This file was created by scanning the printed publication. 

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Technical Editors

Robert T. Meurisse (Retired) is formerly Regional Soil Scientist, USDA Forest

Service, Pacific Northwest Region, Portland, Oregon.

William G. Ypsilantis is District Soil Scientist, USDI, Bureau of Land

Management, Coeur d'Alene, Idaho.

Cathy Seybold, is Soil Scientist, USDA Natural Resources Conservation

Service, Soil Quality Institute, Oregon State University, Corvallis, Oregon.

Papers were provided in camera-ready form for printing by the authors, who are therefore

responsible for the content and accuracy. Opinions expressed may not necessarily reflect

the position of the U.S. Department of Agriculture.

The use of trade, firm, or corporation names in this publication is for the information and

convenience of the reader. Such use does not constitute an official endorsement or

approval by the U.S. Department of Agriculture of any product or service to the exclusion

of others that may be suitable.
Proceedings: Pacific Northwest Forest

& Rangeland Soil Organism Symposium

Robert T. Meurisse, William G. Ypsilantis, and Cathy Seybold, Technical

Editors

Pacific Northwest Forest and Rangeland Soil Organism Symposium:

Organism Functions and Processes, Management Effects on Organisms and

Processes, and Role of Soil Organisms in Restoration

LaSells Stewart Center, Oregon State University,

Corvallis, Oregon

March 17-19, 1998

U.S. Department of Agriculture

Forest Service

Pacific Northwest Research Station

Portland, Oregon

General Technical Report PNW-GTR-461

April 1999
ABSTRACT

Meurisse, Robert T.; Ypsilantis, William G.; Seybold, Cathy, tech. eds.

1999. Proceedings, Pacific Northwest forest and rangeland soil organism

symposium; 1998 March 17-19; Corvallis, OR. Gen. Tech. Rep. PNWGTR-461. Portland, OR: U.S. Department of Agriculture, Forest Service,

Pacific Northwest Research Station. 215 p.

Soil organisms have become a focus of attention for addressing issues of soil

quality and health, and ecosystem sustainability. Land managers are

challenged to ensure that their actions are beneficial to belowground

organisms and processes in the long-term. Research about soil organisms,

their populations, roles, and management effects is fragmented and often

esoteric. The diversity of soil organisms is tremendous. The principal

organisms include bacteria, actinomycetes, fungi, arthropods, nematodes,

worms, insects, and mammals. In order to provide a "state-of-the-science"

about soil organisms related to management, a symposium was convened to

address soil organism functions and processes, management effects on soil

organisms and processes, and describe some roles of soil organisms in

restoration and applications for land management. The focus of the

symposium was on Pacific Northwest forest and rangeland ecosystems.

Keywords: Soil organisms, soil quality, nutrient cycling, nitrogen

transformations, foodwebs, ecosystem integrity, soil resilience, ecosystem

resilience, bacteria, fungi, arthropods, root diseases, soil crusts, disturbance,

interactions
Contents

Opening Remarks

Opening Remarks 1

William G. Ypsilantis

Assessment of Below-ground Organism Functions and Processes

A Non-Value Based Framework for Assessing Ecosystem Integrity 3

Kristina A. Vogt, Daniel J. Vogt, Paul Boon, Anna Fanzeres, Philip Wargo, Peter A. Palmiotto, Bruce Larson,

Jennifer L. O'Hara, Torel Petel-Weynand, Eva Cuadrado, and Joyce Berry

Soil Quality and Health-Some Applications to Ecosystem Health and Sustainability 21

Robert T. Meurisse

Soil-dwelling Arthropods: Their Diversity and Functional Roles 33

Andrew R. Moldenke

Biological Diversity and Ecosystem Functions of Forest Soil Fungi: Management Implications 45

Randy Molina, Thomas O'Dell, Susie Dunham, and David Pilz

Soil Bacteria: A Dynamic Pool of Soil Organic Matter and Catalysts of Key Belowground Processes 59

Jeanette M. Norton

Decomposition of Woody Material: Nutrient Dynamics, Invertebrate/Fungi Relationships and Management

in Northwest Forests 68

Robert L. Edmonds and James L. Marra

Invertebrate Biodiversity, Forestry and Emulation of Natural Disturbance: A Down-to-earth Perspective 80

John R. Spence, Christopher M. Buddie, Kamal J.K. Gandhi, David W. Langor, W. Jan A. Volney, H.E. James

Hammond, and Greg R. Pohl

Management Effects on Soil Organisms and Processes

Soil Organisms and Rangeland Soil Hydrological Functions 91

Jeffrey E. Herrick

Effects of Disturbance on Microbial Activity and N-Cycling in Forest and Shrubland Ecosystems 101

John M. Stark and Stephen C. Hart

Nitrogen Transformations and Management Interactions in Forests 106

David D. Myrold

Wildfire vs. Clearcutting: Impacts on Ectomycorrhizal and Decomposer Fungi 114

S. Visser and D. Parkinson

Decomposition and Nitrogen Mineralization in Forests of British Columbia: Effects of

Forest Management Practices 124

Cindy E. Prescott and Lisa M. Zabek

Tree Species Composition Change-Soil Organism Interaction: Potential Effects on Nutrient Cycling

and Conservation in Interior Forests 137

Alan E. Harvey, Russell T. Graham, and Geral I. McDonald

Out of Sight, but Not Out of Mind: Soil Microorganisms in Moisture-limited Forests 146

Matt D. Busse

Managing Root Diseases by Managing Stands 156

Walter G. Thies

Role of Soil Organisms in Restoration and Applications to Management

Structure and Function of Biological Soil Crusts 161

Jayne Belnap, John Williams, and Julie Kaltenecker

The Role of Soil Organisms in Restoration 179

Michael P. Amaranthus, Efrén Cáizares, and David A. Perry

Soil Organisms-Applications for National Forest Management 190

Karen Bennett
Soil Organisms-Applications to Bureau of Land Management Rangeland Management 193

Mike Pellant

Natural Resources Conservation Service-Databases and Analytical Services 198

Carol D. Franks

Soil Organisms: Functions and Processes-Management Implications. A Synopsis 205

Martin F. Jurgensen

Abstracts of Papers Not Submitted

Soil Foodwebs 211

Elaine Ingham

Nutrient & Soil Organism Population Dynamics-The Basis of Ecosystem Stability 212

John C. Moore

Forest Management Interactions with Soil Organisms 214

James R. Boyle

Acknowledgments 215
OPENING REMARKS
OPENING REMARKS

William G. Ypsilantis1

I'm pleased to join the rest of the steering committee

and speakers in welcoming each and every one of you

to the Pacific Northwest Forest and Rangeland Soil

Organism Symposium. It's encouraging to see so many

of you in attendance at this important symposium in

spite of the tight travel restrictions we all are enduring.

I understand we have over 200 registrants for the

symposium which is fantastic. We have attendees

from several agencies, universities, and the private

sector, and from as far away as Dundee, Scotland.

The purpose of this symposium is to present state-ofthe-science information on the important roles and

functions of soil organisms and some key management

implications in Pacific Northwest forest and rangeland

ecosystems. When Bob Meurisse, US Forest Service's

Region 6 soil scientist, and I first got together over a

year ago to discuss soil organisms, we felt it was

important to focus the efforts of a symposium towards

forest and rangeland managers and a broad array of

resource specialists. The steering committee wanted

to make sure that this symposium helped make a

difference in how we do business in managing forests

and rangeland and was not just another symposium to

be soon forgotten.

A diverse partnership of organizations joined together

to sponsor this symposium including the Bureau of

Land Management, US Forest Service, Natural

Resource Conservation Service's Soil Quality Institute,

Oregon State University and Oregon Society of Soil

Scientists. Our steering committee put together a

wonderful program and guest list of speakers. The

enthusiastic participation by our outstanding lineup of

speakers and poster presenters was especially

gratifying. World renowned researchers from the

Pacific Northwest, Canada, and such distant points as

Yale University and Sofia, Bulgaria will present papers

and posters. Elaine Ingham and Andy Moldenke

graciously agreed to organize the field tour to provide

opportunities for attendees to view field indicators for

forest soil organisms. After many months of hard work

and coordination by the steering committee and the

Forestry Conference Office at Oregon State University

we have finally arrived at this momentous kick-off to

our event.

Our goal for this symposium is to share knowledge

with you about a topic that has only recently begun to

be recognized and understood as vital to the

sustainability and health of ecosystems. When I was in

school we didn't spend that much time studying soil

organisms nor have we dealt with them that much in

forest and rangeland management up until fairly

recently. They have come to the forefront of peoples

awareness and now seems a perfect time for this

symposium. Decision makers want to know the long

range implications of their management actions. It is

up to the scientific community to provide the

knowledge needed to understand the functions and

processes of the soil biological components of the

forest and rangeland ecosystems. Another role of the

scientific community and resource specialists is to

convey this information to the decision makers in a

clearly understandable manner that relates to the

bottom line of ecosystem health and sustainability.

Hopefully, the outcome of this symposium will be to do

exactly that, to convey some of this knowledge in an

easily understandable and digestible manner. Another

hoped for outcome is that it brings together some of

the stakeholders in a way that will lead to partnerships

and sustainable resource management on our forest

and rangeland ecosystems.

This is a very ambitious program we have put together.

The realm of soil organisms is a vast topic that we

can't hope to cover in depth in a 3 day symposium. If

time permits a few questions will be answered by the

speakers immediately after their presentations.

Otherwise, at the end of each half day session there

will be 15 minutes set aside for questions for the

speakers. We hope that this symposium will spark

additional communications between research and

applied management about soil organisms so we have

provided biographical sketches of our speakers in your

handouts with phone numbers and e-mail addresses.

Please feel free to keep up the dialogue started here.

Proceedings will be published for this symposium to

provide the valuable information shared here to those

who were not able to attend. All registered attendees

will receive a copy of the proceedings.

Don't forget the poster session and mixer tonight at

O'Callahan's restaurant at the Ramada Inn. This is a

1William G. Ypsilantis is a soil scientist with the Upper Columbia - Salmon Clearwater Districts, USDI Bureau of Land Management,

1808 N. Third St., Coeur d'Alene, Idaho 83814.

1
wonderful opportunity to meet and share ideas with

some of your peers and others who share a common

interest with you in an informal setting.

I apologize that we could not accommodate all the

attendees who wanted to participate in the field tour.

We felt that we could most effectively communicate

with small groups in such a setting. Those of you who

are registered for the field tour on Thursday afternoon,

we will be boarding the busses in front of the LaSells

Stewart Center and departing at 1:00 o'clock sharp.

Be sure that you are prepared for foul weather as the

tour will be conducted rain or shine.

In summary, I encourage each of you to participate in

the symposium as much as possible and to have a

productive and enjoyable three days here. So come

with us to "Discover Another World;11 the remarkable

world of soil organisms. Thank you.

2
ASSESSMENT OF BELOW-GROUND ORGANISM

FUNCTIONS AND PROCESSES
A NON-VALUE BASED FRAMEWORK FOR ASSESSING ECOSYSTEM INTEGRITY

Kristiina A. Vogt1, Daniel J. Vogt2, Paul Boon3, Anna Fanzeres4, Philip Wargo5, Peter A. Palmiotto6, Bruce Larson7,

Jennifer L. O'Hara8, Toral Patel-Weynand9, Eva Cuadrado10, and Joyce Berry11

ABSTRACT

Indicators of ecosystem state change are currently selected based on the human value for some resource (e.g.,

biodiversity, water quality, etc.) or the human desired end-point of a system (e.g., existence of late successional

forests, lack of insect or pathogen outbreaks, etc.). Examples are used from ecosystem studies and forest certification

protocols to highlight the problems of selecting indicators based on our derived values. Selection of value-based

indicators may misidentify the important constraints controlling the functioning of a system and may result in poor

predictions of system sustainability. Several examples are presented on how to develop a non-value based framework

of analysis to select parameters capable of predicting how a system functions, or to identify the dominant constraints

for each site. The value of using large data sets that can be analyzed to select non-human value-based constraining

variables at different scales is discussed. This process of selecting the dominant constraints also highlights why the

same suite of indicators cannot be used automatically for all sites.

Keywords: indicators, value-based indicators, biodiversity, late successional forests, forest certification protocols,

ecosystem constraints, sustainability

INTRODUCTION

There is a great deal of interest in identifying and

defining criteria that can be used to determine or

assess whether an ecosystem is sustainable under

different types of management activities, or to select

indicators that will sensitively reflect potential changes

in ecosystem condition (i.e., ameliorating or degrading)

(Bormann and others 1993; Mangel and others 1996;

EPA 1996; Vogt and others 1997; Wijewardana and

others 1997). Emphasis on indicators to assess

ecosystem sustainability and/or change has resulted in

the identification of a broad suite of indicators as part

of the assessment process, selection of indicators 'a

priori', and the use of the same indicators at all sites.

This suite of "generic" indicators are not specific to any

site but have been identified to be important in

controlling system function in some way. Several

questions need to be addressed prior to the acceptance

of such an approach to assess sustainability. For

example: have the appropriate indicators that identify

system sustainability been included in the analysis? Is

the emphasis on understanding an acknowledged

problem going to result in the identification of a wrong

set of indicators (value-based) which are not sensitive

to detecting that the system resistance or resilience is

changing? Is it appropriate to assume that the same

suite of generic indicators will automatically transfer to

all sites? What scale of analysis is most appropriate?

Will we have to always scale information from the

small scale (with its high data needs) to the large scale

(smaller data needs) to make it relevant for policy?

To answer these questions we need to examine how

indicators are chosen, and how and when they reflect a

human value that we have for a resource and when

they do not reflect the driving variables maintaining

system resistance and resilience (Vogt and others

1997). Using a value-based approach to drive the

selection of indicators may not be a problem when the

I

 Kristiina A. Vogt a professor of ecosystem ecology at the School of Forestry and Environmental Studies (FES), Yale University,

New Haven, CT.

2Daniel J. Vogt a professor in ecosystem ecology and forest soils at Yale FES.

3

 Paul Boon a research technician in GIS at Yale FES.

4Anna Fanzeres a tropical ecologist, forester and Research Assistant at Yale FES.

5Philip Wargo a plant pathologist and Project Leader at the USDA FS Northeastern Center for Forest Health Research, 51 Mill Pond

Road, Hamden, CT.

6Peter A. Palmiotto a tropical and temperate forest ecologist and Research Assistant at Yale FES.

7Bruce Larson a professor of stand development and Director of Yale School Forests at Yale FES.

8

Jennifer L. O'hara a tropical and temperate forest ecosystem ecologist and Research Assistant at Yale FES.

9

Toral Patel-Weynand a tropical and temperate forest ecosystem ecologist and Research Assistant at Yale FES.

10Eva Cuadrado an agricultural ecologist and Research Assistant at Yale FES.

11Joyce Berry a social ecologist and Lecturer at Colorado State University, Ft. Collins.

3
identified value-based indicators is related at the

holistic level to the condition of the system. For

example cryptogamic crusts are not only highly valued

visual aspects of desert ecosystems but also play an

important functional role in sustaining the system by

deterring the loss of soil by wind erosion (West 1982).

In this case, the identification of a value-based

indicator is highly relevant and an extremely useful way

to identify an important constraining or functional

indicator of site condition.

When human value-based indicators do not reflect

system integrity, the ability of detecting potential

system degradation becomes difficult. Frequently, the

value identified indicators may not change in a

measurable manner during the time-scale of the

assessment period or changes in the indicator do not

reflect system degradation or loss of system resilience

(Vogt and others 1997). One such indicator is Net

Primary Production (NPP) which can be perceived as a

value-based indicator since it documents the amount of

potential wood product output of a site and how well a

system is growing ecologically; typically higher NPP is

translated as being positive since more wood can be

harvested from a site and a system must be functioning

well since sufficient photosynthate is present to

maintain all components of the system (Vogt and

others 1997). However, when measuring the impact

of atmospheric N additions into ecosystems, early

monitoring of NPP would reflect the initial fertilizer

effect of these inputs but not the potential future

decreased NPP and possible "impending degradation"

of the ecosystem. A focus on measuring NPP would

not initially detect the fact that continuous atmospheric

inputs of N will eventually result in leaching losses of

base cations from the soil which results in nutrient

deficiencies limiting plant growth (e.g., decreases in

NPP) (Ingestad and Agren 1992; Lawrence and others

1999).

Problems that can occur when selecting human valuebased indicators over non-value-based indicators are

highlighted in studies using insect and pathogen

outbreak status as an indicator of forest or ecosystem

condition. The absence of serious insect pests or

pathogen problems would indicate a "healthy" system

whereas presence or abundance would indicate

degradation in the system. However, in some systems,

by the time pests are obvious, significant changes may

have already occurred (e.g., secondary chemical

production for defense, carbon allocation changes, loss

of symbionts on root systems, soil chemistry changes,

etc.). However, the pests may only be acting as

secondary agents and merely responding to some other

ecosystem changes (Wargo 1979, 1981; Lawrence and

others 1999; Shortle and others 1999). Even

organisms considered as "primary stress agents" (i.e.,

those capable of affecting healthy trees) may be

responding to changed ecosystem conditions. In the

past, research focused on controlling the pest rather

than understanding its relationship to forest or

ecosystem condition.

