2005), resulting in an immediate, yet often shortterm, increase in nitrogen availability (Monleon et al. 1997; Neary et al. 1999; DeLuca and Zouhar 2000; Neary et al. 2005). Because nitrogen is the primary limiting nutrient in these forests (Mandzak and Moore 1994), this nitrogen pulse in turn influences understory and stand growth rates (Hart et al. 2005). Exclusion of fire from fire-maintained ponderosa pine ecosystems results in a reduction in nitrogen availability and a shift in community structure to a more shrub-dominated understory and a greater abundance and volume of Douglasfir (Mackenzie et al. 2004; DeLuca and Sala 2006; Keeling et al. 2006). This shrub-dominated understory may further reduce nitrogen availability, because it is associated with a greater presence of soluble phenolic compounds in the forest floor which immobilize nitrogen (MacKenzie et al. 2004). In addition to fire, many other natural disturbances including insects, pathogens, and windstorms influence a forest’s composition, structure, and landscape pattern (Lundquist 1995). For example, bark beetles and defoliators,

PAGE 11

such as western spruce budworm (Choristoneura occidentalis) and Douglas-fir tussock moth (Orygia pseudotsugata), can cause widespread tree mortality, reduce tree density, and open up forest canopies over hundreds of thousands of acres (Schmid et al. 1994; Schmid and Mata 2005). Many overstory-understory relationships are associated with tree death and falls (Denslow and Spies 1990). Historically, insects, disease, and windstorms interacted with fire and topography to create a complex mosaic of forest structures, with canopy gaps, multi-aged stands, and variations in forest composition. Restoration of managed forest ecosystems must consider the role of all these disturbances in sculpting the character of natural forest ecosystems of the Northern Rockies.

FIGURE 2.

Species Closely Associated with Ponderosa Pine Forest in the Northern Rockies

Flammulated Owl

Forest Structure and Wildlife Habitat

Wildlife species depend on the diverse landscape characteristics of the dry montane forest produced and maintained by fires and other disturbances of varying intensity and frequency. Different species respond to different elements in the mosaic—the large contiguous patches of varying tree densities, clumps of large trees with interlocking crowns that produce structurally complex canopies, a patchy distribution of plants in the understory, and variable densities of snags and coarse woody debris (Saab and Powell 2005). Flammulated owls (Otus flammeolus) and Idaho ground squirrels (Spermophilus brunneus) are associated with mature and old-growth xeric ponderosa pine/Douglas-fir stands with low to moderate canopy closure, but are absent from higher-elevation, more mesic ponderosa pine/Douglas-fir stands (McCallum 1994; Wright et al. 1997; Idaho Department of Fish and Game 2005).

Large, contiguous, late-seral stands of ponderosa pine are especially important for sustaining certain species of birds (Figure 2) and other wildlife (Dixon 1995; Idaho Department of Fish and Game 2005; Saab and Powell 2005; Squires and Kennedy 2006). The Idaho Department of Fish and Game has identified six species as “closely associated” with large, contiguous patches of old-growth ponderosa pine forests, including the white-headed woodpecker (Picoides albolarvatus), flammulated owl, pygmy nuthatch (Sitta pygmaea), northern goshawk (Accipiter gentilis), great gray owl (Strix nebulosa), and long-legged myotis (Myotis volans) (Mehl and Haufler 2001). Smaller stands of ponderosa pine are often unused by white-headed woodpeckers, for example, even though other habitat characteristics are present (Dixon 1995).

Elk (Cervus elaphus) and deer (Odocoilius spp.) are dependent upon variable forest structure due to their preference for “edge” habitat, which enables them to browse in open stands while using clumps of trees for cover. Research in southwestern Idaho demonstrated that high densities of standing snags created by stand-replacing burns in ponderosa pine/Douglas-fir forests provide crucial nesting habitat for cavity-nesting birds, such as Lewis’s (Melanerpes lewis) and white-headed woodpeckers, and are also important foraging habitat for several woodpecker species (Dixon and Saab 2000; Nappi et al. 2004). Large-diameter snags are especially important because they remain standing longer than smaller ones (Ganey and Vojta 2005).

White-headed Woodpecker Flammulated Owl Pygmy Nuthatch Northern Goshawk Great Gray Owl Long-Legged Myotis Lewis’ Woodpecker White-Breasted Nuthatch Yellow Pine Chipmunk Idaho Ground Squirrel

White-headed Woodpecker

Mehl and Haufler 2001 and Idaho Partners in Flight 2000


Past Land Management Activities and Their Effects in Rocky Mountain Dry Montane Forests

Many management activities have interacted to influence and alter the stand structure, composition, and function of forests throughout the United States. In the dry montane forests of the Northern Rockies, these primarily consist of logging, road-building, fire suppression, grazing, and the introduction of nonnative (especially invasive) plants. The uneven distribution of these activities across the landscape has produced a variety of effects on ecological integrity, the most dramatic of which occur in relatively accessible low-elevation forests (Hann et al. 1997).

Vast areas of the Northern Rockies were subject to intense grazing pressure beginning in the 1800s (Borman 2005). Large cattle herds overgrazed and trampled vegetation across large areas of public lands, particularly in the dry forests and riparian zones where they could access water. Because sheep require less water and can graze more successfully throughout the landscape than cattle, introductions of sheep boomed in the mid- to late 1880s with the decline of cattle and the completion of key railroad lines. Widespread high-grade logging and selective cutting began in the late 1880s through early 1900s throughout the Northern Rockies. Most of this logging occurred at low elevations in more xeric forests and on relatively gentle slopes (gradients less than 35 percent). Many of the oldest and largest ponderosa pines were removed. This type of logging also left behind high fuel loads that contributed to the large, human-caused fires that swept across the landscape during the early 1900s.

Between the end of World War II and the early 1960s, markets for timber species other than ponderosa pine were expanded and improved. Much of the readily accessible ponderosa pine was gone by the 1960s (Robbins and Wolf 1994; Langston 1995; Robbins 1999), but harvest of Douglas-fir, western larch, and other species continued to increase through the 1980s, when large-scale logging began to decrease (Hessburg and Agee 2003).