Current analyses of pest problems within ecosystems

have taken a more holistic approach (Wargo and others

1993; Shortle and others 1999). The insect or

pathogen is no longer automatically looked upon as the

problem. The treatment of the value-indicator results in

the reduction or eradication of the pest. Rather the

pests can be evaluated as an indicator of other

problems or changes within the system that allows the

pests to reach outbreak conditions. This approach

leads to the development of other indicators of change

that occur prior to the subsequent problem of pest

outbreaks. However, some pests can become

established in response to ecosystem change and

subsequently they can become the constraining

variable in that system. Armillaria root disease in

western conifer ecosystems is an excellent example of

this relationship (Shaw 1980; Wargo 1981). However,

it is still important in these systems to identify and

define other potential constraints which could function

as controlling variables that maintain system function

(e.g., soil chemistry) and perhaps even override the

effect of the pathogens (Vogt and others, in review).

In other situations, management may produce value

driven end-points that may result in a desire to maintain

a system in a particular condition that cannot be

sustained by the driving variables. This may result in

system instability that may proceed undetected

because the wrong indicators were chosen.

Maintenance of a human value indicator may actually

result in the degradation of a site or may require large

expenditures of energy to keep the system in the

desired or valued state because this management goal

may conflict with the natural dynamics of the system.

A good example of this is the late successional forests

in the Pacific Northwest in the United States. In this

case, the value indicator is the maintenance of late

successional or old growth forests for the associated

old-growth dependent species (e.g., spotted owl,

marbled murrelet, etc.) (Hansen and others 1991; Spies

and Franklin 1991; FEMAT 1993). However, most

systems are so dynamic, or our knowledge may be

insufficient, so that it is extremely difficult to maintain

certain stages of forest succession. In this case, the

valued condition is late successional forests to be set

aside as reserves and maintained as part of the forest

landscape - the spatial mapping of these forests with

potential corridors identified and establishment of

reserves with no tree harvesting as critical elements of

this approach. However, the future presence of each

forest fragment will vary between fragments depending

on their location within the matrix landscape, the

quality of its edge environment and legacies persisting

in the system (which is controlled by past land-use and

the disturbances particular to that region) (Franklin and

4
Forman 1987; Franklin and Spies 1991; Vogt and

others 1997; Vogt and others, unpublished data).

Natural disturbance cycles and the manner in which our

landscapes have been fragmented have resulted in

development of late successional forest fragments

which have a higher rate of tree turnover than the

about one percent turnover seen in a less fragmented

landscape (Franklin and Forman 1987; Vogt and others

1997).

VALUE-BASED INDICATORS OF TWO

FOREST CERTIFICATION PROTOCOLS

The recent push to develop certification protocols for

forest management (Upton and Bass 1996) also

demonstrates the problem of how to select and use

indicators of changing forest conditions that reflect

sustainable social and natural systems under different

management regimes (Vogt and others 1997). These

certification protocols are being developed because

there has been a loss of 'trust' by the public in how our

natural resources are being managed and harvested.

Since many of the certification protocols were designed

by groups with particular philosophies, they strongly

reflect the values of the organization designing the

protocols. For example, the goals of the Silva Forest

Foundation protocols are to have ecologically

responsible forest use (e.g., ecosystem based approach

to plan and carry out human uses giving priority to

protection, maintenance and restoration of ecosystems)

and ecologically responsible timber management (e.g.,

timber management plans and activities developed and

carried out to protect, maintain and restore fully

functioning forest ecosystem at all temporal and spatial

scales) (Hammond 1997). The goal of Smartwood -

Rainforest Alliance protocol is to sustain the capacity

of forests to maintain their health, productivity,

diversity and overall integrity for long temporal scale

while being managed for human activity and use (Smart

Wood Certification Program 1997).

The existing forest certification protocols generally

identify two categories of certifier organizational

generated values to monitor: 1) maintain biodiversity

and water quality and 2) alleviate poverty of local

people. These values are reflected in the amount and

type of data required to be collected for the two

certification protocols (table 1). The other fact that is

important to note is the scoring system used for these

two certification protocols results in some information

receiving a greater weight and therefore having a

greater chance of influencing the success or failure of

being certified. The Silva Forest Foundation protocol

has a scoring system that varies from -3 to 3. As part

of this system, scores are given for the listed indicators

which are eventually tallied and used to determine

whether a threshold of points had been achieved and

whether the forest should be certified. The Smart

Wood Rainforest Alliance protocol uses a scoring

system that goes from 1 (extremely weak) to 5

(outstanding performance) with a minimum score of 3

required for certification. Again, the total number of

points obtained from assessing the different indicators

are used to assess whether sufficient points were

achieved for certification. The scoring is very

subjective since clear thresholds that indicate when one

should move from one scoring level to another is

difficult to identify. The bias towards biodiversity can

be seen in several categories in which species

inventories or species composition appears in the

protocol; this bias results in weighting the information

needs toward the values embedded in the protocol.

The type of information required for these protocols

can be separated into several different groups and is

summarized in table 1.

After examining table 1, one should immediately notice

the tremendous amount of data that are required as

part of each of the two protocols presented above and

how the data needs are relatively similar for both

protocols. However, these protocols are not designed

to identify if different driving variables are controlling

the functioning of a system (i.e., the non-value based

indicators of ecosystem integrity). In fact, there exists

a perception that if an assessment framework is to be

credible, a clear list of indicators must be used that do

not vary between different assessments.

Certain types of information are not required to be

collected as part of the data needs of the certification

protocols. For example, no knowledge of the following

ecosystem components is required: microbial

components, decomposers, symbionts, soils (physical

or chemical characteristics), pests and pathogens, and

nutrient or trace element (e.g., Al, Zn, Pb, etc.) cycling

or pools. These components are not included because

they require more complicated approaches or

specialized knowledge since they cannot be analyzed

using visual criteria. The absence of these components

is noteworthy since there is a great deal of

documentation on the impact of human activities on

these variables and they have been identified as the

driving variables causing changes in ecosystem

resilience and resistance (Vitousek 1994; Vogt and

others, 1997; Shortle and others 1999; Lawrence and

others 1999).

ANALYSIS OF PROTOCOL EFFICACY

An analysis of the two protocols in table 1 shows the

strong bias towards measuring and documenting

5
Table 1 —The value- based indicators required to be collected for the Silva Forest Foundation and Smart Wood - Rainforest Alliance

certification protocols (Hammond 1997; Smart Wood Certification Program 1997).

Values

Law enforcement

Adequate rules

Land security

Tracking products

Credibility for

consumers

Silva Forest Foundation

1.2/4.3. Mission and objective statements are clear

and ecosystem-based. There is a commitment to

constant improvement.

1.3 Timeframe of plan is at least 1 50 years.

Projections of successional processes and

provisions for the maintenance of successional

composition and structure are described.

Schedule of economic development,

infrastructure maintenance, and ecological

restoration is defined and includes predictions of

management impacts.

2.1.3/2.2.2/2.3.5/2.4.3/2.5.3/5.1.3 / 5.2.2

Permitted uses are designed to maintain full

ecosystem functioning. Agriculture, ranching,

mining, buildings, roads, vehicles, and logging

generally are not permitted (except for

watercourse crossings in 5.1.3).

2.3.6 If old growth is rare, all old growth stands

and fragments are protected.

2.5.1 At least 5% of applicant lands are in whole

protected watersheds.

3.1 Allocation of forest uses across the landscape

is ecologically responsible and balanced.

3.3.1 Large protected reserves are at least 5 times

the size of the largest natural disturbance to be

expected in the landscape. The overall size of

the landscape in question will influence the

actual size of large protected reserves.

3.3.2 At least 5% of applicant lands are in large

protected reserves.

3.3.3 Permitted uses of reserves are designed to

maintain full forest functioning.

3.3.4 Logging and mining are not permitted in large

protected reserves.

4.4. Strategies and prescriptions can be expected

to achieve objectives.

5.2 Ecologically sensitive (ES) sites are protected.

7.2 Pesticides, synthetic fertilizers, and other

chemicals are not used except in special cases

described in standards document.

7.2.1 No pesticides, herbicides, or synthetic

chemical fertilizers are used.

7.5 Wastes are properly handled and disposed of.

7.5.3 Applicants participate in recycling and reuse

programs where they exist.

9.2.3 No exotic or genetically engineered species or

stock are planted.

9.3.2 No mechanical site preparation is done.

9.3.3 No non-tree species are targeted for

elimination.

9.3.4 No pesticides, herbicides, or synthetic

chemical fertilizers are used.

10.1 Salvage logging is in agreement with all other

SFF standards except as described below.

10.2 No live trees are cut in salvage operations.

11.3/1 2.3 No logging is done on areas under

dispute.

13.2 Health and safety standards meet all

applicable laws and regulations.

14.2 Financial considerations do not dictate

management decisions.

Smart Wood - Rainforest Alliance

1.0 General Information: name, address, person in

charge, total volume produced per year,

processing facilities, owner, ownership,

identification of local communities affected by

the project, etc.

1.8 Percentage of timber supply from: - forest land

that is directly managed by the applicant; - forest

land that is managed by a known third party, or;

- open market log timber purchases (from

independent loggers).

2.1 Land title is clear and legally secure, boundaries

are identified on the ground and timber harvest is

legally permitted.

2.4 A maximum of 10% of the lands (in area and

number) under management supervision leave

the program each year, with specific records

kept explaining the reasons for leaving the

program.

2.5 No enrollment of "single harvest only

"individual forest ownership is taking place.

5.1 Forest management operations exceed or

comply with current Federal, State and Municipal

laws and regulations including Best Management

Practices (BMP) and all necessary permits have

been obtained.

5.4 Roads, stream crossing and log landings are

designed to meet or exceed State Best

Management practices.

5.1 5.iv) roads into closed compartments blocked to

prevent unauthorized entry.

6.10 Based on the identification of fragile,

endangered, unique or uncommon natural

communities, appropriate areas, considered

within a landscape ecosystem context, are

designated as "Reserves or Special Management

Zones "on maps and in the field.

6.13 Operation has a policy for soil and watershed

management, addressing riparian buffer zones

and watershed protection. At minimum, the

policy complies with State BMPs to maintain or

improve water quality and aquatic health, and to

minimize on-site and downstream damages

caused by harvesting operations.

6.14 Operation has a policy for steep slope

harvesting. Policy is written and implemented.

6.15 If chemicals are used, operation has a written

policy addressing the ecological rationale(s) for

their use, including the types used and methods

of application. Operation implements the policy.

7.5 Taxes on forest land and timber are paid in a

timely manner in accordance with state and local

tax laws.

7.6 Known archeological sites, and sites of cultural,

historical, community or aesthetic significance

are included in special management zones or

protected during harvest operations.

9.1 Documentation of management activities on all

forests blocks should be available to facilitate

monitoring and certification audits.

9.2 Certified forest products should be clearly

identified through marks, labels, tags, or

6
Values

Empowering the poor

Rights of indigenous

peoples/local

communities

Labor rights

Other social benefits

Documentation/data

collection

Planning

Maps

Silva Forest Foundation

18.2 Indicators of impacts due to timber

management are appropriate, cost-effective, and

adequate.

1.6 Social and economic needs of local

communities are analyzed. Provisions for these

needs are clearly stated.

3.1.1 All local ecologically responsible forest user

groups are given access to the landscape in

proportion to their needs, zoning ensures fair,

balanced land bases for ecologically responsible

users.

3.1.3 Zoning of uses serves the specific social and

economic needs of local communities.

4.7 Social and economic needs of local

communities are analyzed, provisions for these

needs are clearly described.

11.1 Indigenous people control activities on their

territories.

11.2 Needs for protection and support of

indigenous culture are met.

12.1 Traditional and customary rights are protected

through participation by local communities in

decision-making and planning.

12.3 Compensation and/or mitigation is made for

avoidable and predictable damage to public

values and resources due to logging.

12.4 Management strives to contribute a maximum

of value and stability to local communities

through ecosystem maintenance, ecosystem

restoration, investment, employment, valueadded manufacturing, etc.

13.1 Remuneration packages and contracting

agreements are fair and equitable to all parties.

13.3 Rights of workers to organize and bargain

collectively are protected.

13.4 Worker sharing in the benefits of business

success is encouraged.

18.1 Management reviews ecological, social, and

economic impacts of activities immediately after

each entry and thereafter on at least an annual

basis.

1.1.1 Ecosystem types, including distribution, tree

age, composition, and area, are described.

1.1.2 Natural disturbance types, patterns,

frequencies, and sizes are described.

1.5 Description and maps of protected areas and

human use areas are adequate to guide on-theground activities.

1.8 Plans for monitoring impacts of management

are clearly described, including indicators to be

monitored and timing of monitoring.

2.2.1 Identification of ES (Ecologically sensitive)

sites is accurate and comprehensive.

2.3.2 As much as possible or at least 5% of lands

are in OG (old growth) nodes and OG

recruitment.

2.3.4 Nodes are designed to represent a maximum

of ecosystem types indigenous to the landscape,

rare, threatened, and endangered types are

preferred.

2.4.2 Corridors are at least 1 50 m wide, plus a 50-

m buffer on either side where only partial cutting

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separately documented loads, in order to

facilitate chain of custody processing at mills.

- Specific Smartwood Chain of Custody Guidelines.

1.21 Local communities that are involved

with/affected by forest management activities:

(name organizations, adjoining owners and

individuals).

6.17 Employees are trained in proper handling,

storage and disposal of chemicals, apply

chemicals according to label direction, and

protective equipment is available and used.

7.1 Where appropriate forests are used as a training

and/or educational resource for local people in

conjunction with schools, community colleges,

and/or other providers of training and education

and/or forest managers contribute in some way

to public education about forestry practices,

sustainable forestry and the values of forest

ecosystems. 7.2 Summaries of forest

management policies are available to local

communities upon request.

7.3 Local communities and businesses are

considered for employment, recreation and

benefits related to forest management activities.

7.4 Forest management considers community goals

for forest and natural resource use and

protection as articulated in current and approved

community plans and is consistent with

community zoning laws.

7.7 Wages, facilities and other benefits (health,

retirement, worker's compensation, housing,

food) are fair and consistent with (not lower

than) prevailing local standards. More successful

operations should provide better benefits.

7.8 Worker safety and training is an explicit

element of all aspects of forest operations.

1.11 Types of forests that are being managed.

1.12 Are there any plantations?

1.13 Total acres being managed (commercial

production vs non-productive lands).

1.14 Are there any management plans? If so, how

many and covering how many acres?

1.15 Total volume of timber coming from each type

of forest.

1.16 Acreage of forest being harvested each year.

1.17 Year in which harvesting operation began.

1.18 Do harvesting operations include construction

of primary roads?

1.19 Modes of timber extraction (mechanical or

manual, specify equipment used).

2.2 Land is dedicated by owners to long-term forest

management.

2.3 A clearly defined pool of lands has been

established.

3.1 Sufficient information exists on which to base

forest management plan.

3.2 Long-term monitoring is implemented to enable

7
Values Silva Forest Foundation

to SFF standards is practiced, corridor frequency

is at least 1 for every 3 km of valley length.

2.5 On lands of over 20,000 ha, whole protected

watersheds are designated.

2.5.2 Whole protected watersheds are as big as

possible or at least 2000 ha.

2.5.4 Whole protected watersheds are designed for

optimal representation of the full range of

ecosystem types found within the larger

landscape.

3.2 Roads and landings are designed to maintain

the full forest functioning.

3.2.1 Roads are generally located on stable and

moderately stable terrain, roads are fit into

terrain, soil, and moisture constraints imposed by

ecosystem characteristics.

3.2.2 Main road and spur road density is defined.

3.3 On lands of over 100,000 ha, large protected

reserves are designated.

3.3.5 Large protected reserves are designed for

optimal representation of the full range of

ecosystem types found within the larger

landscape.

4.1 Description of character and condition of

applicant lands is adequate to guide management

activities.

4.1.1 Ecosystem types, including distribution and

area, are described.

4.1.2 The frequency of non-tree plant species are

described by layer or height class.

4.1.3 Tree species, heights, diameters, ages,

growth rates, and general health are described.

4.1.4 Snag species, heights, and diameters are

described; large fallen tree species, lengths,

diameters, and decay classes are described.

4.5 Descriptions and maps of location and human

uses of the protected network of ecosystems are

adequate to guide on-the-ground activities.

4.8 Plans for monitoring impacts of management

are. clearly described, including indicators to be

monitored and timing of monitoring.

5.1 Protected riparian ecosystems (PRE's) are

designed to maintain full ecosystem functioning.

5.1.1 Protective measures in PRE's are justified

from a site-specific ecosystem-based

perspective.

5.2.1 Identification of ES (Ecologically Sensitive)

sites is accurate and comprehensive.

6.1 At least 10% of dominant and co-dominant

trees are marked/mapped and reserved as fullcycle trees.