From the 1940s to the 1980s, continually improving technology allowed the construction of extensive road networks to access remote areas for timber removal, fragmenting contiguous forests into smaller and smaller patches. For example, the Boise National Forest in Idaho has 5,477 miles of inventoried roads, the majority of which were built for timber and mineral extraction. Like the Boise National Forest, many public lands throughout the Northern Rockies have experienced substantial, and in some cases extremely heavy, road construction since 1950, especially on relatively low-gradient, accessible slopes. This road network has come at a cost, including degradation of natural landscapes and other environmental impacts (Foreman and Wolke 1992; Forman and Alexander 1998; Trombulak and Frissell 2000).

Fire suppression in the Northern Rockies began in 1910, but did not have much effect until technology improved in the 1940s. The modern fire exclusion period included the suppression of lightning-ignited fires and the cessation of widespread burning by humans (accidental as well as intentional ignitions by Native Americans and early settlers, the latter of whom often used fire to clear land for agriculture and road-building). The magnitude of twentieth century fire sup-

PAGE 13

pression varies along an elevational gradient in Rocky Mountain ponderosa pine forests, with greater effects at low elevations (Steele et al. 1986; Brunelle et al. 2005).

Over the past two centuries, unusually favorable climatic conditions for tree regeneration, heavy and unregulated grazing by horses, cattle, and sheep, and the elimination of low-intensity surface fires (which previously served to thin dense stands of young trees) have all contributed to increased tree density in Rocky Mountain ponderosa pine forests and the decline in forest integrity (Borman 2005). The effects of these factors, as well as of road-building and the spread of invasive species, have been unevenly distributed across the landscape, with the greatest effects in the most accessible areas—that is, at low elevations and on gentle slopes. The increased stand densities we see today in previously logged ponderosa pine forests can be partially traced to recovery processes after logging (Kaufmann et al. 2000; Baker et al. 2006). Logging quickly removed many of the oldest and largest ponderosa pine trees, and resulting canopy gaps eventually filled in with young ponderosa pine and Douglas-fir. Historical photographs from the 1930s depict the removal of canopy trees, with the dense understory of ponderosa pine and Douglas-fir saplings purposefully left behind to promote regeneration (U.S. Forest Service 1937). This repeated selection cutting caused forests to develop an unnatural stand structure and progression (Hessburg et al. 2005, Hessburg et al. 2008).

The history of fire exclusion has also contributed to altered stand dynamics in ponderosa pine forests. On gentle slopes and at low elevations, where fires formerly were frequent but generally of low severity, fire suppression apparently has permitted a buildup of woody fuels, which in turn has led to increased fire severity. At higher elevations and on steeper slopes, fire suppression has not had the same effect because these forests were historically characterized by a mixed-severity fire regime (Veblen et al. 2000; Baker et al. 2006), and while

PHOTO BY GREG APLET/TWS

Euro-American settlement of the American West in the mid- to late 1800s also resulted in the establishment and expansion of many non-native, invasive plants (Hann et al. 1997; Hessburg and Agee 2003). The depletion of native bunchgrasses by widespread livestock grazing prompted the deliberate introduction of non-native forage species (Mack 1981, 1986), and railways and logging roads served as conduits for their spread. In addition, the seeds of non-native plants were often distributed within crop seeds and livestock forage. By the late nineteenth century, numerous invasive plants, such as bull and Canada thistles (Cirsium vulgare, Cirsium arvense), cheatgrass (Bromus tectorum), Dalmatian and yellow toadflax (Linaria dalmatica, Linaria vulgaris), and spotted knapweed (Centaurea esula), had become established (Wissmar et al. 1994; Langston 1995; Hann et al. 1997). The negative ecological impacts of invasive plants include altered fire regimes, loss of native species, reduced forage quality, reduced water availability, and altered nutrient cycles (Duncan et al. 2004).

A logged stand on the Boise National Forest, Idaho, depicting large old ponderosa pine with younger Douglas-fir and abundant brush.


fuels may have built up significantly on these sites (Franklin and Agee 2003) the fire regime likely remains predominantly mixed severity.

Because of the steep topography of these unroaded and unlogged areas — known as “roadless areas” — the fire regime was and remains mixed severity.

Both managed and unmanaged forests exposed to multiple fires over the past 100 years have lower stand densities than forests that experienced no fires during this same time period (Keeling et al. 2006). However, a history of logging may make forests more susceptible to high-severity fire than unmanaged forests, and plantation forests are uniquely susceptible to high-severity fire (Odion et al. 2004; Stephens and Moghaddas 2005; Thompson et al. 2007). It is highly likely that historical logging activities increased fuel accumulation beyond that caused by fire exclusion alone (Kaufmann et al. 2000). Most of the trees in unmanaged, fire-maintained forests are codominant, whereas trees in managed stands occur as dense ladder fuels. Naficy and Sala (2007) found that tree density in unburned, logged stands was approximately twice that in unburned, never-logged stands, and almost four times that in never-logged, fire-maintained stands. While fire exclusion increases stand density by promoting the growth of shade-tolerant trees, logging greatly compounds this effect. Relative to unlogged stands, logged stands exhibit a higher density of small trees and a higher density of small dead trees. This suggests that logging of forest stands for fuel reduction may actually create greater potential for highseverity fires in the future. On many western public lands, grazing by large numbers of sheep resulted in severe erosion and significant changes in vegetation composition, structure, and function, including the establishment of high tree densities, shifts in understory species composition, and reduced stand productivity and forest health (Borman 2005). The removal of native bunchgrass and pinegrass communities by intensive grazing pressure reduced competition for shrubs and seedlings, allowing more trees and shrubs to successfully establish (Pearson 1942; Rummell 1951; Madany and West 1983), though the density of the resulting understories is highly variable (Schubert 1974; Veblen and Lorenz 1986; Heidman 1988). In addition, grazing reduced the incidence of low-severity fire and this, combined with years of effective fire suppression further facilitated by the extensive road network, allowed tree seedlings to survive (Sherriff 2004) and grow into lethal ladder fuels (Agee and Skinner 2005). These processes combined to convert the forest from relatively open stands of large-diameter, old-growth trees to crowded stands of small trees. Grazing would have had its greatest impact in forests where fire was previously frequent (i.e., those with open, grassy understories), and these impacts would have been compounded by the fact that grazing was generally conducted on the same low-angle slopes as those that were targeted for logging.