6.2 Full-cycle trees represent all dominant and codominant species present in the stand previous

to logging.

6.3 At least 25% of the pre-logging number of

snags are retained, including at least 50% of

large-diameter-class snags.

6.4 At least 50% of the pre-logging volume of large

fallen trees are retained.

6.7 Trees are felled, limbed, and bucked on-site. no

whole-tree harvesting is done.

7.1.2 Only partial cutting to SFF standards is

practiced, windfirm trees are favored.

7.1.4 All fallen trees and 80% of pre-logging

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adjustment of management plans and strategies.

3.5 Forest management plan is implemented in the

field as written. Necessary modifications and/or

revisions are appended to the plan.

3.6 Maps are produced at an adequate detail, scale,

and provide operational guidance for

management activities and facilitate on-site

monitoring.

4.3 Management strategies prevent over harvesting

of individual tree species.

4.4 Management strategies emphasize long term

stand quality.

3.4 A forest management plan is written and

available that includes:

i) multi-year planning,

ii) clear landowner and management objectives,

iii) analysis of cruise and inventory data relating to

target species including regeneration,

iv) description of specific silvicultural prescriptions

to achieve desired future condition,

v) description of access issues and constraints,

vi) protection measures identified for special

management areas,

vii) forest insect and disease conditions that may

affect long term forest health condition,

viii) forest management practices for soil

conservation, water quality protection, sensitive

sites, unique natural communities, aesthetics,

chemical use and handling,

ix) monitoring and update schedule.

5.2 The road, log landing and skidding are

designated prior to harvest and are minimized,

while providing transportation safety and hauling

needs.

5.3 Maps have been prepared before harvesting or

road construction, which specify areas which are

suitable only for dry-weather and/or frozen

conditions and indicate locations of truck roads,

landings, main skid trails, no-cut zones, and

other conservation areas.

5.5 A pre-harvest inventory and sale area

reconnaissance is implemented.

5.6 Trees are marked or clearly designated prior to

harvest.

5.7 An operating/harvesting plan is written,

available, and used in the field that includes:

i) silvicultural objectives,

ii) volume and basal area targets,

iii) residual species composition, and

iv) transportation and access issues.

5.8 Pre-harvest silvicultural activities are

undertaken, as appropriate to the operation's

needs, scale and intensity.

5.11 Skid trail placement is acceptable for site

conditions and minimized.

5.13 Log landing and hauling activities are

appropriate for the scale of the operation.

5.15 Appropriate timber sale area closure actions

are taken in areas that are no longer in use.

8
Values

Production

Economic returns

Finances

Silva Forest Foundation

number of snags are retained.

7.3 Roads and landings are designed to maintain

full forest functioning.

7.3.1 Road and landing layout agrees with the

landscape plan.

7.3.3 Right-of-way generally is less than 12 m,

with no unnecessary disturbance on cut slopes,

fill slopes, and road surface .

7.3.8 roads and landings are built to last as

permanent structures.

7.3.9 Old roads are rebuilt wherever this will reduce

overall degradation of lands.

7.3.10 Roads are not used during periods of

instability (e.g. during freeze-up or break-up).

7.3.11 Impacts from construction are minimized,

and less than 10% of lands with in a logging unit

are affected by all roads taken together,

including skid roads and landings.

7.3.12 Maintenance of roads and landings when

not in use is adequate to prevent an increase in

impacts through deterioration.

7.4.1 Ground-based systems are used only on

slopes less than 40%, with skid roads on

microslopes less than 20%. slopes between

40% and 60% are logged only with cable or

aerial systems.

7.4.2 All spur roads, skid roads, and landings are

pre-located in the field.

7.4.3 Area of disturbance by spur roads, skid roads,

and landings is less than 10% for ground-based

systems and less than 7% for cable and aerial

systems within a logging unit.

7.4.4 Spur roads, skid roads, and landings are

designed as permanent structures.

8.2 Cutting over any period removes no more than

80% of the growth for the same period.

9.1.1 Management for natural regeneration is

integrated with the planning and implementation

of logging activities.

9.1.2 Damage to advance regeneration during

logging is minimized.

18.3 Reserved areas, composition, and structures

are clearly marked, mapped, or otherwise

designated to guide on-the-ground activities now

and in the future.

8.3.2 tree selection will result in improvement of

the stand's timber value.

10.3 Salvage logging is necessary from an

economic viewpoint.

14.1 The applicant has the financial means to carry

out the management plan.

15.1 Product wastage is minimized.

1 5.2 Systematic efforts are made to diversify

product types and to maximize the value of

extracted wood.

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1.6 Total volume (actual and estimated) of timber

produced per year. Includes different type(s) of

product(s) being produced and the volumes of

each type.

3.7 Non-timber forest products have been

considered and their management is incorporated

into planning process where appropriate. Where

applicable this may include: ginseng,

mushrooms, maple syrup production, or provision

of recreational opportunities.

4.1 An allowable cut has been derived based on

well-documented estimates of growth and yield

to provide a non-declining sustained yield of

forest products and this target is being followed

in harvest planning.

8.1 Stumpage rates being paid to landowners are

fair and representative given logging conditions,

timber quality, volume and local experience with

log markets.

9
Values

Biodiversity - (plants,

animals, composition,

structure)

Silva Forest Foundation

1.1.3 Relationship between ecosystem types and

disturbances are analyzed.

1.1.4 Present condition is compared with natural

ranges of variability as estimated by the best

available information.

1.4 Identification and accommodation of ecological

limits are accurate and ecosystem-based.

1.7 Minimum landscape-level requirements for

stand-level applicants are met.

1.7.1 Description and analysis of successional

processes on applicant lands are adequate to

guide ecosystem-based planning.

1.7.2 Plans to mimic and maintain successional

processes are clearly described.

2.3 Old growth nodes and old growth recruitment

will provide specialized habitat throughout the

landscape.

2.3.1 nodes are as big as possible or large enough

to provide at least 50% forest-interior habitat

(areas over 250 m from nearest edge).

2.3.3 Nodes are designed to provide habitat for

indigenous rare, threatened, and endangered

species.

2.4 Cross-valley corridors can be expected to

provide opportunity for movements of all forest

organisms known to require mature or

undisturbed forest.

2.4.1 Corridor design provides useable linkages

among components of the protected landscape

network and other ecologically important

landscape ecosystems.

2.5.5 Whole protected watersheds are designed to

provide habitat for rare, threatened, or

endangered species.

3.1.2 Zoning of uses is designed to maintain full

forest functioning across the entire landscape.

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8.2 Revenue received is sufficient to financially

support long-term forest management, e.g.

planning, inventory, resource protection and

post-harvest management activities such as

silvicultural treatments, road maintenance,

growth and yield monitoring, and other actions

to ensure long-term forest health.

8.3 Financial benefits of forest management are

reinvested into long-term management.

8.4 Financial situation, investment objectives, and

debt load do not place undue pressure on the

operation to sacrifice management for short term

high output harvesting production.

8.5 Under-utilized timber species are being

evaluated for harvest and commercial

opportunities.

8.6 Forest operation encourages, where

appropriate, the utilization of non-timber forest

products and recreation to enhance forest

potential.

8.7 Merchantable log loss and waste in the forest is

minimized.

8.8 Logs and lumber are handled so as to minimize

potential loss in value from degrade, exposure,

and parasites.

8.9 Forestry operation seeks the highest and best

uses for individual tree and timber species and

products.

1.20 What ecologically sensitive , critical or unique

natural areas are known to exist in the vicinity of

harvesting?

4.2 Boundaries of harvest areas are respected and

controlled between cutting cycles so as to

prevent disturbance to regeneration.

5.10 Felling techniques minimize damage to

residual forest.

6.2 Management activities maintain continuity of

forest cover at the watershed level and minimize

forest fragmentation.

6.3 Management addresses the diversity,

composition and structure of the forest at the

stand, watershed, and landscape levels.

6.4 Gap sizes and canopy openings are appropriate

to the local ecology and sufficient to regenerate

the stand while minimizing fragmentation of the

structure and diversity of the forest at the

watershed level.

6.6 Management practices addresses the retention

of cavity, den and/or snag trees for wildlife

habitat.

6.7 Management practices actively addresses the

management and/or enhancement of existing

wildlife habitat elements.

6.8 Management activities reduce risks from

invasion or expansion of exotic species in the

forest.

6.9 Conservation of threatened, rare, endangered

and unusual plant and animal species, natural

communities and critical habitats, are explicitly

incorporated into management and harvesting

plans.

6.12 Design and layout of reserves or special

management zones is considered at the

watershed and landscape level. Connectivity of

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11

Values

Water and soil

Silva Forest Foundation

agriculture, mining, buildings, and ranching are

not permitted in the PLN (Protected Landscape

Network).

3.3.6 Large protected reserves are designed to

provide habitat for rare, threatened, or

endangered species.

4.1.7 Rare, threatened, and endangered species are

inventoried.

5.3 Rare, threatened, and endangered species and

ecosystems are protected from logging impacts.

6.5 Species, decay class, and size diversity of

retained snags and large fallen trees

approximates pre-logging diversity, with

emphasis given to larger size classes.

6.6 Spatial distribution of retained composition and

structures approximates patterns caused by

natural disturbance.

7.4.5 Damage to retained trees is less than 5% of

residual stems.

8.1 Logging is designed as much as possible to

mimic and maintain natural disturbance regimes

and the ecosystems they form.

8.1.1 Disturbances being mimicked and maintained

are natural to the site.

8.1.2 Cutting levels do not cause significant

interruption or alteration of natural nutrient

storage and cycling processes.

8.1.3 Cutting patterns mimic common natural

disturbance patterns as much as possible.

8.1.5 Composition and structures are permanently

reserved as required by protected network of

ecosystems.

8.3 Selection of cut trees maintains the natural

biodiversity of the stand, improves tree quality,

and develops a viable economic resource.

8.3.1 Selection favors trees resistance to

windthrow, disease, insects, etc.

8.3.3 Tree selection imitates successional

processes at the individual, species, and

community levels.

9.2.2 Planted trees represent the full range of

species for the successional phase being

mimicked, planting stock is taken from within

the local biogeoclimatic subzone. unfertilized

bare-root planting stock is preferred.

9.3 Brush control and site preparation are designed

to maintain full forest functioning.

9.3.1 Brush control and site preparation mimic

natural successional processes and patterns.

10.4 Salvage logging removes less than 50% of the

fallen trees and less than 50% of the snags

remaining after the disturbance. Trees removed

are be representative of natural species,

diameter, and height class distribution on the

site.

2.1.1 Protective measures in PRE's are justified

from a site-specific ecosystem-based

perspective.

2.1.2 Average riparian buffer width is at least 25 m

on all sides of watercourse or water body, with

none less than 15 m. no-logging portion is at

least 10 m wide, cutting in remainder agrees

with standards in 6.1.

4.1.5 Terrain slope, stability, and aspect are

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forested areas should be considered in planning

of reserves or special management zones.

5.9 Strategies for protecting highly erodible areas

or designated buffer zones and wetlands are

being followed during harvest.

5.12 Skidder operators work with sufficient care to

avoid residual stand damage and minimize

impacts to soil and water resources.

5.15 i) temporary stream crossing structures

removed and stream banks restored,

ii) landings reshaped to secure adequate surface
Values

Miscellaneous

management

practices

Silviculture methods

Chemicals/wastes/

pollutants

Restoration

Plantations

Silva Forest Foundation

described.

4.1.6 Soil types, drainage, stability, and depths are

described.

4.6 Descriptions of protective measures for soil,

water, and permanently reserved composition

and structure are clear, and estimation of

impacts due to timber management are accurate

(within 5% of actual impacts).

5.1.2 Average riparian buffer width is at least 25 m

on all sides of watercourse or water body, with

none less than 15 m. no-logging portion is at

least 10 m wide, cutting in remainder agrees

with standards in 6.1.

7.1 Where permitted, logging in riparian ecosystems

is designed primarily to maintain full functioning

of those ecosystems with particularly large

margins of precaution.

7.1.1 Cutting patterns and levels are justified from

a site-specific ecosystem-based perspective.

7.1.3 Exposure of mineral soil and/or soil

compaction is less than 5%.

7.1.5 Felling is designed to minimize degradation,

felling is away from watercourse or water body

and away from no-logging reserves.

7.1.6 No roads or vehicles are permitted in

protected riparian ecosystems, except for

watercourse crossings.

7.1.7 No products are piled in protected riparian

ecosystems, felling equipment is refueled in

specified zones outside of protected riparian

ecosystems.

7.3.2 Roads are located on stable areas with

minimal cut and fill, roads are fit into terrain,

soil, and moisture constraints imposed by

ecosystem characteristics.

7.3.4 Road drainage structures, including road

profile, road cross-section, ditches, culverts,

bridges, etc. maintain natural drainage patterns

and minimize concentration of flow.

7.3.5 Watercourse crossings are minimized.

7.3.6 Watercourse crossings accommodate 100-

year floods and span the entire riparian zone (but

not necessarily the riparian ecosystem).

7.3.7 Watercourse crossings maintain fish passage

and watercourse bed structure.

7.4 Skidding or yarding systems are designed to

maintain soil and water structure and

functioning.

7.4.6 Bonus points may be awarded for specific

measures taken to protect soil and water during

skidding or yarding.

7.2.2 If a chemical is used with ecological

justification, choice of chemical and method of

application are designed primarily to minimize

damage to ecosystems and non-target impacts.

7.5.1 Toxic and inorganic wastes are disposed of in

approved disposal sites, generally located outside

the forest.

7.5.2 Pollution is minimized in all phases of

operation.

8.1.4 Silvicultural treatments (e.g. prescribed burns)

imitate natural disturbance regimes.

9.1 Natural regeneration is preferred over planting.

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drainage and reseeded, preferably with nonexotic species,

iii) functional cross drains constructed on roads

where appropriate,

v) active erosion control measures (mulching, silt

fence) are implemented as needed in accordance

with State BMPs.

6.5 Management practices addresses structural

heterogeneity, potential for soil improvement and

nutrients enrichment on the forest floor by

maintaining or restoring the presence of coarse

woody debris (CWD).

6.11 River and stream corridors, steep slopes,

fragile soils, wetlands, vernal pools, lakes and

pond shorelines, and other hydrologically

sensitive areas are automatically designated as

special management zones.

Generic Guidelines for Assessing Plantations

3.3 Forest management objectives and

prescriptions are documented and based on site

history, owner's objectives, forest condition,

ecology, management needs, and other relevant

factors. Silvicultural prescriptions are

substantiated using current research, silvicultural

guides and applicable growth and yield models.

4.5 Management addresses the restoration of degraded or low quality forest stands.

4.6 Planting incorporates ecological factors or is utilized as a silvicultural tool to restore high-graded

12
Values Silva Forest Foundation

9.2 Where tree planting is done, it is designed to

maintain natural biodiversity.

9.2.1 Planting is justified from a site-specific

ecosystem-based perspective.

14.3 Investment in equipment and infrastructure

will minimize pollution and degradation.

16.1 Restoration is designed to mimic, maintain,

and, where necessary, reintroduce disturbance

regimes natural to the site and landscape.

16.2 Where practiced, artificial restoration is

justified as a necessary interim stage to allow

eventual restoration of natural composition,

structures, and functioning.

17.1 Restoration activities do not create negative

impacts in excess of expected long-term

benefits.

17.2 Restoration activities account for the

reestablishment and maintenance of natural

biodiversity at all scales of time and space.

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stands to natural species mix and forest cover.

4.7 Planting where it occurs, utilizes mixed species

of native stock, rather that exotics.

5.14 Post-harvest silvicultural activities are implemented where planned with reliability and success,

6.1 Silvicultural prescription have a primary objective

of perpetuating a sustainable forest ecosystem.

Silviculture is practiced with the goal of conserving and restoring ecological diversity and ecosystem function.

6.16 A constant effort is made to: minimize the use

of chemicals, use targeted application methods,

use the least toxic chemical possible, and phase

in non-chemical alternatives.

Waste (e.g. sawmill residues, chipping, debarking,

waste oil, and landing debris) is being properly

disposed.

biodiversity-a value based indicator. Central to the

implementation of the certification analysis is the

measurement, maintenance and restoration of

biodiversity which reappears several times in the

different categories of required data presented in

table 1. However, the effectiveness of focusing on

biodiversity measurements to assess system

sustainability is debatable. It has been assumed that

biological diversity is directly correlated to the

maintenance of system carbon and nutrient cycles

(Karieva 1994). The ability to characterize a site

condition by focusing on species numbers and

assuming that there is a direct cause and effect link

between species diversity and function would simplify

our capability to assess the effect of site disturbance or

a particular management regime. Unfortunately such a

direct relationship has not been shown for most plant

communities (Johnson and others 1996). This lack of

clear links between biodiversity and ecosystem function

is unfortunate since it would greatly simplify our ability

to detect or even predict changes in ecosystems.

It has also been suggested that the biodiversity of

ecosystem components other than plants should be

pursued as indicators of change in system function.