Finally, the introduction and spread of non-native forage species and weeds resulting from widespread livestock grazing, road-building, and logging operations altered the structure, composition, and successional pathways of these forests at landscape and regional scales by causing the exclusion of native species and altering fire regimes and other natural disturbance patterns (Belsky and Blumenthal 1997). Once again, low-elevation forests on gentle slopes experienced the greatest degree of impact, in this case from invasive species which were introduced along roads and grazing allotments.

PAGE 15

The Restoration Imperative: A Call for Holistic Landscape Management

The combined effects of past land management activities have produced two distinctly different conditions within the dry montane forest of the Northern Rockies. In relatively accessible locations on slopes less than 35 percent, forests have been converted from open stands of large trees, maintained by frequent, low-severity fires, to dense forests of second growth, without sufficient understory to carry surface fire but with enough vertical and horizontal fuel continuity to sustain lethal crown fire.

PHOTO BY JOHN MCCARTHY/TWS

Elsewhere, where steep slopes prevented road-building and subsequent logging, large trees remain. There, grazing and fire suppression have resulted in some increase in the forest’s density and vertical fuel continuity, but have not fundamentally altered forest structure and function. Because of the steep topography of these unroaded and unlogged areas (known as “roadless areas”), the fire regime was and remains mixed severity. While the surface fire component has been missing for a century, expected fire behavior there is likely not far outside its historical range of variability, although continued fire

PHOTO BY JOHN MCCARTHY/TWS

These two photos were taken before and after a prescribed fire was used to reduce the understory density in a dry montane forest in the Deadwood Roadless Area, located in Idaho's Boise National Forest. The photo above, taken in 2005, shows a dry montane forest with young ponderosa pine and Douglas-fir in the understory. The photo to the left, taken in 2006, shows the stand structure after the fire burned through the understory and reduced the density of young ponderosa pine and Douglas-fir.


exclusion threatens to convert these forests from a mixed-severity regime to a lethal crown-fire regime.

Therefore, the Northern Rockies, it would seem, require two distinctly different approaches to ecological restoration. The low-elevation, roaded and logged forests may require thinning and prescribed fire to establish sufficiently open conditions to reintroduce natural fire and accelerate the development of oldgrowth structure. Roadless areas require a different approach, focused on the reintroduction of mixed-severity fire and the prevention of future catastrophic loss of existing old growth.

Such an approach to forest restoration would be justified even in a static climate; however, the looming challenge of climate change compels urgent action. The overall importance of climate and its influence on forest structure and ecosystem processes underscores the urgency of ecological restoration to mitigate ecological impacts of climate change in forests that have undergone substantial alterations due to past land uses (Westerling et al. 2006). Recent warming trends have contributed to an increased occurrence of wildfires in the western United States (Westerling et al. 2006). These trends are predicted to continue, raising the possibility that stand-replacing fires in a warmer climate could push forest ecosystems into a state of instability where they regenerate in a state not similar to their prior condition. This type of effect has been modeled for the Canadian boreal forest, demonstrating a shift in forest composition to a suite of species that have lower carbon storage potential (Bond-Lamberty et al. 2007). Although there are many uncertainties associated with climate change-induced shifts in fire regimes, moisture and temperature conditions, and growing season length, the potential impact of climate change on forest composition and structure further emphasizes the need to encourage the return of fire-resilient conditions to our western forest ecosystems, and calls for particularly rapid action in those forests where current conditions of structure and vegetation composition fall outside of the natural range of variability. In more mesic forests, climate is a more potent driver of fire frequency and behavior than fuel load, thus rendering fuel reduction treatments far less effective at influencing fire behavior (Agee and Skinner 2005; Noss et al. 2006). Restoring natural conditions and ecosystem processes to dry montane forests will likely increase their ability to adapt to, and survive, a changing climate regime. Conversely, leaving these forests in their current state will likely make them more susceptible to stand-replacing fires, potentially leading to slow recovery or establishment of an altered ecological state that is neither natural nor desirable. Forest restoration that maximizes diversity, minimizes species loss, and returns to natural fire regimes will yield the greatest resilience to the impacts of climate change (Noss 2001). To accomplish this, restoration strategies must take into account the histories and effects of past management activities at different sites in the Rocky Mountain ponderosa pine environment.

PAGE 17

Opportunities for Dry Montane Forest Restoration in the Northern Rocky Mountains

To investigate how the two approaches to dry montane forest restoration described above might play out on the ground, we conducted a spatial analysis of ponderosa pine-dominated forests across public and private lands in the states of Idaho and Montana. These two states represent the bulk of the Northern Rockies region, and are also the states where good data exist on the distribution of dry montane forests. Our analysis shows where dry montane forest is found, where it may need to be managed to minimize the wildfire threat to communities, and where forests in need of restoration can be easily accessed for treatments. As seen below, our analysis demonstrates the need for two distinct approaches to restoration depending on the management history of the particular forest. There are many areas in these two states that have been logged and may be appropriate for thinning and prescribed fire to restore the forest mosaic. Where thinning is appropriate, it may contribute to the wood products sector of the economy. At the same time, however, thinning would be inappropriate for forests in roadless areas and wilderness. These forests need only the return of fire to sustain their ecological integrity.

Because communities have so frequently developed at the lower elevations of mountain ranges in the West (leaving ponderosa pine very often concentrated within just a few miles of urban and rural towns), and because restoration of dry montane forests will require the reintroduction of fire, restoration opportunities must be assessed in the context of surrounding communities. Using housing density data from the U.S. Census Bureau and land ownership data acquired from the states, we classified the dry montane forest of Idaho and Montana according to a three-zone approach to fire management developed by The Wilderness Society (Wilmer and Aplet 2005; Aplet and Wilmer 2006). According to this system, the area closest to communities, which must be managed for community protection from wildfire, is known as the “Community Fire Planning Zone” (CFPZ). Here, communities must plan the necessary infrastructure, conduct homeowner education, and facilitate fuel reduction to minimize the probability of structure loss during wildfires. To define this zone for our study area, we first identified all census blocks in which private land housing density exceeded one house per 40 acres (following the density threshold for wildland-urban interface communities provided in the January 4, 2001 Federal Register). Next, we buffered the private land in these census blocks by one-half mile to represent the area surrounding communities in which vegetation may need to be modified for community protection.

In the remotest parts of the landscape, fire does not threaten communities, and natural fires can be managed for their ecosystem benefits. We call this part of the landscape the “Fire Use Emphasis Zone” (FUEZ) to emphasize the opportunities for using fire alone to achieve ecological restoration. To define the FUEZ for our study area, we had to identify a distance from the CFPZ beyond which it may be acceptable to allow a natural ignition to burn. In the absence of any clear guidance on such a distance, we selected five miles. Admittedly, this is an arbitrary distance; in practice, selected distances will be publicly negotiated among community members, land managers, and the public through the land


management planning process and other agreements. However, five miles provides a reasonable starting point for analysis.

In between these two zones, in the five-mile radius between the Community Fire Planning Zone and the Fire Use Emphasis Zone, is an area that is close enough to communities that wildfires will not be welcome but that is far enough away that the condition of the vegetation poses no substantial threat to communities. We call this area the “Wildfire Resilience Zone” (WRZ), because wildfires will generally be suppressed, but vegetation should be managed (with both thinning and prescribed fire) to be ecologically resilient to wildfires that do escape “initial attack.” The WRZ, because of its proximity to communities and the presence of roads that have facilitated effective fire suppression, has suffered the most from fire exclusion in the past. Thus, despite the nearness of forests in the WRZ to human settlement, restoration activities here have the potential both to improve landscape health and provide critical wildlife habitat. Over these three zones, we laid information on land cover/vegetation types and land ownership. Land cover was broken down by major physiognomic groups, including urban, agricultural, grassland, shrubland, and forest. Forest types were further classified as “dry montane forest” and “other forest.” In Idaho, dry montane forests include Ponderosa Pine Forest and Mixed Xeric Forest, which includes dry forests dominated by ponderosa pine and Douglas-fir. In Montana, which employs a different vegetation classification scheme, we identified Ponderosa Pine Forest, Mixed Xeric Forest, and Low Density Xeric Forest. In addition to these vegetation classes, we looked at major land ownership classes, including private, state, tribal, and major federal agencies.

Finally, we calculated acreages of unroaded dry montane forests using national park, wilderness, and national forest inventoried roadless area land jurisdictions. These areas have not experienced the same logging and road-building regime as other federal lands, and thus represent areas where restoration using fire alone is appropriate. This may be either natural fire or prescribed fire, which does not require road access in order to carry out. By contrast, areas that have been previously roaded are likely to require thinning as part of their restoration plan.

FIGURE 3.

Our analysis indicates that there are approximately 6.6 million acres of dry montane forest in the states of Idaho and Montana, representing about one-sixth (16.5 percent) of the approximately 40 million acres of forest in the two states (Figure 3). About 11 percent (717,000 acres) of the dry montane forest exists in national parks, wilderness, and national forest roadless areas (433,000 acres in Idaho and 284,000 acres in Montana) (Figure 4). This forest is likely to be unaltered by past logging n Dry Montane Forest activity and could be restored through reintron Other Forest duction of fire alone. The remaining 5.8 million acres of dry forest exists on roaded federal lands (32 percent) or on non-federal lands (68 percent) (Figure 4). It is possible that some of this acreage, especially the remote non-federal lands in eastern Montana, also contains large trees, an Montana open forest structure, and areas where logging

Dry Montane and Other Forest Acreage in the Northern Rockies 18,000,000 16,000,000 14,000,000

Acres

12,000,000 10,000,000 8,000,000 6,000,000 4,000,000 2,000,000 0 Idaho

PAGE 19

and road-building did not occur. However, it is clear that most of the dry montane forests of the Northern Rockies are likely to have suffered effects from past land management activities that make them vulnerable to fire, and may require thinning prior to the reintroduction of fire.

FIGURE 4.

Ownership and Status of Dry Montane Forest in the Northern Rockies

3,500,000 3,000,000 2,500,000

Acres

When these patterns are considered in 2,000,000 the context of the three-zone fire management system, additional conse1,500,000 quences for restoration emerge. The 1,000,000 majority of dry montane forest in Idaho and Montana occurs within approximate500,000 ly five miles of communities (Figure 5). 0 In Idaho, the CFPZ contains more than 320,000 acres of dry montane forest (less than 3 percent roadless), and the FIGURE 5. WRZ encompasses almost 1.5 million acres (16 percent roadless) of this forest (Figure 5a). In Montana, about 330,000 acres of dry montane forest occur in the a. Idaho CFPZ (less than 3 percent roadless), and 800,000 more than 2.2 million acres occur in the 700,000 WRZ (8 percent roadless) (Figure 5b). In 600,000 the portion of these lands that are roadless, fire alone may be used to reduce 500,000 fuels and achieve restoration. In the 400,000 roaded portion, thinning may be combined with fire to achieve restoration 300,000 goals. All told, the two states contain 200,000 more than 4.3 million acres of dry mon100,000 tane forest where fuel treatment should be emphasized (either for community 0 protection in the CFPZ, or ecological restoration in the WRZ). Only 10.1 perb. Montana cent of this acreage is roadless. The 1,600,000 remaining 3.9 million acres of dry montane forest represent a sizable opportuni1,400,000 ty for landscape-scale restoration outside 1,200,000 of parks, wilderness, and roadless areas 1,000,000 where thinning may be appropriate. In addition to these priorities for dry mon800,000 tane forest restoration, the CFPZ contains 600,000 other kinds of forests that may also need 400,000 to be thinned to provide community protection from fire. Forest types other than 200,000 dry montane in the CFPZ amount to 0 740,000 acres in Idaho and 940,000 acres in Montana.

n Roadless Federal n Roaded Federal n Non-Federal

Idaho

Montana

Dry Montane Forest in Three Fire-Management Zones in Idaho (a) and Montana (b)

Acres


CFPZ

WRZ

FUEZ

WRZ

FUEZ

Acres


CFPZ


In both states, about two-thirds of the CFPZ and three-quarters of the dry montane forest in the CFPZ is under non-federal ownership, indicating that community protection from fire in the Northern Rockies represents both a fuel treatment challenge for private landowners and an opportunity for private land restoration. In addition, more than half of the dry montane forest in the WRZ occurs on non-federal land. Although The Wilderness Society focuses on activities within federal lands, we recognize the importance of natural conditions and processes across the broader landscape. Private lands managed in a manner that restores and sustains resilience to wildfire without the loss of old-growth trees will protect these important genetic and ecological resources and further reduce the potential for movement of large-scale stand-replacing fires across ownership boundaries.