For example, similar to the biodiversity relationships

suggested for plants growing in grasslands (Tilman and

Downing 1994), there has been interest in using

symbiont diversity as an indicator of how the system

was responding to anthropogenic activities. In the

field, the diversity of mycorrhizal species can be high

and has been shown to vary from 7 to over 20

different fungal species found colonizing the root

system of one tree species at any given time period

under undisturbed conditions (Vogt and others 1981;

Gardes and Bruns 1996). However, preliminary

information suggests that the diversity of

ectomycorrhizal fungal species found on plants' roots

may not be directly related to how successfully plants

occupy a site. In one greenhouse study where the

diversity of symbionts on roots were controlled,

researchers were unable to show that increasing the

diversity of fungi on roots conferred an advantage to

the seedlings (Benecke and Gobi 1974; Parladé and

Alvarez 1993). A focus on diversity of species found

on root systems does not address the fact that the

functional diversity may be much more useful to

monitor, especially since the efficiency at which

different fungal species acquire abiotic resources has

been documented to be highly variable (Harley and

Smith 1983). If one knew how spatially heterogeneous

the abiotic resource availabilities were in the ecosystem

and the individual efficiencies of the different fungi,

one might begin to predict how many functional groups

would be needed to optimize resource acquisition. For

plants growing in grassland ecosystems, increasing the

number of plant species beyond 10 did not confer

increased stability to that system suggesting a

threshold for this system (Tilman and Downing 1994).

Another example with Bishop pine (Pinus muricata) also

highlights the importance of assessing factors other

than diversity to monitor ecosystems. In this case, the

ability of Bishop pine to have up to 20 different species

of mycorrhizal fungi on its roots (Gardes and Bruns

1996) was not relevant in assessing its effectiveness in

competing for abiotic resources and outcompeting

other plant species at the same site (Northrup and

others 1995). The success of Bishop pine and its

ability to acquire abiotic resources was dependent not

on the diversity of fungal species present but on the

ability or efficiency at which the symbionts utilized

organic N forms {Northrup and others 1995). In this

example, it was the ability of Bishop pine to produce

litter higher in polyphenolics with slower decomposition

rates that result in most of the available N being

13
present in the organic N form. In this example, several

indicators could have been used to assess the

sustainability of Bishop pine on the site - tree NPP, N

availability in the soil, water availability, mycorrhizal

diversity, litter decomposer biomass or community

composition or decomposition rates. Out of all these

potential indicators, the most sensitive to assess

Bishop pine health would be changes in polyphenol

levels in decaying litter and the presence of symbionts

capable of utilizing organic forms of N. In this case,

the determination of NPP, mycorrhizal diversity or the

development of detailed N budgets would probably not

have been sensitive indicators to monitor; even if

changes in these parameters would be measurable,

they are not the driving variables controlling how the

system responds.

ALTERNATIVE FRAMEWORK FOR

DETERMINING NON-VALUE BASED

INDICATORS

To shift these protocols away from the human value

based approach, we developed a framework that

produces a "road map" identifying what information is

relevant to each site and also shows that the information used to evaluate a site might change between

different locations (Vogt and others, unpublished data).

Our framework utilizes a process of identifying legacies

and the key driving variables controlling the resistance

and resilience of an ecosystem; those indicators that

are typically not identified when using a value-based

approach. These non-value based indicators constrain

the ability of a system to function and influences the

response of the system to perturbations. Our

framework will address the problem of the large data

needs of the Certification protocols by identifying those

site specific variables relevant to monitor.

One of the problems identified with the certification

protocols (table 1) is the need for large amounts of

documented data. The presence or absence of these

data and their content determines whether or not a site

will be certified as sustainable. This need for large

amounts of data makes it almost impossible to assess

many sites because some data may not be available at

some given site. This requires the acquisition of

missing data to fulfill the certification requirements

without even knowing if any of the missing data are

relevant for the site being assessed. This highlights the

need for a method that can determine what information

is associated with or influences the functioning of a

system. It is also important to accept the fact that the

information needs will vary between different sites and

we should not be attempting to use the exact same

information to assess different sites (Vogt and others

1997). The framework presented in Vogt and others

(in review) is important because it develops a road map

that can be used to assess what the information needs

are for different sites. This framework takes one from

assessing a site based on a pre-requisite set of tools

that are used at all sites to a researcher assessing a

specific site by identifying what the driving variables

are that control or constrain how the system functions.

Large data sets which already exist can be quite useful

(Vogt and others 1996) for identifying what are the

constraints controlling the functioning of a site.

Identification of these constraints is especially

important because they frequently are not characterized

by the value-based indicators but function in controlling

or driving how that system functions (Vogt and others,

unpublished data). These constraints are also

important to identify since they are the driving variables

that would be important to monitor to determine if the

condition of the site is changing.

Determining the constraints of an ecosystem using a

large data set would then help identify the non-value

based indicators that control system function. It is

important that analyses do not occur only at higher

spatial scales (see table 2) but that analyses are

conducted at scales at which the system is sensitive to

a particular management activity (table 3). For

example, climatic variables may be identified as

important driving variables controlling the productivity

that is achievable on a site at higher spatial scales

(table 2). However, at smaller scales, productivity may

be poorly predicted by climatic variables (see table 3).

The scale sensitive to assessing the system must be

the one utilized for analysis - whether this occurs at the

smaller or larger scale. Therefore, when climatic

variables exert a strong influence on the functioning of

that system and climate change has been recorded, a

manager's focus should expand to assessing the impact

of climate change on ecosystem processes. In this

example, the influence of climate would have to be

considered as a constraint on the site, even though it

cannot be managed, since it changes how the system

responds.

The information synthesized in table 2 from Vogt and

others (1996) highlights how variable the important

driving parameters are in controlling the aboveground

productivity of sites in general (Vogt and others 1996).

These data encompassed information from the boreal

to the tropical forests and showed that climatic

variables explained about one third of the changes in

Aboveground Net Primary Production (ANPP). At this

scale, information on organic matter accumulation in

the soil or on the surface of the soil explained about

one-fifth of the variability recorded in ANPP (table 2).

These variables do not appear to be useful to predict

how ANPP might change in response to management

activities at these larger spatial scales. Interestingly,

past analyses of these large data sets revealed that

climatic variables were effective at explaining the

14
amount of surface organic matter (e.g., forest floor) but

not soil organic matter (SOM) accumulations (Vogt and

others 1996).

Table 2 —Variables predicting Aboveground Net Primary

Production when data were combined from all sites

given in Vogt and others (1996).

a

 MRT = mean residence time (years)

The information in table 2 shows that the mean

residence time (MRT) of N in the surface organic

horizons may be quite useful as an indicator to monitor

systems. If forest floor mass is useful for predicting

NPP, litter decomposition rates may be useful as

indicators of the productivity changes on the site.

Since it was the mean residence time of N in the forest

floor that explained a greater proportion of the changes

in ANPP, it may be worthwhile monitoring the

decomposer organisms located in the surface organic

horizons as indicators. Whether it would be useful to

monitor the decomposer organisms directly associated

with the litter turnover or whether changes in forest

floor N MRT would be sufficient and a more sensitive

parameters to monitor is not clear. This suggests that

we need to be able to detect changes in the forest floor

N MRT at a level that is sensitive to the changes

occurring in the ecosystem. This analysis, is of course,

at a very large spatial scale and the ability to go to any

given site to predict how the system is responding to a

disturbance may be difficult. They do give at a large

spatial scale, the large scale constraints that may exist

on a site and are therefore useful indicators of variables

that should be monitored.

The information needs will vary depending on the scale

at which the system is being analyzed. It is usually

generalized that less information is needed at larger

spatial scales and that the information needs increase

dramatically as one scales down (Vogt and others

1997). The generalizable relationships shown with

ANPP and all sites combined (table 2) do not transfer

readily to all systems located at smaller spatial scales

(table 3). These relationships have been quantified

using a large data set (Vogt and others 1996) and

expressed at different scales of analysis. Further

scaling down of these predictive relationships show

that the correlations between SOM and ANPP vary

depending on the forest climatic type (Vogt and others

1996). When all forest climatic types were pooled

together, about a fourth of the ANPP variability could

be explained by SOM levels. However, in the warm

temperate, subtropical and tropical climatic forest

types, there existed no predictive ability between these

variables. This contrasted with the cold temperate

climatic type where 34% of the variability in ANPP

could be explained by the SOM levels and the boreal

climatic type where 26% of the variability could be

explained. Pooling of data from all climatic types

suggested a certain predictive capability and what

might be general constraints on the growth achievable

on all forests. However, zoning down to lower scales

immediately showed that certain forest types do not

have the surface organic matter or soil organic matter

accumulations functioning as constraints on ANPP.

Interestingly, the ability to predict total NPP using the

forest floor mass accumulations (r2 = 0.43, p= 0.01)

or SOM levels (r2 = 0.53, p = 0.01) was higher than

for ANPP in the boreal forest climatic type (Vogt and

others 1996). In the boreal forests, these results

suggested that decomposition variables are very useful

indicators of system response. In fact, research from

specific sites suggested that most of the system

energetics flux through the decomposers. This

contrasted with the cold temperate climatic types

where the forest floor accumulations had a poorer

predictive ability while the ability of SOM accumulation

to predict total NPP was the same as for ANPP. This

suggests that monitoring the decomposers might not

be as useful in many of the cold temperate forest

climatic types. In the cold temperate forest

ecosystems, data shown in table 3 suggest that plants

might be a better indicator of changing conditions in

the site.

If total NPP was predicted using forest floor

accumulations and the data are separated by soil order

instead of climatic types, the ability to identify strong

predictive relationships under certain conditions is

highlighted (fig. 1; Vogt and others 1996). For

example, forests growing on an Inceptisol soil order

had strong relationships between forest floor mass

accumulation and total NPP with 62% of the variation

in total NPP explained. No significant relationships

were found for any of the other soil orders. The

15

PREDICTIVE VARIABLES

Forest Floor Mass Content

Soil Organic Matter Content

Minimum Annual Temperature, C

Mean Annual Temperature, C

Forest Floor N MRTa

r

2

(p = 0.001)

0.22

0.24

0.31

0.37

0.67
Table 3—Variables predicting Belowground Net Primary Production (BNPP)

and Belowground Biomass (BB) are shown by Forest Climatic Type (e.g.,

Boreal, Cold Temperate Needleleaf, Warm Temperte Broadleaf Deciduous,

Subtropical Broadleaf Evergreen, Tropical Broadleaf Evergreen) (data from

Vogt and others 1996).

a

 FFN MRT = forest floor nitrogen mean residence time

b

 FFN = forest floor nitrogen

relationship found for Inceptisols is probably related to

the fact that these are young soils and are

characterized by low nutrient availability. The other

soils orders may also have low nutrient availabilities but

are dominated by vegetation and symbionts that have

evolved with the nutrient levels inherent to that site.

Further identification of indicators that would suggest

what type of variables should be monitored can be

identified by examining the information given in table 3

(data from Vogt and others 1996). This information

highlights nicely how the same indicators should not be

automatically transferred to all sites to assess the

condition of a site or how it responds to disturbance.

The data show how indicators vary depending on the

climatic forest type and by whether biomass or

productivity data are being assessed (table 3).

Information given in table 3 also showed how variable

the type of information used to predict the productivity

and biomass of fine roots across these different forest

climatic types can be. When aboveground litterfall

nutrients are more effective in predicting biomass or

productivity, this is probably reflecting the

plant/symbiont level adaptations to the site. When

temperature or the mean residence times of surface

organic matter appear as useful predictors of

achievable biomass or productivity, decomposer

variables may be quite useful as indicators to monitor.

16

Forest climatic type

• Independent variables

Boreal

• max. annual air temperature

• min. annual air temperature

• aboveground litterfall N

• aboveground litterfall P

• FFN MRTa

Cold Temperate Needleleaf

• annual precipitation

• aboveground litterfall N

• aboveground litterfall K

• aboveground litterfall P

Warm Temperate Broadleaf Deciduous

• soil texture

• FFNb

Warm Temperate Needleaf

• mean annual temperature

• soil texture

Subtropical Broadleaf Evergreen

• soil texture

Tropical Broadleaf Evergreen

• maximum annual temperature

• aboveground litterfall N/P ratio

• aboveground litterfall P

Dependent variable

r

2

 (p = 0.01)

BNPP

0.83

0.83

—

—

—

0.13

0.15

—

—

0.75

0.66

0.61

—

—

-

0.99

—

BB

—

—

0.99

0.95

0.63

—

—

0.74

0.64

—

—

—

0.73

0.84

0.81

—

0.53
Table 4—Some general factors controlling soil organic matter

accumulation that are external or internal to the soil

(Stevenson 1965; Lugo and others 1986; Oades 1988;

Johnson 1992, 1995; Vogt and others 1995).

Forest Floor Mass (Mg ha-1)

Figure 1—Net Primary Production predicted using forest floor

mass where the data were grouped by soil order (data

synthesized in Vogt and others 1996 was used).

The scale of measurement analysis can be critical in

determining when a sensitive level of analysis has been

identified and whether the right indicator was selected.

If the incorrect scale is chosen, analysis of impact may

show no impact of an activity when in fact the system

may already be changing. A good example of the

varying sensitivity of detecting system response can be

seen by looking at how organic matter in the soil is

determined - as a total mass or by fractionating it into

its components. For example, determination of total

organic matter in the soil may not show a response to

management since such a high proportion of the carbon

may be in a recalcitrant pool (Oades 1988; Johnson

1992). Thus, an analysis of its component parts may

be extremely useful and may reflect that the system is

changing in the short time scale in response to some

management activity. SOM is generally composed of

different stages of decomposed organic matter each of

which have different turnover rates (Baldock and others

1992).

This raises the question of identifying the scale at

which the system is sensitive to an activity and then

measuring the system at the appropriate scale. This

will probably vary depending on the activity being

measured. The information needs will also vary by

scale and again more factors can be shown to be

important in controlling a pool size. As an example, at

smaller scales, many factors have been identified as

important in affecting the stability of soil organic

matter as shown in table 4 (Stevenson 1965; Lugo and

others 1986; Oades 1988; Johnson 1992, 1995; Vogt

and others 1995). Each one of these factors may

function as a constraint or driving variable in different

sites and thus become crucial to measure and monitor

for changes in the system. In most cases, one will

probably not measure these factors unless one of them

has been shown to be sensitive to management activity

(e.g., acid deposition and decreases in soil pH and Ca

levels; Lawrence and others 1999).

In other cases, a focus on the composition or pool size

of SOM may better integrate many of the factors listed

in table 4; the analysis framework should be used to

determine at what level of analysis is the system most

sensitive to and reflecting the changing system. For

example, when human activities have changed the pool

sizes and fluxes of nutrients, changes in SOM may be a

very useful parameter to measure. Also some factors

such as N availability may have a significant effect on

the total pool of organic matter found in the soil (Vogt

and others 1995). Vegetation type also controls the N

availability and the type of carbon compounds

produced (e.g., simple and complex) both of which

directly affect the organic matter accumulation in the

soil. For example, SOM accumulation has been shown

to be higher in deciduous dominated forests and forests

composed of a mixture of deciduous and evergreen

species than in evergreen dominated forests where N

availabilities are lower (Vogt and others 1995). In this

case, the effect being measured is excess N inputs but

the sensitive variable to monitor in the ecosystem could

be leaching losses of N below the rooting zone or

changes in the SOM content of soil.

CONCLUSION

An examination of indicators used in the existing

frameworks for assessing ecosystem health or

sustainability shows that they are mostly human valuebased. These value-based indicators selected for use in

the protocols for forest certification generally reflect

17

External to Soil

Temperature

Precipitation

Vegetation type

• deciduous, evergreen

• litterfall chemical quality

Disturbance

• type

• time since disturbance

Type of Management

Internal to Soil

Clay composition &

dominance

PH

Base saturation percent

N, Ca availability

Decomposer types and

dominance

Al, Fe - form and

abundance

Soil texture
the interests of the organizations developing them and

are unfortunately considered to be uniformly applicable

across all sites examined. Most of these protocols

focus on biodiversity which has not been shown to be

directly linked to ecosystem sustainability except in a

few cases. These assessment protocols also have high

data needs which reduces their utility in many systems

since the data may not be available or would cost too

much to collect. The perception that much data should

be monitored is augmented by our uncertainty of how a

system functions and our inability to begin to identify

relevant indicators. To develop an efficient framework

for certifying forest management one needs to create

as few protocols and/or indicators to check, measure,

or monitor in a forest site as possible and yet still be

able to insure that ecologically, environmentally,

silviculturally, etc. sound forest management has been,

or is being, conducted. One step to accomplish this in

part would be to determine what indicators are directly

or closely linked to the functions of a system, such as

those that drive or constrain the system. It is not

surprising that many of theses indicators might be

associated with soil ecology. Monitoring these types of

indicators would allow one to know if the system is

changing and, if so, whether the changes are a

consequence of natural events and/or forest

management practices.

Acknowledging the role of value-based indicators in the

selection of indicators to assess system health is

critical because they may incorrectly identify a suite of

indicators which are not sensitive to site conditions.