Beyond five and a half miles from communities, in the FUEZ, there is relatively little opportunity for dry montane forest restoration in Idaho; roaded federal and non-federal lands each contain less than 100,000 acres of this forest, and federal roadless lands contain less than 200,000 acres (Figure 5a). In Montana,

FIGURE 6.

Dry Montane Forests in Montana and Idaho

Coeur d’Alene • Missoula • Helena • Billings •

Boise •

0

Community Fire Planning Zone Dry Montane Wildfire Resilience Zone Forests Fire Use Emphasis Zone include ponderosa Federal lands pine and dry mixed conifer 50 100 Miles

PAGE 21

where almost 2 million acres of dry montane forest exist in the FUEZ, the opportunity for restoration may be substantial, especially on private land (Figure 5b) In these remote areas, some thinning may be appropriate where past logging has altered forest structure, but opportunities should also be sought to use natural fire to achieve forest restoration goals under the right conditions.

In summary, the states of Idaho and Montana contain 6.6 million acres of dry montane forest, only about 11 percent of which exists on roadless federal lands where forest structure should be restored through the use of fire alone (Figure 6 and 7). The remaining 5.9 million acres of dry montane forest have likely been affected by past management activities such that restoration may require thinning in addition to fire. Of these lands, two-thirds occur within five and a half miles of communities, where active thinning is most appropriate. The remainder, mostly remote non-federal lands in eastern Montana, may require some thinning to address past management effects, but may also present opportunities to restore forest structure through the use of natural fire.

FIGURE 7.

Dry Montane Forests in Montana and Idaho

Coeur d’Alene • Missoula • Helena • Billings •

Roadless lands Roaded Federal lands Non-federal lands

Boise •

0

Federal lands National Parks, Wilderness, and Roadless Areas 50 100 Miles

Dry Montane Forests include ponderosa pine and dry mixed conifer


Elements of a Successful Restoration Program

A central tenet of ecological restoration holds that the conditions and processes that sustained ecosystems historically can sustain them again in the future if restored through management. Under the increased fire activity expected due to climate change, restoring historical, fire-resilient forest structure takes on even greater importance. Restoration activities include both active and passive management approaches. Active management strategies apply cultural operations and forest management strategies such as timber harvest, tree planting, thinning, prescribed burning, fertilization, grazing, weed control, improving wildlife habitat, stream channel reconstruction, erosion control, decommissioning of roads, trail and road maintenance and construction, and recreation resource maintenance and improvement. Passive management strategies allow natural structure, composition, and function to operate and exist in the landscape, such as by not suppressing wildfire, allowing natural plant succession to proceed on abandoned roads, and allowing stream channels to be restructured by flood events. As described above in the holistic vision section, different combinations of these two approaches will be relevant for different areas of the Northern Rocky Mountain ponderosa pine ecosystem. Whether active or passive management strategies are employed, restoration forestry must attempt to recreate natural processes through manipulation of forest structure, composition, and function, and these modifications must be conducted in a manner that maximizes landscape connectivity between restored and natural landscapes. This requires attention to several key elements, detailed below.

Use Adaptive Management

FIGURE 8.

Adaptive management is an approach that considers each management action as an experiment that will inform future activities (Shindler et al. 1999). An increasing number of federal forest management plans include the concept of adaptive management, but current management often falls short of this goal (Bormann et al. 2007). Adaptive management is a circular, fluid, and iterative process involving a continuum of evaluation, planning, action, and monitoring (Figure 8). Once initiated, the adaptive management approach requires continual monitoring, evaluation, and, potentially, reorientation, if the ecosystem’s trajectory is not meeting objectives. Each element of this process is fundamental to the success of the approach, and ignoring or short-changing any one element greatly reduces the likelihood of success. There must be an acknowledgement of our inability to predict future conditions (due to a changing climate, our limited Planning understanding of forest succession patterns, and introduction of new variables such as exotic species), as well as our limited knowledge of the efficacy of different restoration treatments. These Implementation uncertainties mean that the flexibility to change with improving knowledge is a crucial element of restoration planning.

The Adaptive Management Process s

s

s

s

Evaluation

Monitoring

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Set Priorities

Priority setting must take into account the landscape-scale variability characteristic of the Rocky Mountain dry montane forest, and ensure that all elements of the forest mosaic are included. For example, some wildlife species require dense stands of ponderosa pine, while other species prefer habitat with open stands and large-diameter trees. Identifying where these habitats currently occur and where certain habitat conditions are currently lacking at the landscape scale can help identify where restoration efforts should be focused.

In addition, planners must take into account the risks created by past management activities. Treatments such as thinning should be focused on stands that have retained large-diameter trees but are threatened by fuel accumulation as a result of logging and fire exclusion (Kolb et al. 2007). Stands containing oldgrowth pines, larch, or Douglas-fir should also garner priority for restoration in order to protect these trees from loss to stand-replacing wildfires. Landscapescale analysis can also indicate where other important genetic resources exist and need to be protected. Lower-elevation forests that remain unroaded and unlogged are relatively rare in the Northern Rockies, and thus will have a high value for conserving biodiversity (Noss and Cooperrider 1994; Strittholt and DellaSala 2001; Crist et al. 2005). They can serve as refugia for sensitive terrestrial and aquatic species, have lower rates of invasions of non-native species, and provide reference conditions for understanding natural ecosystem processes. Restoration of these sites should focus on the reintroduction of mixed-severity fire and protection from activities that may cause degradation or loss of existing old growth.

By evaluating ponderosa pine and Douglas-fir forests through a series of landscape and ecological filters, we can begin to narrow down where restoration should be a high priority and identify which methods can be used to achieve the goal of restoration. Current U.S. Forest Service efforts to prioritize forest restoration at a region-wide scale (e.g., Values at Risk in Region 1, Integrated Restoration and Protection Strategy in the Northern Region, Southwest Idaho Ecogroup Aquatic Restoration Strategy {see U.S. Forest Service 2003}) use a combination of GIS data layers such as road density, municipal watersheds, fisheries, fire history, and winter range elk habitat (see, for example, http://fsweb.r1.fs.fed.us/gis/R1_values_at_risk_strategy/). This type of coarse-

PHOTO BY GREG APLET/TWS

To achieve ecological restoration ,land managers should begin by developing a restoration plan that identifies the most important pieces of the landscape for restoration. Planning at broad spatial scales can facilitate the development of multi-objective restoration strategies where different needs can be met. Projects should be selected based on their potential for success and their potential for impact in a larger ecosystem context.