The value based indicators need to be monitored in any

assessment but if they do not reflect the system

holistically, priority should be given to monitoring those

non-value based indicators. These non-value based

indicators would be identifying those factors that

constrain the functioning of a site. Using a framework

that allows for non-value based indicators to be

identified is crucial for producing effective assessment

protocols.

ACKNOWLEDGMENTS

The ideas developed in this paper occurred while

conducting research supported by the National Science

foundation grant to the Long-term Ecological Research

Program in the Luquillo Experimental Forest, Puerto

Rico and DEB9306758 to Yale University, and from the

USDA Forest Service Northeast Forest Experiment

Station and Northeast Global Change Program. Many

of the ideas also developed as part of Yale University

School of Forestry and Environmental Studies class on

Certification where the teaching assistants Jessica

Lawrence and Fran Raymond acted as catalysts in

synthesizing information on certification. The Yale

Forest Forum directed by John Gordon and Gary

Dunning contributed to the development of this paper

by creating a stimulating environment where these

ideas could be discussed and by obtaining financial

resources to bring in guest speakers for the

Certification class.

LITERATURE CITED

Baldock, J.A.; Oades, J.M.; Waters, A.G.; Peng, X.;

Vassallo, A.M.; Wilson, M.A. 1992. Aspects of the

chemical structure of soil organic matter materials

as revealed by solid-state

 13C NMR spectroscopy.

Biogeochemistry. 16:1-42.

Benecke, U.; Gobi, F. 1974. The influence of different

mycorrhizae on growth, nutrition and gas-exchange

of Pinus mugo seedlings. Plant Soil. 40:21-32.

Bormann, B.T.; Brookes, M.H.; Ford, E.D.; Kiester,

A.R.; Oliver, CD.; Weigand, J.F. 1993. In Eastside

Forest Ecosystem Health Assessment, Volume V.

US Department of Agriculture, Forest Service,

pp. 62..

FEMAT (Forest Ecosystem Management Assessment

Team). 1993. Forest Ecosystem Management: An

Ecological, Economic and Social Assessment.

Washington DC. U.S. Government Printing Office.

No. 1993-793-071.

Franklin, J.F.; Forman, R.T.T. 1987. Creating

landscape patterns by forest cutting: Ecological

consequences and principles. Landscape Ecology

1:5-18.

Franklin, J.F.; Spies, T.A. 1991. Composition,

function, and structure of old-growth Douglas-fir

forests. Pp. 71-79. In: Ruggiero L.F.; Aubry K.B.;

Carey A.B.; Huff M.H. (eds.). Wildlife and

Vegetation of Unmanaged Douglas-fir Forests.

USDA For. Serv. Gen. Tech. Rep. PNW-GTR-285.

Portland, Oregon.

Gardes, M.; Bruns, T.D. 1996. Community structure of

ectomycorrhizal fungi in a Pinus muricata forest:

above- and below-ground views. Canadian Journal

Botany. 74:1572-1583.

Hammond, H. 1997. Silva Forest Foundation

certification program evaluation scoring checklist.

Draft 1.16 October 1997. 16 pp.

Hansen, A.H.; Spies, T.A.; Swnason, F.J.; Ohmann,

J.L. 1991. Conserving biodiversity in managed

forests. BioScience. 41:382-392.

18
Harley, J.L.; Smith, S.E. 1983. Mycorrhizal

Symbiosis. Academic Press, London. 483 pp.

Ingestad, T.; Agren, G.I. 1992. Theories and methods

on plant nutrition and growth. Physiol. Plant.

84:177-184.

Johnson, D.W. 1992. Effects of forest management

on soil carbon storage. Water, Air and Soil Pollution.

64:83-120.

Johnson, K.H.; Vogt, K.A.; Clark, H.J.; Schmitz, O.J.;

Vogt, D.J. 1996. Biodiversity and the productivity

and stability of ecosystems. TREE. 11: 372-277.

Johnson, M.G. 1995. The role of soil management in

sequestering soil carbon. Pp. 351-363. In: Lal R.;

Kimble J.; Levine E.; Stewart B.A. (eds.). Soil

Management and Greenhouse Effect. Advances in

Soil Science. CRC Press & Lewis Publishers. Boca

Raton, Fl.

Karieva, P. 1994. Diversity begets productivity.

Nature 368:686-687.

Lawrence, G.; Vogt, K.A.; Vogt, D.J.; Tilley, J.; Wargo,

P.M.; Tyrrell, M. 1999. Chapter 5. In: Responses

of Northern Forests to Environmental Change.

Springer-Verlag. New York.

Lugo, A.E.; Sanchez, M.J.; Brown, S. 1986. Land use

and organic carbon content of some subtropical

soils. Plant Soil. 96:185-196.

Mangel, M.; Talbot, L.M.; Meffe, G.K.; Tundi Agardy,

M.; Alverson, D.L.; Barlow, J.; Botkin, D.B.;

Budowski, G.; Clark, T.; Cooke, J.; Crozier, R.H.;

Dayton, P.K.; Elder, D.L.; Fowler, C.H.; Funtowicz,

S.; Giske, J.; Hofman, R.J.; Holt, S.J.; Kellert, S.R.;

Kimball, L.A.; Ludwig, D.; Magnusson, K.,

Malayang, B.S.I.; Mann, C; Norse, E.A.; Northridge,

S.P.; Perrin, W.F.; Perrings, C; Peterman, R.M.;

Rabb, G.B.; Regier, H.A.; Reynold, J.E.I.; Sherman,

K.; Sissenwine, M.P.; Smith, T.D.; Starfield, A.;

Taylor, R.J.; Tillman, M.F.; Toft, C; Twiss, J.R.J.;

Wilen, J.; Young, T.P. 1996. Principles for the

conservation of wild living resources. Ecological

Applications 6:338-362.

Northrup, R.R.; Yu, Z.; Dahlgren, R.A.; Vogt, K.A.

1995. Polyphenol control of nitrogen release from

pine litter. Nature. 377:227-229.

Oades, J.M. 1988. The retention of organic matter in

soils. Biogeochemistry. 5:35-70.

Parladé, J.; Alvarez, I.F. 1993. Coinoculation of

aseptically grown Douglas-fir with pairs of

ectomycorrhizal fungi. Mycorrhiza. 3: 93-96.

Shaw, C.G. III. 1980. Characteristics of Armillaria

mellea on pine root systems in expanding centers of

root rot. Northwest Science. 54:137-145.

Shortle, W.C.; Smith, K.T.; Minocha, R; Minocha, S.;

Wargo, P.M.; Vogt, K.A. 1999. Chapter 3. In:

Responses of Northern Forests to Environmental

Change. Springer-Verlag. New York.

SmartWood Certification Program. 1997. Northeast

Regional Guidelines for the Assessment of Natural

Forest Management. The Rainforest Allinace's Smart

Wood Program. Richmond, Vermont. 22 pp.

Spies, T.A.; Franklin J.F. 1991. The structure of

natural young, mature, and old-growth forests in

Washington and Oregon. Pp. 90-109. In: Ruggiero

L.F.; Aubry K.B.; Carey A.B.; Huff M.H. (eds.).

Wildlife and Vegetation of Unmanaged Douglas-fir

Forests. USDA For. Serv. Gen. Tech. Rep. PNWGTR-285. Portland, Oregon.

Stevenson, F.J. 1965. Origin and distribution of

nitrogen in the soil. In: Bartholemew W.V.; Clark

F.E. (eds). Soil Nitrogen. Am. Soc. Agronomy.

Madison, Wl.

Tilman, D.; Downing J.A. 1994. Biodiversity and

stability in grasslands. Nature. 367:363-365.

Upton, C; Bass S. 1996. The Forest Certification

Handbook. St. Lucie Press. Delray Beach, Florida.

219 pp.

U.S. Environmental Protection Agency. 1996.

Proposed Guidelines for Ecological Risk Assessment.

August 1996, EPA/630/R-95/002B. US EPA,

Washington, DC.

Vitousek, P.M. 1994. Beyond global warming;

ecology and global change. Ecology 75:1861-

1902.

Vogt, K.A.; Edmonds, R.I.; Grier, C.C. 1981.

Dynamics of ectomycorrhizae in Abies amabilis

stands: the role of Cenococcum graniforme [Sow.]

Fred. & Winge. Holarctic Ecology 4:167-173.

Vogt, K.A.; Gordon, J.C.; Wargo, J.P.; Vogt, D.J.;

Asbjornsen, H.; Palmiotto, P.A.; Clark, H.J.; O'Hara,

J.L.; Keeton, W.S.; Patel-Weynand, T.; Witten, E.;

and others. 1997. Ecosystems. Balancing Science

with Management. Springer. New York. 470 pp.

Vogt K.; Vogt D.; Brown S.; Tilley J.; Edmonds R.;

Silver W.; Siccama T. 1995. Dynamics of forest

floor and soil organic matter accumulation in boreal,

temperate, and tropical forests. Pp. 159-178. Ch.

14. In: Lai R.; Kimble J.; Levine E.; Steward B.

19
(eds.). Soil Management and Greenhouse Effects.

CRC Press. Lewis Publishers. Boca Raton, Florida.

Vogt, K.A.; Vogt, J.D.; Palmiotto P.A.; Boon, P.;

O'Hara, J.; Asbjornsen H. 1996. Review of root

dynamics in forest ecosystems grouped by climate,

climatic forest type and species. Plant Soil. 187:

159-219.

Wargo, P.M. 1979. Defoliation-induced chemical

changes in sugar maple roots stimulate growth of

Armillaria mellea. Phytopathology. 62:1278-1283.

Wargo, P.M. 1981. Defoliation and secondary-action

organism attack: with emphasis on Armillaria

mellea. Journal of Arboriculture. 7:64-69.

Wargo, P.M.; Bergdahl, D.R.; Tobi D.R.; Olson C.W.

1993. Root vitality and decline of red spruce. In:

Fuhrer E.; Schutt P. (eds.). Contributiones Biologiae

Arborum Vol. 4. Ecomed. Landsberg/Lech. Germany

134 pp.

West N.E. 1982. Intermountain salt-desert

shrublands. Pp. 375-398. In: Ecosystems of the

World. Elsevier Scientific. Amsterdam.

Wijewardana, D.; Caswell S.J.; Palmberg-Lerche C.

1997. Criteria and Indicators for Sustainable Forest

Management. FAO XI World Forestry Congress.

Turkey. 13-22 October 1997.

20
SOIL QUALITY AND HEALTH-SOME APPLICATIONS TO ECOSYSTEM HEALTH AND

SUSTAINABILITY

Robert T. Meurisse1

ABSTRACT

Sustainability of forest and rangeland ecosystems is intricately linked with healthy and productive soils. Soils vary

widely in their quality. Their health is a condition of the soil relative to their inherent quality and ability to perform vital

soil and ecosystem functions. Soils in Pacific Northwest forest and rangeland ecosystems are particularly

heterogeneous. For example, selected soils from six Ecoregions, have organic carbon contents that range from about

40,000 to more than 800,000 kg/ha. Total nitrogen content ranges from less than 2,000 to more than 50,000 kg/ha.

There is a close relationship between soil quality and health and ecosystem quality and health. Ecosystem stresses can

be increased by management activities that compact soils and alter soil moisture and nutrient status and processes.

Forest health in Inland Western forests is intricately linked with carbon and nutrient cycling. A high percentage of soil

organisms make their home in the soil and play a critical role in the functions of ecosystems. Alteration of soil physical

and nutrient conditions can alter the incidence of soil borne diseases which, in turn, affect the health of ecosystems.

Specific quantitative soil standards, or thresholds, are needed to determine detrimental conditions. The National Forests

have established a variety of measures of detrimental conditions. A variety of physical measures, such as bulk density

changes, displacement and erosion, and detrimental burn conditions are emphasized.

Soils vary widely in their resiliency, or ability to rebound from impacts. A model of soil resiliency adopted from

Szabolcs (1994), includes physical, chemical and biological buffering and pedological and anthropological soil fluxes.

Several soils from six Pacific Northwest Ecoregions are rated for their resiliency. Criteria for chemical, biological and

physical buffering is included.

Keywords: soil quailty, soil health, soil functions, soil resiliency, ecosystem health, ecoregion

INTRODUCTION

Concerns about sustaining the productivity and health

of forest ecosystems have been the subject of much

research, debate, and federal actions (Perry and others

1989; Gessel and others 1990; Everett and others

1994; Harvey 1994; Denniston 1995; Jurgensen and

others 1997). Forest and livestock grazing

management practices and their impacts on the

environment are scrutinized daily by the media. A

committee of the National Research Council was

convened to examine the scientific basis for

inventorying, classifying and monitoring rangelands. It

included a strategy for evaluating rangeland health.

Recent flooding and landslide events in the west have

resulted in debates about watershed health. The

decline of salmonids in northwestern streams is

attributed to declines in watershed health, among other

factors. Although often overlooked, soil conditions and

soil degradation often underlie the concerns about

ecosystem health and attempts to restore ecosystem

health must consider and understand soil conditions

and processes (Meurisse and Geist 1994). Soil quality

and health, and methods for assessing them, have been

the subject of several symposia and publications

(Griffith and others 1990; Doran and others 1994;

Greenland and Zabolcs 1994; Doran and Jones 1996).

Soil health and quality are implicitely embedded in

statute for the National Forests and Grasslands of the

United States. But what do we mean by ecosystem

health, watershed health, soil health, and soil quality?

When there is not even complete agreement about

human health (Clements and Hales 1997), can we

reach consensus about soil health or ecosystem health?

My objectives are to define soil quality and health and

link them to vital soil functions. Secondly, I will

describe and illustrate some relationships between soil

health and ecosystem health, including a discussion of

soil quality standards as thresholds for soil health.

Thirdly, I will discuss soil resiliency concepts, some

criteria and ratings of soils in Pacific Northwest

1

 Robert T. Meurisse is Regional Soil Scientist, Retired, USDA Forest Service, Pacific Northwest Region, Portland, OR 97208

21
ecosystems and the importance of resiliency to

restoration of ecosystem health and sustainability.

REASONS FOR CONCERN ABOUT SOIL

QUALITY AND SOIL HEALTH

Soil and Civilizations

"For humankind, soil is the essence of life and health."

This opening sentence of the Preface to the Soil

Science Society of America Publication, Defining Soil

Quality for a Sustainable Environment, captures the

significance of the soil resource for sustainability of

human life and economic well being (Doran and others

1994). The importance of definition of soil quality and

establishment of quantitative indicators or thresholds

can not be overemphasized (Doran and Parkin 1994).

Definition is important for communication.

Establishment of quantitative measures is critical for

monitoring conditions. The importance of sound soil

and water management in the development and

advancement of civilization, particularly in the Middle

East, is described by Hillel (1994).

Soil Degradation

Worldwide soil degradation is estimated to be nearly

two billion ha. (Szabolcs 1994). Principal causes are

from water and wind erosion, with lesser amounts from

chemical and physical causes. Asia and Africa are

estimated to have more than 700 and 400 million

hectares of degradation, respectively. South America is

third with about 200 million hectares of degradation.

Although Mexico and Central America and North

America are among the least degraded of all Regions,

many millions of hectares are degraded. In North

America, there is little measurable amount estimated

for physical and chemical degradation. Studies have

shown that many areas have reduced productivity and

quality largely because of erosion and physical

degradation from soil compaction (O'Laughlin and

Pearce 1984; Froehlich and McNabb 1984; Sullivan

1988; Berg 1988; Geist and others 1991). Most of the

degradation in forest systems is due to erosion and

physical degradation, namely soil compaction (Sullivan

1988; Froehlich and McNabb 1984) and erosion

(Megahan 1981; O'Laughlin and Pearce 1984; Berg

1988). On rangelands, erosion and sedimentation is the

principal source of degradation (Smith 1989).

SOIL QUALITY AND HEALTH-WORKING

DEFINITIONS

Terms such as soil quality, soil health, forest health,

and ecosystem health, are anthropomorphic concepts

with relative values, just as human health is. The

terms soil quality and health often are used

interchangeably. However, subtle differences exist

between them. A 1995 National Cooperative Soil

Survey Soil Quality Committee defined soil quality as:

"the capacity of a specific soil to function, within

natural or altered land use boundaries, to sustain or

improve plant and animal productivity, water, air

quality, and human health and habitation." This

definition is similar to one defined by Doran and Parkin

(1994). Pedogenesis imparts specific qualities to soils

with resultant differences in their inherent capacity to

function. The American College Dictionary defines

health as "freedom from disease or ailment; the general

condition of the body or mind with reference to

soundness or vigor. So, I define soil health as "the

condition of the soil with reference to its inherent

quality and ability to perform vital ecosystem

functions." The distinction is the capacity to function

relative to a soils' inherent qualities.

Relationship of Soil Quality and Health to Vital

Soil Functions

Definitions of soil quality and health are linked to the

ability to perform vital soil functions. There is general

agreement among soil scientists about the major

functions of soils, though they may be stated in

different ways. I characterize them as follows: (1)

Sustain biological productivity, activity, and diversity;

(2) store and cycle nutrients and other materials; (3)

partition water, energy and solute flow; (4) filter,

buffer, immobilize, and detoxify organic and inorganic

materials; and (5) support structures and protect

archeological treasures. Therefore, we can say a soil is

unhealthy if soil properties are altered to the extent

that vital soil functions are impaired to some

measurable degree. Further, where specific soil quality

standards, or thresholds, are established, and those

standards are exceeded so that the ability to perform

vital functions is impaired, we can consider the soil as

unhealthy.

Soil properties that affect the capacity to function vary

depending on the specific function. For example,

properties important to productivity or biological

activity usually will be somewhat different from those

functioning to detoxify, buffer or immobilize materials.

But some properties are common to multiple functions.

Properties such as organic carbon, total nitrogen,

porosity, soil depth, particle size distribution, acidity,

and ion exchange capacity are examples of properties

that are critical to multiple soil functions, though their

relative values for a given function may be significantly

different. Additionally, soil moisture and temperature

regimes are important soil variables contributing to the

capacity of a soil to function. Soil quality generally can

be judged by soil properties that regulate the ability to

store and cycle nutrients and carbon, store and release

22
water, and provide favorable gas and heat exchange for

specific functions.

Soils vary widely in their quality and capacity to

perform vital functions. This is particularly true in the

Pacific Northwest where there is extreme heterogeneity

in the factors of soil formation and soil forming

processes. In particular, rates and magnitudes of

carbon and nitrogen cycling and accumulation vary

widely (Meurisse and others 1991). For example, soil

organic carbon ranges from about 40,000 kg/ha in

some Interior Northwest soils to nearly a million kg/ha

in some coastal Andisols (table 1). Similarly, total

nitrogen ranges from less than 2,000 to more than

50,000 kg/ha. Variation within a soil series may be

related to presence or absence of N-fixing species such

as snowbrush (Ceanothus velutinus). This is illustrated

for the Lapine series (table 1). The coastal Andisols

contain some of the highest amounts of organic carbon

in the world (Homann and others 1995). Those soils

also are among the most productive in the world

(Meurisse 1988a). Shallow and moderately deep

Mollisols contain amounts of organic carbon and

nitrogen comparable to deep and very deep Andisols in

Ecoregion M242C and M332G (table1). Grasses and

forbs are the dominant lifeform on these soils.

SOME SOIL HEALTH AND ECOSYSTEM

HEALTH CONNECTIONS

There is a close relationship between soil quality and

health and the quality and health of forest, rangeland,

and watershed ecosystems. Forest health usually is

considered in terms of insects and diseases and their

effects on tree mortality. These often are only

symptoms to more underlying stresses regulated by

soil/climate systems. Such stresses often are the result

of some management action. Management activities

that change soil moisture and nutrient status through

soil compaction, displacement and severe burning often

are major contributors to stresses that cause mortality

(Everett and others 1994). For example, the effects of

soil compaction and related loss of soil structure on

tree establishment and growth is well documented

(Froehlich and McNabb 1984; Clayton and others

1987; Powers and others 1990). Compaction increases

soil strength, decreases root penetration, decreases

moisture and nutrient availability and decreases gas

exchange. These all contribute to plant stress and loss

of vigor. Stress induced in part of an ecosystem

permeates the entire system (Waring 1985).

Moderate and severe soil compaction significantly

reduced nonmycorrhizal root tip abundance on both

Douglas-fir and western white pine after one year of

outplanting in northern Idaho (Amaranthus and others

1996). Similarly, ectomycorrhizal root tip abundance

and diversity was reduced on Douglas-fir seedlings in

severely compacted soils. These unhealthy soil

conditions directly effect forest health and

sustainability by directly affecting vital functions of

productivity, diversity, and biological activity.

Management activities affect long-term productivity

and sustainability of forest ecosystems by altering soil

conditions and processes. Where soil health is

maintained or improved, productivity potential will be

maintained or improved. When soil functions are

imbalanced and soil health degraded, plant vigor and

productivity is reduced (Waring 1985). Several

symposia have addressed relationships between

management activities, soil conditions and processes

and long-term productivity of Western Montane and

Northwest forest ecosystems (Perry and others 1987;

Gasboro and Slaughter 1988; Gessel and others 1990;

Harvey and Neunschwander 1991). Jurgensen and

others (1997) provide a comprehensive and critical

review of timber harvesting impact on soil organic

matter, nitrogen, productivity and health of Inland

Northwest Forests. They conclude that maintenance of

adequate soil organic matter levels is critical for

sustaining forest health and productivity under the

variable soil moisture and temperature conditions of the

region. Soil organic matter plays a major role in forest

soils and their productivity, activity and diversity and is

a major factor in the global carbon cycle (McFee and

Kelly 1995). Detrimental displacement, erosion, severe

burns, or acidification results in loss of organic matter

and nutrients and affects the ability of soils to replace

losses because of decreased plant productivity.

Soils vary widely in carbon sequestration processes and

magnitudes. Similar variations occur with nutrients,

especially nitrogen (table 1). The flow of energy

through the soil system is intricately linked with carbon

mineralization, or decomposition. Even though many

PNW soils have high reservoirs of total N, most PNW

forest sites are nitrogen deficient (Chappell and others

1992; Moore and others 1994). However the

deficiency often is related to availability rather than

total supply. Forest health in Inland Western forests is

intricately linked with carbon and nutrient cycling.

Constrained cycling, due to fire suppression and

increased activities of endemic insects and diseases,

are critical factors in these stressed ecosystems

(Harvey 1994).

Soil organisms are increasingly the subject of study for

their effects on forest soils and forest health. Soils

have a high degree of biological diversity (Richards

1987; Allen 1991; Molina and Amaranthus, 1991). A

high percentage of all organisms make their home in

the soil. There is an evolving body of literature defining

ecological linkages between soil processes, plant

growth, and community dynamics in the rhizosphere

23
Table 1 —Classification, depth, organic carbon (OC) and total nitrogen (TN) contents of selected soils in Pacific Northwest ecosystems. D

Soil Survey laboratory characterization. Except, Tolo, Klicker and Helter data from Geist and Strickler 1987).

SECTION

a

M242A

M242B

M242C

M261A

M332G

M333A

SERIES

KLISTAN

SLICKROCK

BOISTFORT

MINGPOINT

SALANDER

KILCHIS

HEMBRE

CINEBAR

COLTER

LASTANCE

ASCHOFF

LAPINE (PUTR)

LAPINE (CEVE)

SHANAHAN

NARD

DESCHUTES

JOSEPHINE

TOLO

KLICKER

HELTER

FIVEBIT

LIMBERJIM

NEVINE

RATHDRUM

SOIL GREAT GRP

HAPLUDANDS

FULVUDANDS

FULVUDANDS

HAPLUDANDS

FULVUDANDS

HAPLUMBREPTS

HAPLUMBREPTS

HAPLOXARANDS

VITRICRYANDS

HAPLOCRYODS

HAPLUMBREPTS

VITRICRYANDS

VITRICRYANDS

VITRICRYANDS

HAPLOXERALFS

HAPLOXEROLLS

HAPLOXERULTS

VITRIXERANDS

ARGIXEROLLS

VITRICRYANDS

HAPLOXEROLLS

UDIVITRANDS

VITRIXERANDS

VITRIXERANDS

FAM PARTICLE SIZE

MEDIAL-SKELETAL

MEDIAL/LOAMY

MEDIAL/CLAYEY

MEDIAL/CLAYEY

MEDIAL

LOAMY-SKELETAL

FINE-LOAMY

MEDIAL

ASHY-PUMICEOUS

LOAMY-SKELETAL

LOAMY-SKELETAL

ASHY-PUMICEOUS

ASHY-PUMICEOUS

ASHY/LOAMY

FINE-LOAMY

COARSE-LOAMY

FINE-LOAMY

ASHY/LOAMY

LOAMY-SKELETAL

ASHY/LOAMY

LOAMY-SKELETAL

ASHY/LOAMY-SKLTAL

ASHY/LOAMY-SKLTAL

ASHY

DEPTH

D&VD

VDEEP

VDEEP

VDEEP

VDEEP

SHALLOW

DEEP

VDEEP

DEEP

D&VD

VDEEP

VDEEP

VDEEP

VDEEP

DEEP

MDEEP

DEEP

VDEEP

MDEEP

D&VD

SHALLOW

DEEP

DEEP

VDEEP

OC (10

3

 kg/ha)

318

806

582

521

812

133

327

269

73

70

90

42

55

61

97

37

96

62

67

124

68

57

82

41

TN (10

2

 kg/ha)

191

536

233

385

323

122

210

282

72

17

44

18

24

23

n/a

50

73

38

37

80

33

50

47

18

MOIS

U

U

U

U

U

U

U

XE

U

U

U

XE

XE

XE

XE

XE

XE

XE

XE

XE

XE

U

XE

XE

Ref. McNab, W. Henry, and P.E. Avers. 1994.
(Molina and Amaranthus 1991). Mychorrizae affect soil

structure by producing humic compounds. Interrupting

flow of organic materials to mychorrizae can cause a

deterioration in soil structure and aggregate stability,

and reduce forest regeneration and root growth

(Amaranthus and others 1990). Healthy populations of

mychorrizal fungi and other soil microbes are essential

for the growth and survival of tree seedlings,

particularly on droughty and nutrient-poor sites.

Soil microorganisms play a critical role in the functions

of ecosystems. They are particularly important for

nutrient and carbon cycling and energy flow, nitrogen

fixation, and nutrient and moisture supply through

mychorrizal symbiosis (Richards 1987; Amaranthus and

others, 1990; Allen 1991: Harvey and others 1994).

Habitat provided by variable soil conditions influences

the species and populations of soil biota. Significant

changes in the habitat from management practices can

greatly alter micro flora and fauna species and their

abundance (Wells and others 1979; Amaranthus and

others 1990; Page-Dumroese and others 1991). Such

changes can significantly change the vital soil function

of carbon and nutrient storage and cycling with

subsequent effects on plant performance (Molina and

Amaranthus 1991; Harvey and others 1994: Jurgensen

and others 1997).

Soil borne diseases may be increased when soil health

is impaired by compaction and changes in soil nutrient

and moisture conditions. Wilks and others (1983)

reported increases in spread of black stain in ponderosa

pine under impaired aeration and high redox potentials.

Such conditions can occur from soil compaction.

Unpublished surveys and personal observations have

revealed black stain in Douglas-fir and in ponderosa

pine in compacted soils. Changes in nutrient status,

especially nitrogen, may affect incidence of other

diseases such as laminated root rot (Phellinus weirii

(Murr.) Gilb.) Nelson (1975) found lower incidence with

high dosages of nitrogen fertilizer. He attributed the

results to antagonists of P. weirii. Similarly, Mika and

others (1992) suggest that stresses from moisture and

nutrient deficiencies can result in increased tree

mortality from diseases. In particular, tree root

chemistry, altered by thinning and N-fertilization,

stimulated incidence of Armillaria ostoyae (Entry and

others 1991). However, fertilization with potassium

resulted in decreased incidence of Armillaria (Moore

and others 1994). Thus, nutrient balance is important

for maintaining healthy soils and healthy ecosystems.

Watershed Health

Watershed health has a direct linkage with soil health.

Hydrologic function and water quality is impaired when

soils are compacted and eroded. The literature is

replete with studies that document effects from

silivicultural practices, land use and fire on soil and

water resources (Megahan, 1981; O'Loughlin and

Pearce, 1984; Berg, 1988 ). Practices that significantly

decrease infiltration and increase volume and timing of

runoff, affect the manner in which water and energy

are distributed within a watershed. Such effects can

lead to increases in peak flows and accelerated erosion,

including gullies and landslides (Swanson and Dyrness

1975; Jones and Grant 1996). When soil compaction

exceeds about 12 percent of a watershed, significant

changes in peak flows have been reported in Western

Oregon (Harr 1976). Accelerated erosion from poorly

designed or improperly applied practices decreases

water quality and productivity (Megahan 1981;

O'Laughlin and Pearce 1984; Berg 1988). Properly

functioning watersheds are dependent on properly

functioning, healthy soil systems.

The connections between forest and rangeland

ecosystems and watersheds are dynamic and complex.

They are highly variable in their qualities and in their

conditions or health. Fundamentally, all are underlain

and regulated largely by the quality and health of the

soil ecosystem and its complex, dynamic web of

properties, processes and functions.

SOIL QUALITY STANDARDS

It is national Forest Service policy to have soil quality

standards to ensure sustainability and long-term

productivity of forest ecosystems. Several Regions

have developed and implemented standards, or

measures, of soil quality since the 1970's. The general

concepts and standards are described by (Griffith and

others 1990). Specific standards vary by region. The

Pacific Northwest Region developed the first

comprehensive set of quantitative soil quality standards

and a procedure for measuring soil physical conditions.

These procedures are described by Meurisse (1988b)

and Geist and others (1991). Sullivan (1988)

presented some of the most complete data that

illustrates the use of the sampling methods and the

effects of some management practices on soil

conditions. Understanding the role of soils is crucial to

understanding stress processes within ecosystems and

establishment of measurable soil quality standards is a

means of expressing desired soil conditions (Meurisse

and Geist 1994).

The most common measures of soil quality used on the

National Forests are various physical properties. There

are others such as erosion, organic matter, and degree

of burning. A general description of the common

measures follows.

Physical measures: Bulk density, porosity, soil

displacement, infiltration, rutting and puddling.

25
Erosion measures: These are more indirect and include

soil loss tolerance, some specified percentage of topsoil

loss, effective soil cover, and some percent of forest

floor removal.

Organic matter measures: X percent loss in surface

layers, amount of organic matter less than some

specified quantity, amount of large woody debris, and

large woody debris sufficiency.

Detrimental burning: This is a measure of loss of O

horizons and signs of oxidation, change in soil color at

the surface together with the next cm depth blackened.

Specific quantitative standards, or thresholds,

determine detrimental conditions. The detrimental

conditions are presumed to be "unhealthy" soil

conditions because they are impaired in their ability to

perform one or more of the vital functions. In most

cases, functions of concern are productivity and

diversity, water quality and changes in water and

energy flows, and effects on nutrient and carbon

cycles.

SOIL RESILIENCE

The concept of resilience is fundamental to assessing

soil quality or soil health. Resilience is particularly

important for considering restoration priorities and

treatments. Resilience is defined variously as the ability

to rebound or recover from some condition or stress.

Szabolcs (1994) defines resilience as follows:

SR = BCph + BCch + BCb + t1

t2

 dPSF/dt + t1

t2

 dASF/dt

Where: SR = Soil resilience

BCph = Physical buffering

BCch = Chemical buffering

Bcb = Biological buffering

PSF = Pedological soil fluxes

ASF = Anthropological soil fluxes

Such an approach to resilience can contribute not only

to its interpretation, but to its modelling and estimation

through further studies. Because the concept of

resilience is crucial to characterization of soil health,

interpretations of soil resilience need to be made.

Some of the important dynamic factors that contribute

to resilience are organic carbon, soil nutrients,

especially nitrogen, and soil structure. Other intrinsic

properties include moisture and temperature regimes,

depth, particle size distribution and permeability. The

relative importance of the factors vary depending on

whether one is rating for chemical, biological or

physical buffering. Thus a soil may have a different

rating for physical buffering than for chemical or

biological buffering. Soil organisms play crucial roles

because of their function in pedoturbation and carbon

and nitrogen cycling.

Soil resilience ratings for selected soils from several

broad ecoregions are displayed in table 2. Rating

classes of very high to low are shown. For chemical

and biological buffering, organic carbon and total

nitrogen contents are major factors, together with

moisture and temperature regimes and soil depth (table

3). Organic carbon, family particle size and moisture

and temperature regimes are important criteria for

physical buffering (table 4). Because the criteria for

chemical, biological and physical buffering include

several pedological properties, I considered the

buffering ratings to be equivalent to resiliency ratings.

These factors that are important for resilience also

contribute directly to soil quality.

Table 2—Soil resiliency ratings as a function of chemical/

biological and physical buffering, for selected soils in six

Pacific Northwest Ecoregions.