This ridge above Long Creek, in Idaho’s Boise National Forest, is a good candidate for restoration. It displays an open forest of large ponderosa pine, relatively gentle slopes, and good road access.


filter approach can provide the starting point for a closer look, but the fuller analysis must include additional crucial aspects such as landscape history, location of important genetic resources, landscape connectivity, location of endangered, threatened, and sensitive species, and future land use plans.

take into account the

Finally, the relationship to existing natural areas can also help guide restoration priorities. For example, ponderosa pine and mixed ponderosa pine/Douglas-fir forests that are located in close proximity to or between roadless or protected areas may be good candidates for restoration, as they will help create a larger effective size of the roadless and/or protected area as well as increase connectivity between these areas. Larger patches will protect more species and more individuals, provide more habitat for species with large home ranges and those requiring large contiguous tracts of ponderosa pine habitat, restore more intact ecosystem processes, and increase landscape connectivity within and amongst these patches. By contrast, random or non-prioritized restoration of individual stands, streams, or roads within a watershed may result in the formation of “islands” of natural process within a larger altered landscape, and reduce the potential for success.

landscape-scale

Take the Long View

Priority setting must

variability characteristic of the Rocky Mountain dry montane forest and ensure that all elements of the forest mosaic are included.

It must be emphasized that restoration in the absolute sense will not be achieved in our lifetimes; instead, restoration efforts place ecosystems on a more natural trajectory. Therefore, restoration strategies must be planned with an understanding of short- and long-term temporal variables (diurnal and seasonal cycles and centennial and millennial climatic cycles).

Sustainable forest management may require that we consider all of these factors within the context of a several-hundred-year perspective. Although ten-year forest plans are appropriate for administrative purposes, this time frame is far too short to effectively address sustainable forest management. Restoration efforts may require tens to hundreds of years prior to establishment of natural nutrient, energy, carbon, and fire cycles. Forest harvests conducted in a context of future “resting” of the landscape prior to re-entry might mean planning over a 250-300 year horizon. A combination of short-range and long-range planning will be necessary to accomplish forest restoration objectives. Short-range planning will allow for projects to be placed on the landscape, while long-range planning will allow forest managers to truly consider sustainable forest management. Restoration efforts must also be considered in the context of climate change. Recent temperature increases and precipitation decreases in the Rocky Mountain West are clearly leading to a shift in fire occurrence and severity on the landscape (Westerling et al. 2006). Climate change may also result in shifts in understory species composition and thus fine fuel accumulation on the forest floor. These changes in fuel loading and character may change fire patterns and severity, which in turn will affect the success of the restoration effort. Overall, it is not clear how restoration efforts will be influenced by climate change, but restoring forests as close as possible to their natural conditions and ecosystem processes will likely increase their ability to adapt to and survive a changing climate regime.

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Consider the Whole Ecosystem

When conducting thinning operations, prescribed fire, or other silvicultural practices as part of forest restoration programs, there are a number of issues that need to be considered to avoid negative environmental effects. An assessment of these potential impacts is required in order to evaluate the potential benefits of a proposed restoration plan. Silvicultural treatments are designed to produce a change in overstory structure and composition, but they also have a less obvious though equally important influence on understory composition, belowground processes, and avian and mammalian habitat and food resources over both the short term and long term.

Effects on Wildlife. Restoration activities that interrupt certain ecological processes in the short term may yet improve habitat conditions for many wildlife species in the long term. For example, restoration that combines burning and thinning produces highly variable effects on populations of small mammals depending on the species and pretreatment conditions, but overall, the increased habitat complexity produced by these treatments has a positive effect, especially over time (Carey and Harrington 2001). By contrast, removing large-diameter trees and snags and altering the natural “clumpy” distribution of trees and snags can have a significant negative impact on many birds, including white-headed woodpeckers, flammulated owls, Lewis’s woodpeckers and pygmy nuthatches (Hutto 2006). Restoration strategies should strive to maintain the diverse, patchy landscape characteristic of the dry montane forest (Dixon 1995; Saab et al. 2002).

Effects of Non-Native Species. Recent literature has clearly demonstrated that active restoration treatments increase the numbers of non-native plant species present (Griffis et al. 2001; Weink et al. 2004; Fulé et al. 2005; Dodson and Fiedler 2006), due to the “opening” of the stand by thinning, prescribed fire, or especially the two in combination. Research in Montana associated with the Fire and Fire Surrogates Study demonstrated that the combination of thinning and prescribed fire increased the cover of non-native species by a factor of four compared to thinning alone or burning alone, and by more than a factor of seven compared to untreated control (Dodson and Fiedler 2006). This response may have been the result of the frequency (two separate disturbance events in two consecutive years) as well as the intensity of the treatment, with many trees removed by the thinning operation and additional trees killed by the prescribed fire (Dodson and Fiedler 2006). The newly opened stand along with soil disturbance created conditions well suited to invasive species establishment and distribution. Such impacts of forest restoration activities must be evaluated for their potential to persist over the long term, and must also be weighed against longterm benefits, assuming that the restoration objectives are ultimately achieved. Effects on Soil and Soil Processes. Restoration activities have the potential to alter soil function through decreased infiltration rates, decreased organic matter inputs, nutrient loss, increased soil erosion, and decreased microbial diversity and activity. These changes may produce long-term degradation of forest function and productivity, and thus failure of restoration objectives. Tractors used in timber harvest cause soil exposure and compaction, and timber export results in removal of carbon and nutrients from the forest environment. However, soil degradation can be greatly minimized by taking simple precautions during forest restoration activities and avoiding highly sensitive land-


scapes. Timber harvest using ground-based equipment should only be conducted under dry or frozen soil conditions, and only on slopes of less than 30 percent. Compaction of soils by harvesting on wet or moist soils may take tens of years to recover, thus reducing the potential for moisture recharge of surface soils (Brais and Camiré 1998).

Forest roads and timber harvest activities may accelerate sediment delivery to rivers and streams, greatly reducing water quality and endangering aquatic organisms (Sharpe 1975; Trombulak and Frissell 2000). Mass wasting and road failure events are often associated with unstable soils, undercut slopes, and extreme weather conditions, and can produce large, acute sediment delivery events. In contrast, soil compaction and exposure within a timber harvest unit is more likely to yield chronic, low-level increases in sediment delivery to neighboring water bodies. Soil exposure and compaction accelerate soil erosion, which is a particular problem on sites with steep slopes, highly erodible soils, and frequent high-intensity rainfall events (Burroughs and King 1989).

Timber harvest in combination with prescribed fire almost always results in a net loss of carbon and nitrogen reserves associated with the soil organic horizon (Gundale et al. 2005). Ground disturbance from timber harvest commonly mixes the organic horizon material into surface mineral soil, resulting in a temporary increase in mineral soil organic matter but also an increase in soil respiration and decomposition rates. This, combined with the increased surface temperatures associated with the opening of the forest canopy, can result in increased degradation of soil organic matter and a net loss of soil carbon. Reintroduction of fire as part of forest restoration may increase nutrient turnover rates (DeLuca and Zouhar 2000; Gundale et al. 2005; Hart et al. 2005), bringing them closer to pre-management conditions, but excessive use of fire may also lead to reductions in total nutrient capital (Hart et al. 2005; Smithwick et al. 2005, Keeling et al. 2006). Many of these changes in soil properties (loss of forest floor carbon, stimulation of nutrient cycling, and changes in soil acidity) are also changes that could be expected after wildfire events. Thus, nutrient and carbon dynamics must be considered within a long-term context that takes into account both the effects of natural disturbances on nutrient cycling, and the unnatural export of nutrients offsite that may be part of restoration activities. Charcoal is a significant byproduct of fire events and one of the only long-term legacies of fire (DeLuca et al. 2006; DeLuca and Aplet 2008). Charcoal is a desirable byproduct of fire, increasing nutrient turnover rates, soil water-holding capacity, cation exchange capacity, and soil warming under sun exposure (due to the darker soil color). Thinning of forests without the reintroduction of fire alters stand structure without depositing this potentially important driver of soil processes. Furthermore, charcoal represents a much more persistent form of carbon than forest biomass or even soil organic matter, and so charcoal formation may result in long-term storage of forest carbon not paralleled in unburned forests (DeLuca and Aplet 2008).

Insect and Disease Considerations. Forest restoration programs must take into account the fact that other natural disturbances such as insect and pathogen outbreaks are an important part of the ecological function of ponderosa pine forests. Such outbreaks, even ones that kill trees over thousands of acres, are natural events and have been occurring for thousands of years

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throughout the Northern Rockies (Swetnam and Lynch 1993). In recent years, the combination of drought conditions and hot summers (Westerling et al. 2006) has probably stressed trees and could make ponderosa pine, Douglas-fir, and western larch more susceptible to insects and pathogens. In addition, the homogeneous forest conditions that have been created in low-elevation forests by past forest management activities and fire suppression efforts have increased insect- and disease-induced mortality in recent years (Hemstrom 2001). Where appropriate, thinning has the potential to improve stand health by reducing basal area and thus tree-to-tree competition, in turn increasing the ability of individual trees to defend against insect attack (Fettig et al. 2007). Reducing ponderosa pine stand density has been shown to reduce the number of beetle attacks on individual trees if conducted sufficiently far in advance of the outbreak (Schmid and Mata 2005; Kolb et al. 2007). However, in such cases it is critical that slash (logging debris) created by the thinning operation be treated immediately to prevent the buildup of bark beetles (e.g., Ips spp.) that can move out of the slash to kill live trees (Graham and Knight 1965). Furthermore, the consequences of careless mechanical thinning activities, such as soil compaction and root damage, may increase the susceptibility of forest stands to subcortical insects and root pathogens (Fettig et al. 2007).

Be Mindful of Socioeconomic Context

Successful restoration projects must also take into consideration the existing infrastructure and economic base of surrounding communities. Restoration efforts based solely on ecological outcomes with no consideration of long-term social

TABLE 3.

Examples of Ecological Variables to be Addressed by Restoration Monitoring Type of Monitoring

Number of Sites/ Data Quality

Ecosystem Health Variables

Aquatic Integrity Variables

Fire Hazard Variables

Biodiversity Indicators

Multi-Party Monitoring

Most sites/Limited data of moderate quality

Tree mortality Disease Soil organic matter Soil compaction

Aquatic invertebrates Water temperature

Surface fuel load Fuel moisture Ladder fuels

Invasive species Tree size, age, diameter

Intensive Sample Monitoring

Few sites/ Extensive data of high quality

Tree mortality/ Cause of death Forest structure Coarse woody debris Regeneration Disease indices Trail widening Soil erosion Nutrient availability Infiltration rates Soil organic matter

Aquatic invertebrates Water temperature Water chemistry Fish counts Turbidity

Surface fuel load Crown base Fuel moisture Live fuels

Vegetation composition Endangered species Tree size, age, diameter Distribution change Indicator species— wildlife or plant

Spatial Monitoring

All sites/Limited data amount and quality

Tree mortality Stress

Turbidity Temperature

Density, foliar moisture

Change in species composition, invasive plants


impacts or economic feasibility are likely to fail. Although ecological processes are the primary focus of the proposed action, public support and economic benefit are necessary for restoration programs to be effective. Ultimately, these projects should lead to increased quality of life for nearby communities, which will then become more connected to and invested in the health of the landscape. Restoration projects should be implemented with ample opportunity for public comment and support from all stakeholders. This approach, which is generally applied in federal stewardship contracting, has been highly successful at engaging local citizens, contractors, and conservationists. Having broad support for a given project will greatly increase the potential for success and greatly reduce the potential for legal battles and negative discourse. Changes to the project that emerge from the monitoring process should also receive public comment and evaluation.