SECTION

M242A

M261A

M242B

M242C

M332G

M333A

SERIES

KLISTAN

SLICKROCK

BOISTFORT

KILCHIS

JOSEPHINE

LASTANCE

ASCHOFF

LAPINE (PUTR)

LAPINE (CEVE)

DESCHUTES

TOLO

KLICKER

HELTER

NEVINE

RATHDRUM

CHEM/BIOLOGICAL

HIGH

VERY HIGH

VERY HIGH

MOD LOW

MODERATE

LOW

MOD LOW

LOW

LOW

MOD LOW

MOD LOW

MOD LOW

MODERATE

MODERATE

MOD LOW

PHYSICAL

MODERATE

VERY HIGH

HIGH

LOW

MODERATE

MOD LOW

MOD LOW

MODERATE

MODERATE

LOW

HIGH

MOD LOW

MODERATE

MODERATE

MODERATE

Restoration of stressed sites, and ecosystem

processes, is an integral component of sustaining forest

ecosystem health (Everett 1994). An understanding of

soil resiliency is critical for implementing cost-effective

restoration programs. Generally, the most resilient

soils are those from the Oregon and Washington Coast

Ranges. The most resilient are Andisols (Slickrock,

Boistfort), with organic carbon content in excess of

500,000 kg/ha and 30,000 kg/ha total nitrogen

(table1). Soils with less than 70,000 kg/ha organic

carbon and 5,000kg/ha total nitrogen are rated

moderately low oil resiliency is critical for implementing

cost-effective restoration programs. Generally, the

26
Table 3 —Resiliency criteria for chemical and biological buffering in selected Pacific Northwest soils.

Rating

Very High

High

Moderate

Mod. Low

Low

OC

(thous kq/ha)

500 plus

200-500

70-200

30-70

< 30

TN

(thous kg/ha)

30 plus

20-30

5-20

3-5

<3

Depth

Class

D & VD

D & VD

MD & D

MD & SH

SH & MD

Moisture

Regime

Udic

Udic

Xeric/Udic

Xeric/Udic

Aridic/Xeric

Temperature

Regime

Iso-mesic,

Mesic

Iso-mesic,

Mesic, Frigid

Mesic, Frigid

Cryic

Mesic, Frigid

Cryic

Cryic/Frigid

Mesic

SH = Shallow; MD= Moderatley Deep; D = Deep; VD = Very Deep

0C = Organic Carbon; TN - Total Nitrogen

Table 4—Resiliency criteria for physical buffering in selected Pacific Northwest soils.

Rating

Very High

High

Moderate

Mod. Low

Low

OC

(thous kg/ha)

500 plus

200-500

70-200

30-70

< 30

Family

Particle Size

Medial, Loamy

Medial/Loamy

Medial/Clayey

Ashy/Loamy

Ashy-Pumiceous

Fine-Loamy

Ashy/L-Skeletal

Medial-Skeletal

Loamy-Skeletal

Skeletal

Coarse-Loamy

Depth

Class

D & VD

D & VD

MD & D

MD&SH

SH & MD

Moisture

Regime

Udic

Udic

Xeric/Udic

Xeric/Udic

Xeric/Aridic

Temperature

Regime

Iso-mesic,

Mesic

Iso-mesic,

Mesic, Frigid

Mesic, Frigid

Cryic

Mesic, Frigid

Cryic

Mesic, Cryic

SH = Shallow; MD= Moderatley Deep; D = Deep; VD = Very Deep

OC = Organic Carbon

27
most resilient soils are those from the Oregon and

Washington Coast Ranges. The most resilient are

Andisols (Slickrock, Boistfort), with organic carbon

content in excess of 500,000 kg/ha and 30,000 kg/ha

total nitrogen (table 1). Soils with less than 70,000

kg/ha organic carbon and 5,000 kg/ha total nitrogen

are rated moderately low and low for chemical and

biological buffering. Some of these same soils, Tolo,

Lapine, Helter, Nevine, are rated moderate to high for

physical buffering, or resiliency. Priorities for

restoration treatments should be based on knowledge

of the soil condition and soil resiliency. Those that are

most resilient, may recover without costly treatments.

SUMMARY

A key to managing the soil resources to sustain their

health, is to have knowledge of the soils and their

behavior in response to management practices. This

includes having knowledge of pedogenic processes and

carrying capacity of the soil. Development of

quantitative soil quality standards, assessment of soil

health conditions and application of sound management

practices is essential for sustainable development and

survival of civilizations.

Sustainable, healthy forest and rangeland ecosystems

and watersheds are intricately linked with healthy soil

conditions. An understanding of the basic soil

ecosystem properties and processes, including soil

resiliency, is essential to sustaining soil quality and soil

health. Life and civilization depends upon sustaining

the quality of soils and their ability to perform their vital

functions.

QUESTIONS FROM THE SYMPOSIUM

FLOOR

Q: Microfauna can survive higher temperatures under

dry conditions, then is fall burning preferable to spring

burning?

A: Not necessarily. It depends upon the conditions of

the soil and organic matter when ignited and it depends

on whether or not the burn is within prescription.

There simply are too many variables to make a

categorical statement that one season is better than

another.

Q: What are some field keys to look at in assessing

organic matter health for forests?

A: There are several simple keys. (1) Observe the

amount and degree of surface soil disturbance relative

to the natural condition of similar soils. Make

estimates of classes of disturbance based on depth and

extent of disturbance. (2) Consider the distribution of

surface organic matter (OM). Is the OM uniformly

distributed, or is it limited to a few concentrations? (3)

Consider whether the amount of OM appears to be

within the range of natural variability.

Q: At what point do we become concerned about OM

levels of a particular stand?

A: I think that if more than 15% of the mineral soil

organic matter is lost, there would be some loss in site

productivity. However, this seldom occurs under

current management regimes. I think that if surface

organic matter, including coarse woody debris, is far

outside the range in natural variability for the site, then

we should be concerned. Many of our forests have

considerably more surface OM than what was present

prior to the aggressive fire suppression activity of the

past 50 to 75 yrs. But where surface OM is reduced to

less than about 10 to 20 T/Ac, some short term

reductions in productivity can be expected. We need

to know and understand the nature of the soils and

vegetation at a given site or stand. But, too often we

make broad generalizations that may not be applicable.

Q: Can you discuss "case hardening" of down woody

debris (DWD) by fire? Do they decompose faster or

slower?

A: Unfortunately, I am aware of no substantive data

to answer this question. My observations suggest

there is slower decomposition, especially on relatively

dry sites common in the Interior Northwest. But I am

not sure if the same relationship exists on the very

moist West Cascade and Coast Range forests.

Q: Please estimate the fraction of PNW forests (total

public and private) that are managed according to

systems that have been demonstrated not to

cumulatively degrade them.

A: (I assume the question means not to degrade the

soils, which may be different from degrading the forest

per se). I have no data to make a credible estimate for

all forests. I can only make an estimate for the National

Forests, and it is only a rough estimate. I believe that

more than 90% of westside national forests are

managed without measurable soil degradation. On the

eastside, where ground impacting equipment is

dominant, multiple entries are common, most of the

monitoring that is done indicates there is commonly

about 20 to more than 30 percent in detrimental

condition prior to rehabilitation. Also, there are

increasingly frequent wildfires of moderate to high

intensity burns that have cumulative adverse impacts.

There is a lot of rehabilitation on national forests, so

the cumulative effect of actively managed forests,

probably is less than 20% degraded.

28
Q: Do you think the role of soils and biological activity

would receive more attention if there was a Clean Soil

Act-like Clean Water Act & Clean Air Act?

A: It probably would increase attention to the soil

resource, including soil biology. There has been some

discussion of promoting a Soil Quality Act, but it hasn't

progressed very far.

Q: Background: An eastside pine community, low to

mod. site; burned during Aug. 1994 - moderate to high

intensity; no burned area emergency rehab. Grass

seeding the next summer; several high intensity

summer thundershowers mobilized 0.5 - 6 inches of

soil. The area was salvage logged during summer of

95 & 96. Replanting took place in Spring of 96 & 97.

The current proposal is to decrease shrub competition

by 100% herbicide application. What are the

implications of the herbicide proposal to soil

ecosystem? What are the cumulative effects to the

soil ecosystem?

A: My initial answer is "I don't know." But, based on

what is emerging from levels of growing stock studies

at Pringle Falls Experimental Forest, removal of shrub

understory, regardless of whether it is done

mechanically or with herbicides, may have some longterm adverse affects on stand productivity. Also, some

of the shrubs likely are N-fixing species that would

normally add replacement nitrogen to the site that may

have been lost in the fire. Herbicide effects on soil

organisms probably would be short term and of low

significance. Further, loss of 0.5 to 6 inches of soil

over areas larger than 100 ft.2 probably would have

significant detrimental effects. So, I would expect

some measurable, cumulative adverse effects under

this scenario.

LITERATURE CITED

Allen, M.F. 1991. The ecology of mycorrhizae.

Cambridge University Press, 40 West 20th St. New

York, NY 1011-4211. 184 p.

Amaranthus, M.P.; Page-Dumroese, D.; Harvey, A.;

Cazares, E.; Bednar, L.F. 1996. Soil compaction

and organic matter affect conifer seedling

nonmycorrhizal and root tip abundance and

diversity. USDA Forest Service, Pacific NW Res.

Station, Portland, OR 97208. Res. Paper PNW-RP494. 12 p.

Amaranthus, M. P.; Molina, R.; Perry, D.A. 1990. Soil

organisms, root growth and forest regeneration. In:

Forestry on the Frontier-Proceedings National

Society of American Foresters National convention.

Spokane, WA. pp 89-93.

Berg, N.H. (tech. ed.). 1988. Proceedings of the

Symposium on Fire and Watershed Management.

October, 1988, Sacramento, CA. USDA Forest

Service, Pacific SW Forest and Range Exp. Stn.

General Tech. Report PSW-109. 164 p.

Chappell, H.N.; Omule, S.A.Y.; Gessel, S.P. 1992.

Fertilization in Coastal Northwest Forests: using

response information in developing stand-level

tactics. In: Forest fertilization: Sustaining and

improving nutrition and growth of western forests.

Chappell, H.N.; Weetman, G.F.; Miller, R.E. (eds.).

Institute of Forest Resources Contrib. 73. College of

Forest Resources, Univ. of Washington, Seattle, WA

98195. 302 p.

Clayton, J.L.; Kellogg, G.; Forrester, N. 1987. Soil

disturbance-tree growth relations in Central Idaho

clearcuts. USDA Forest Service, Ogden,UT.

Intermountain Research Station. Res. Note INT-372.

6 p.

Clements, M.; Hales, D. 1997. How healthy are we?

Report on American's health practice, beliefs, and

opinions-including some surprising findings from a

new national survey. In: Parade, September 7.

Denniston, D. 1995. Sustaining mountain peoples and

environments. In: State of the World. A

WorldWatch Institute Report on Progress Toward a

Sustainable Society. W. W. Norton and Co. New

York, NY 10110. 255 p.

Doran, J.W.; Coleman,D.C; Bezdecek, D.F. Stewart,

B.A. (eds.). 1994. Defining soil quality for a

sustainable environment. SSSA Special Publication

No. 35. SSSA, Inc., Madison, Wl. 244 p.

Doran, J.W ; Jones, A.J. (eds.). 1996. Methods for

assessing soil quality. SSSA /special Publication No.

49. SSSA, Inc. Madison, Wl. 410 p.

Doran, J.W.; Parkin, T.B. 1994. Defining and

assessing soil quality. In Doran, J.W.; Coleman,

D.C.; Bezdecek, D.F.; Stewart, B.A. (eds.). Defining

soil quality for a sustainable environment. SSSA

SpecialPublication No. 35. SSSA, Inc., Madison,

Wl. p 3-21.

Entry, J.A.; Cromack, K. Jr.; Kelsey, R.G.; Martin, N.E.

1991. Response of Douglas-fir to infection by

Armillaria ostoyae after thinning plus fertilization.

Phytopathology 81:682-689.

Everett, R. 1994. Eastside Forest Health Assessment.

Vol. IV: Restoration of Stressed Sites and

Processes. USDA Forest Service, PNW Res. Stn.,

Portland, OR 97208-3890. Gen. Tech. Rep. PNWGTR-330. 123 p.

29
Everett, R.; Hessburg, P.; Jensen, M.; Bormann, B.

1994. Eastside Forest Health Assessment. Vol. I:

Executive Summary. USDA Forest Service, PNW

Res. Stn., Portland, OR 97208-3890Gen. Tech.

Rep. PNW-GTR-317. 61 p.

Froehlich, H.A.; McNabb, D.S. 1984. Minimizing soil

compaction in Pacific Northwest Forests. In: Forest

Soils and Treatment Impacts. Stone, E.L. (ed.).

Proceedings of the Sixth North American Forest

Soils Conference, Univ. Tennessee, Knoxville. June,

1983. Dept. of Forestry, Wildlife and Fisheries, The

Univ. of Tenneessee, Knoxville, TN 37996-4500.

454 p.

Geist, J. M.; Strickler, G.S. 1978. Physical and

chemical properties of some Blue Mountain soils in

Northeast Oregon. USDA Forest Service, Portland,

OR 97208. Pacific Northwest Research Station,

Res. Paper PNW236. 19 p.

Geist, J. M.; Meurisse, R.T.; Max, T.A. 1991.

Monitoring forest soil properties to maintain and

productivity. In: Harvey, A.E.; Neuenschwander,

L.F. (eds.). Proceedings of a Symposium,

Management Productivity of Western Montane

Forest Soils, Boise, ID. April, 1990. Intermountain

Res. Stn., Ogden, UT. 8440. Intermountain Res.

Stn. General Technical Report INT-280. 254 p.

Gessel, S.P.; Lacate, D.S.; Weetman, G.F.; Powers,

R.F. (eds.). 1990. Sustained Productivity of Forest

Soils. Proceedings of the Seventh North American

Forest Soils Conference, Univ. British Columbia,

1988. Faculty of Forestry Pub, Vancouver, B.C.

525 p.

Greenland, D.J.; Szabolcs, I. 1994. Soil resilience and

sustainable land use. Proceedings of a Symposium,

October, 1992. Budapest, Hungary. CAB

International, Wallingford, Oxon OX 10 8DE, UK

561 p.

Griffith, R.W.; Goudey, C; Poff, R. 1990. Current

application of soil quality standards. ln:Proceedings

of the soil quality standards sysmposium. SSSA

meeting, San Antonio, TX, October 21-27, 1990.

USDA Forest Service, Watershed and Air

Management Staff. Washington, D.C. WO-WSA-2.

80 p.

Harr, D.R. 1976. Forest practices and streamflow in

Western Oregon. USDA Forest Service, Pacific NW

Forest and Range Exp. Stn. Portland, OR. General

Technical Report PNW-49. 18 p.

Harvey, A.E. 1994. Integrated roles for insects,

diseases and decomposers in fire dominated forests

of the Inland Western United States: Past, present

and future forest health. J. Sustainable For. 2:211-

220.

Harvey, A.E.; Neuenschwander, L.F. (eds.). 1991.

Proceedings-Management and Productivity of

Western Montane Forest Soils. Proceedings of a

Symposium, Management and Productivity of

Western Montane Forest Soils, Boise, ID, April,

1990. USDA Forest Service, Intermountain Res.

Stn., Ogden, UT 84401. Intermountain Res. Stn.

General Technical Report INT-280. 254 p.

Harvey, A.E.; McDonald, G.I.; Jurgensen, M.F.; Larsen,

M.J. 1994. Microbes: drivers of long-term

ecological processes in fire-influenced cedarhemlock-white pine forests in the Inland Northwest.

In: Baumgartner, D.A.; Lotan, J.E.; Tonn, J.R.,

(comp. and eds.). Proceedings, Interior cedarhemlock-white pine forests: ecology and

management. March 2-4, 1993, Spokane, WA.

Pullman, WA: Washington State University: 157-

163.

Hillel, D. 1994. Introductory overview: Soil, water,

and civilization. In: Soil and water science: Key to

understanding our global environment. Proceedings

of a Symposium. SSSA meeting, Cincinnati, OH,

Nov. 1993. SSSA Special Publication No. 41.

SSSA, Inc. Madison, Wl. 103 p.

Homann, P.S.; Sollins, P.; Chappell, H.N.;

Stangenberger, A.G, 1995. Soil organic carbon in

a mountainous, forested region: relation to site

characteristics. Soil Sci. Soc. Amer. J. 59:1468-

1475.

Jones, J.A.; Grant, G.E. 1996. Peak flow responses

to clear-cutting and roads in small and large basins,

western Cascades, Oregon. Water Res. Res. Vol.

32 (No. 4): 959-974.

Jurgensen, M.F.; Harvey, A.E.; Graham, R.T.; PageDumroese, D.S.; Tonn, J.R.; Larsen, M.J.; Jain, J.B.

1997. Impacts of timber harvesting on soil organic

matter, nitrogen, productivity, and health of Inland

Northwest Forests. Forest Science, Vol. 43, No. 2:

234-251.

McFee, W.W.; Kelly, J.M. (eds.). 1995. Carbon forms

and functions in forest soils. Proceedings of the

Eighth North American Forest Soils Conference,

Univ. FL, 1993. Soil Sci. Soc. Amer., Madison, Wl.

594 p.

McNab, W.H.; Avers, P.E. 1994. Ecological

subregions of the United States: Section

descriptions. Administrative Publication WO-WSA5. Wash. D.C. USDA Forest Service. 267 p.

30
Megahan, W.F. 1981. Effects of silvicultural practices

on erosion and sedimentation in the Interior West-A

case for sediment budgeting. ln:lnterior West

Watershed Management, Baumgartner, D. (compiler

and ed.). Proceedings of a Symposium, April 1980,

Spokane, WA. Wash. State Coop. Extension.

Pullman, WA 99164. 288 p.

Meurisse, R.T. 1988a. Soil Productivity and the

management of U.S. Andisols. In: Kinloch, D.I.;

Shoji, S.; Beinroth, F.H.; Eswaran, H. (eds.).