Monitor Restoration Success

Restoring a forest’s historical structure and function is a long-term process that only begins with reintroducing fire, thinning, pruning, road removal, and other mechanical actions. Because restoration is a new enterprise, monitoring must be conducted to assess impacts and guide future improvement. Monitoring is absolutely fundamental to the cyclical process of adaptive management (Shindler et al. 1999). In addition, changing climatic conditions emphasize the need for monitoring of restored forests to allow for the modification of treatments within the region should existing restoration tools or approaches lead to unexpected outcomes.

Unfortunately, very few comprehensive monitoring efforts have taken place to date, limiting the ability to assess the efficacy of forest restoration (US-GAO 2004). Most of the monitoring programs that have been undertaken could be termed implementation monitoring (Block et al. 2001), or compliance monitoring, which assesses whether or not a management action has been carried out as designed. What is needed, however, is effectiveness monitoring, which evaluates whether or not a management action has achieved its ultimate objective (Block et al. 2001). There are few examples of any extensive effectiveness monitoring conducted in relation to forest restoration efforts on federal lands over the past 20 years. Effectiveness monitoring has four fundamental components (Block et al. 2001): 1. A clear objective based on historical reference conditions and an appropriate vision of how that condition will exist on the landscape. 2. Selection of a meaningful, quantifiable, and realistic set of response variables that are easily observed and are repeatable. The Fire and Fire Surrogates Study (http://www.fs.fed.us/ffs/) identified a number of appropriate response variables that are sensitive to forest restoration treatments. We propose monitoring response variables within each of four fundamental ecosystem attributes: (a) Ecosystem health; (b) Aquatic integrity; (c) Fire hazard; and (d) Biodiversity. Effective multiparty monitoring is much more likely to be successful if the monitoring program provides meaningful incentives that lure participants back and encourage them to take the process seriously. (DeLuca et al. 2008). (See table 2 3, page 27).

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3. Adequate funding for the completion of all monitoring tasks. The up-front costs of monitoring are commonly viewed as an impediment, but in fact monitoring programs associated with major environmental policy implementations have proven to account for a fraction of the total cost and yield extremely important information (Lovett et al. 2007). 4. A dedicated system for recording, analysis, synthesis, and reporting of monitoring data.

Forest restoration activities might be effectively monitored using one of the following approaches (DeLuca et al. 2008): (1) Multi-party monitoring conducted by a collective of stakeholders including citizens, conservation groups, timber interests, and agency personnel; (2) Intensive sample monitoring, meaning detailed, federally supported ecosystem and socioeconomic monitoring of a limited number of representative restoration sites; or (3) Spatial monitoring, analysis using remote sensing of ecosystem response to restoration activities. Table 3 provides examples of the ecological variables that could be monitored under each of these three approaches. Although spatial monitoring could potentially be conducted broadly across landscapes, multi-party monitoring or intensive federal monitoring cannot reasonably be conducted on every forest restoration project conducted throughout the Rocky Mountain West. Where multi-party monitoring efforts are undertaken, non-profit organizations could play a key role in their design, organization, and implementation. Teams could be built primarily depending on “citizen scientists;” however, these types of efforts require support to ensure data quality and consistent results.


Conclusions

Dry montane forests in the northern Rocky Mountains have been greatly altered by management activities over the past two centuries, including logging, grazing, road-building, and broad-scale fire exclusion. These activities have resulted in dense growth of more shade-tolerant species and a greater percentage of small diameter trees than was typical of the same forests in the years before European settlement of the American West. Historically, the majority of these forests existed in disequilibrium conditions, subject to periodic fires that burned with varied severity and frequency and resulted in a mosaic of stand structures across the landscape. However, changes due to past management activities, as well as recent and ongoing climate change, have made these forests more vulnerable to severe, stand-replacing burns. In the past, forest restoration efforts aimed at reconstructing historical stand structure and function in the Northern Rockies have greatly depended upon examples from the southwestern United States, where stands of dry, pure ponderosa pine experienced recurrent, sub-lethal fires on a relatively short, consistent interval. A different, mixed-severity fire regime is characteristic of forests in the Northern Rockies, and forest restoration efforts there must be designed with these ecological differences in mind.

Restoration efforts must also take into account how the current conditions differ at different sites in the Northern Rockies. Lower-elevation forests that have been more severely affected by past management activities may need a combination of thinning and prescribed fire to bring stand structure closer to the historical condition before natural fire can be reintroduced. Higher-elevation, less accessible forests that have not been changed as significantly by past management activities may need only the return of natural fire to sustain their ecological function.

Humility is a fundamental ingredient of an effective forest restoration strategy. We must acknowledge our limited knowledge of future conditions (due to a changing climate, long-term forest succession patterns, and the introduction of new variables such as invasive species) and our limited knowledge of the efficacy of restoration treatments. Effective restoration management must be designed with an inherent flexibility that allows land management agencies and landowners to respond to improved knowledge in these areas by altering management strategies and restoration treatments if necessary. This type of flexibility requires a clear and consistent monitoring of ecosystem response to restoration activities, a central assumption of adaptive management. Achieving restoration of Rocky Mountain dry montane forest ecosystems, or at least setting the forests on a more natural trajectory, will require that land managers adequately analyze historical patterns within specific regions, use timber harvest as a stand treatment only where appropriate, reintroduce fire as a natural process whenever possible, and apply a flexible, adaptive approach to restoration. Under these conditions, forest restoration will have the potential to maintain wildlife populations, build resiliance to a changing climate, and create the sustainable conditions desired by land managers, scientists, and the people who will continue to enjoy these forests in the future.

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COVER PHOTOS: Top: Boise National Forest, Idaho. Photo by John McCarthy, TWS Right: Boise National Forest, Idaho. Photo by John McCarthy, TWS

Michele R. Crist Ecologist [email protected] The Wilderness Society 950 West Bannock, Suite 605 Boise, ID 83702 208-343-8153 Thomas H. DeLuca, Ph.D. Senior Ecologist [email protected] The Wilderness Society 503 West Mendenhall Bozeman, MT 59715 406-586-1600 Bo W ilmer Landscape Ecologist [email protected] The Wilderness Society 950 West Bannock, Suite 605 Boise, ID 83702 208-343-8153

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Gregor y H. Aplet, Ph.D. Director of Ecology [email protected] The Wilderness Society 1660 Wynkoop Street, Suite 850 Denver, CO 80202 303-650-5818