Proceedings of the Ninth International Soil

Classification Workshop, Japan, 20 July to 1

August, 1987. Publ. by Japanese Committee for the

9th International Soil Classification Workshop, for

the Soil Management Support Services,

Washington, D.C., USA.

Meurisse, R.T. 1988b. Soil productivity protection

and improvement: objectives, policy, and standards

in the Pacific Northwest Region of the Forest

Service. In: Slaughter, C.W.; Gasbarro, T. (eds.).

Proceedings of the Alaska Soil Productivity

Workshop, April, 1987. USDA Forest Service,

Portland, OR 97208-3890. Pacific NW Res. Stn.

General Technical Report, PNW-GTR-219. 120 p.

Meurisse, R.T.; Robbie, W.A.; Niehoff J.; Ford, G.

1991. Dominant soil formation processes and

properties in Western-Montane forest types and

landscapes-some imlications for productivity and

management. In: Harvey, A.E.; Neuenschwander,

L.F. (eds.). 1991. Proceedings-Management and

Productivity of Western Montane Forest Soils.

Proceedings of a Symposium, Management and

Productivity of Western Montane Forest Soils,

Boise, ID, April, 1990. USDA Forest Seervice,

Intermountain Res. Stn., Ogden, UT 84401.

Intermountain Res. Stn. General Technical Report

INT-280. 254 p.

Meurisse, R.T.; Geist, J.M. 1994. Conserving soil

resources. In: Everett, R. (comp.). Eastside Forest

Health Assessment. Vol. IV: Restoration of

Stressed Sites and Processes. USDA Forest Service,

PNW Res. Stn. Portland, OR 97208-3890. Gen.

Tech. Rep. PNW-GTR-330. 123 p.

Mika, P.G.; Moore, J.A.; Brockley, R.P.; Powers, R.F.

1992. Fertilization response by interior forests:

when, where, and how much. In: Forest fertilization:

Sustaining and improving nutrition and growth of

western forests. Chappell, H.N.; Weetman, G.F.;

Miller, R.E. (eds.). Institute of Forest Resources

Contrib. 73. College of Forest Resources, Univ. of

Washington, Seattle, WA 98195. 302 p.

Molina, R.; Amaranthus, M. 1991. Rhizosphere

biology: Ecological Linkages between soil processes,

plant growth, and community dynamics. In: Harvey,

A.E.; Neuenschwander, L.F. (eds.). Proceedings of

a Symposium, Management and Productivity of

Western Montane Forest Soils, Boise, ID. April,

1990. USDA Forest Service, Intermountain Res.

Stn., Ogden, UT 84401. Intermountain Res. Stn.

GeneralTechnical Report INT-280. 254 p.

Moore, J.A.; Mika, P.G.; Schwandt, T.W.; Shaw, T.M.

1994. Nutrition and forest health. In: Baumgartner,

D.M.; Lotan J.E.; Tonn, J.R. (eds.). Interior cedarhemlock-white pine forests: ecology and

management. Symposium proceedings. Washington

State University, Pullman, WA. 365 p.

Nelson, E.E. 1975. Survival of Poria Weirii in wood

buried in urea-amended forest soil. Phytopathology.

65:501-502.

O'Laughlin, C.L.; Pearce, A.J. (eds.). 1984.

Proceedings of a Symposium on Effects of Forest

Land Use on Erosion and Slope Stability. May,

1984. Environment and Policy Institute, East-West

Center, Univ. HA, Honolulu, HA. 310 p.

Page-Dumroese, D.; Graham, R.T.; Harvey, A.E. 1991.

Soil organic matter effects on Douglas-fir growth in

Northern Idaho soils. In: Harvey, A.E.;

Neuenschwander, L.F. (eds.). 1991. ProceedingsManagement and Productivity of Western Montane

Forest Soils. Proceedings of a Symposium,

Management and Productivity of Western Montane

Forest Soils, Boise, ID, April, 1990. Intermountain

Res. Stn. General Technical Report INT-280.

Intermountain Res. Stn., Ogden, UT 84401. 254 p.

Perry, D.A.; Meurisse, R.;Thomas, B.; Miller, R.; Boyle,

J.; Means, J.; Perry, C.R.; Powers, R.F. (eds.).

1989. Maintaining the long-term prodcutivity of

Pacific Northwest Forest Ecosystems. Proceedings

of a sysmposium, Maintaining the Long-term

Productivity of Pacific Northwest Forest

Ecosystems. In cooperation with the College of

Forestry, Oregon State Univ. April, 1987. Timber

Press, 9999 S.W. Wilshire, Portland, OR. 97225.

256 p.

Powers, R.F.; Alban, D.H.; Miller, R.E.; Tiarks, A.E.;

Wells, C.G.; Avers, P.E.; Cline, R.G.; Fitzgerald,

R.O.; Loftus, N.S. Jr. 1990. Sustaining site

productivity in North American Forests: Problems

and Prospects. In: Gessel, S.P.; Lacate, D.S.;

Weetman, G.F.; Powers, R.F. (eds.). Sustained

Productivity of Forest Soils. Proceedings of the

Seventh North American Forest Soils Conference,

Univ. British Columbia, 1988. Faculty of Forestry

Pub., Vancouver, B.C. 525 p.

31
Richards, B.N. 1987. The microbiology of terrestrial

ecosystems. John Wiley and Sons, Inc. New York,

NY 10158. 399 p.

Smith, L.E. 1989. Range condition and secondary

secession: A critique. In: Lauenroth, W.K.; Laycock,

W.A. (eds.). Secondary Succession and the

Evaluation of Rangeland Condition. Boulder, CO.

Westview.

Sullivan, T. 1988. Monitoring soil physical conditions

on a natioanl forest in eastern Oregon: A case

study. In: Slaughter, C.W.; Gasbarro, T. (eds.).

Proceedings of the Alaska Soil Productivity

Workshop, April, 1987. Pacific NW Res. Stn. General

Technical Report, PNW-GTR-219. 120 p.

Swanson, F.J.; Dyrness, C.T. 1975. Impact of clearcutting and road construction on soil erosion by

landslides in the western Cascade Range, Oregon.

Geology 3(7):393-396.

Szabolcs, I. 1994. The concept of soil resilience. In:

Greenland, D.J.; Szabolcs, I. (eds.). Soil esilience

and sustainable land use. Proceedings of a

Symposium, October, 1992. Budapest, Hungary.

CAB International, Wallingford, Oxon OX 10 8DE,

UK 561 p.

Waring, R.H. 1985. Imbalanced forest ecosystems:

assessments and consequences. Forest Ecology and

Management, 12:93-112. Elsevier Sci. Pub.

Amsterdam.

Wells, C.G.; Campbell, R.E.; DeBano, L.F.; Lewis, C.E.;

Fredricksen, R.L.; Franklyn, C.C.; Froelich, R.C.;

Dunn, P.H. 1979. Effects of fire on soil. A stateof-knowledge review. In: National Fire Effects

Workshop, 10-14 April 198, Denver CO. U.S. Forest

Service, Wash. Off. Gen. Tech. Rep. WO-7.

Wilks, D.S.; Gersper, P.L.; Cobb, F.W. Jr. 1983.

Relation of soil redox potential to infection of

ponderosa pine by Ceratocystis wageneri.

Phytopathology 73:1120-1125.

32
SOIL-DWELLING ARTHROPODS: THEIR DIVERSITY AND FUNCTIONAL ROLES

Andrew R. Moldenke1

ABSTRACT

Soil-dwelling arthropods are extremely diverse and abundant. Northwest forest soils may support an average of

250,000/m2. Knowledge of functional roles within the soil ecosystem is limited. Some of the prominent functional

groups are fungivores, micro-predators, macro-predators, and shredders.

Soil food webs involve complex functions and the basic structure requires a minimum of four components. These

components are (1) the living plant root, (2) soil bacteria and fungi, (3) fauna that graze on microbial biomes, and (4)

predator fauna. Some critical environmental determinants for ensuring persistence of functional links are described.

Three categories of determinants are epigeic habitats, soil surface phenomena, and hypogeic habitats.

Keywords: arthropods, keystone species, fungivores, predators, shredders, nutrient cycling, food web, epigeic habitats,

hypogeic habitats

INTRODUCTION

The majority of the world's arthropods live within the

soil (Southwood 1978). The vast majority of all animal

life are species of arthropods (Wilson 1992). This

amounts to a fabulously rich soil fauna in all terrestrial

ecosystems (Petersen and Luxton 1982). In the forests

of the western United States there may be as many as

several hundred species of arthropods (insects, mites,

spiders, centipedes, millipedes, skunk-spiders, and their

relatives) in every square meter of soil (Moldenke

personal observation). The identity and relative

abundance of these myriad creatures vary in response

to every climatic microhabitat. Thus, the species

richness encompassed within several square miles is at

least several thousand (Parsons and others 1991). The

great majority of these species are widespread

geographically, inhabiting territory encompassing

thousands to tens of thousands of square miles. If we

assume that these species could not coexist unless

several aspects of their biological functions were

unique, then the roles these creatures play must be

remarkably diverse (Moldenke and others 1994).

Not only are soil arthropods diverse, but they are

numerically significant as well. A single square meter of

Northwest forest soil may support an average of

250,000 arthropods (Moldenke 1990). The most

numerous are the turtle-mites (perhaps 100-

150,000/square meter) and springtails (50-

100,000/square meter)(Petersen and Luxton 1982). All

of these small creatures sum to a significant biomass.

The length of adult turtle-mites span the range from

only 100-1200 microns, averaging about 300 microns;

however the living biomass of the turtle-mites alone in

a Northwest conifer forest soil is equivalent to 156

entomologists per square mile (Moldenke 1989). Nearly

all of the species of vertebrates inhabiting the Pacific

Northwest derive a significant portion (or all) of their

nutrition from arthropods (yes, even spotted owls do

too; Meslow personal communication). Even though

abundance, diversity and biomass of arthropods are

high, it is quite likely that food limitation is a significant

limiting factor for many vertebrates. Not all invertebrate

species are alike in nutritive value, and few are easy to

capture. During any given season, the majority of the

diets of most vertebrates are probably confined to a

very restricted list of arthropod taxa. As entomologists,

wildlife ecologists and land managers, we know very

little specific information about the diet of

"insectivorous vertebrates." Many of the "sensitive "

and "endangered" taxa are largely insectivorous; to

manage them effectively we need to know what they

consume.

Due to both the exceptional abundance and high

diversity of arthropods, it is probably safe to assume

that there is no ecological process in which numerous

arthropods do not play a significant role. All too

frequently, theoretical "process-centered" ecological

discussions degrade into whether species A

"regulates/drives" species B, or B "regulates" A. In the

real world, both species A and species B are codependent and neither is likely to function normally

without the other. This is especially important for

"ecosystem managers" to comprehend, since the

Andrew Moldenke is Research Associate in Entomology, Department of Entomology, Oregon State University, Corvallis, OR 97331

33
nature of nearly all real-life processes are better

represented as complex food-webs rather than as

simple food-chains.

We could theoretically simplify the forest

entomologist's realm by stating that: a) Douglas-fir

needles fix carbon by photosynthesis; b) these needles

are fed upon by sucking insects (aphids) or chewing

insects (caterpillars); c) these caterpillars and aphids are

fed upon by small insect predators (lady beetles and

ichneumon wasps); d) these predators, in turn, are fed

upon by thrushes; e) which are in turn fed upon by

small hawks, etc., etc. [the classical food-chain model].

But, even when a single herbivore species is preeminently abundant (i.e., the Douglas-fir tussock-moth),

feeding relations are extremely complex. Studies by

Fichter (1984) have shown that: 1) under most nonoutbreak conditions, arthropod predation is sufficient to

keep significant persistent populations of the pest in

check; 2) that control is exerted by a diversity of

predators; and 3) that it is the diversity, itself, which is

most significant, not the individual species of

predators. For instance: first instar caterpillars are

consumed primarily by orb-web spinning spiders which

intercept the wind-dispersed larvae; by day, larger

caterpillars are captured as they feed by sit-and-wait

crab spiders and, as they locomote, by visual-pursuit

jumping spiders; by night, the active hunting

Anyphaena spiders and the concealed pouncing

gnaphosid spiders exact a toll. Even larger caterpillars

disperse to lower branches to feed and are intercepted

by the sheet-web spiders. Any one form of predation is

insufficient to control populations; loss of any hunting

strategy would allow the tussock-moth to escape in

either space or time and for populations to increase

dramatically. It takes a diversity of hunting strategies to

achieve "control".

Entomologists have a limited knowledge of the

functional roles of soil-dwelling arthropods. We lump

them into diverse groups, such as: fungivores, micropredators, macro-predators, shredders, etc. This

categorization, overly simplified by the abysmal lack of

knowledge about their behavior and functional linkages

within the soil ecosystem, leads to the logical

questioning "Well, if so many species all do this same

thing (i.e., fungivory), aren't they 'redundant'? Couldn't

the ecosystem function just as well without large

numbers of these species?"

I can't answer this question because: 1) this

redundancy is largely artifactual, caused by scientists'

lack of knowledge; and 2) the "redundancy" of, say,

fungal-feeding arthropods is closely tied to the

"redundancy" of soil-dwelling fungi. Tiedge (1994)

estimates there may be as many as 40,000 species of

bacteria and fungi in a single teaspoon of forest soil.

The more silviculturists study the symbiotic mycorrhizal

fungi, the more they are convinced that diversity is

necessary for forest health (Trappe 1977; Molina and

others 1982; Amaranthus and Perry 1989). Numerous

investigations have shown the positive feedback

between diverse fungivores and the diversity of the

fungus community (Newell 1984; Visser 1985; Mitchell

and Parkinson 1976). It is greatly upsetting for a

scientist to say that we don't understand the

significance of diversity, but that we "believe" it is

important. But in 1998, that is primarily the case. We

know that an attribute of all natural ecosystems is high

species diversity, and that the attribute of all "efficient"

arthropogenic systems is species-reduction. As

scientists, we hedge and say that in some way high

species richness confers long-term stability on the

ecosystem. But, we don't as yet understand the

complex values of biodiversity.

In the event that loss of a single species from a

complex ecosystem engenders an unexpectedly large

system-wide effect, ecologists hold that such is not the

usual case and they coin the expression "keystone

species" to emphasize the uniqueness of species with

such system-wide importance. There is no way to

determine how many keystone species inhabit a given

environment. There may be many; there may be few.

Only the loss of one (or more) will tell. As Aldo Leopold

(1987) says, "The intelligent tinkerer keeps all the

parts."

THE CAST OF CHARACTERS

Fungivores

Turtle-mites (Oribatida or Cryptostigmata; fig. 1) are

major elements of soil fauna worldwide, but nowhere

more so than in the temperate conifer forests (Petersen

and Luxton 1982). Most species possess an extremely

impenetrable exoskeleton as adults, and like turtles

have forsaken speed and agility as an escape strategy.

There may be as many as 4 to 6 dozen species in a

square meter of forest soil (Moldenke 1989).

Historically, most have been described as generalist

fungivores. Rearing experiments by Mitchell and

Parkinson (1976) revealed that individual cohabiting

species prefer different species of fungi, and more

recent studies on gut enzymes reveal that some species

gain nutrition primarily from the plant matter in their

diets, some from the fungus matter, and some from a

mixture of both (Seipel and de Ruiter-Dijkman 1993).

The composition of the liquid diet of the more primitive

species is not known, nor is the degree to which some

species feed upon other arthropods (Moldenke and

others 1998).

The commonest turtle-mites around the world are in

the genus Oppiella. Studies by Stefaniak and Senizak

(1978) upon their feeding biology are relevant to the

34
Figure 1 —A fungivorous turtle-mite, Belba. Mites have 4

pairs of legs. This species carries its former shed skins (4) on

top of its back.

discussion of species redundancy. Stefaniak and

Senizak found that Oppiella feed upon a number of

fungi in the environment, but that they consistently

refused to eat four species and would choose to starve

to death in their presence rather than consume them.

They would feed upon several other species, not

eagerly, but the fungi could sustain their growth.

Several others would be fed upon eagerly, the mites

could mature, but no viable eggs could be laid (and

length of time required for maturation was prolonged

an additional 25-50%). Only three individual species, of

all the ones tested, were able to sustain both growth

and reproduction. In the soil, rather than a Petri-plate,

turtle-mites probably never encounter enough of a

single species of fungus to feed upon for their life span

(from 1 to 12 months), so they are continually forced

to prune whatever is available. Judging from Stefaniack

and Senizak's experiments, most of what turtle-mites

find is barely capable of maintaining them. Soil is filled

with fungi, but through the evolutionary miracle of

biodiversity, each fungivore finds the various fungus

species differentially appetizing.

Springtails (Collembola; fig. 2) are the most universal of

all the soil arthropods, living even in Antarctica and the

snowfields of Mt. Everest (Hopkin 1997). Their name

derives from the appendage at the rear of the body,

which is held doubled-back under the abdomen during

normal activity. Since springtails have an exoskeleton,

they are able to achieve a heightened blood pressure by

contracting the major longitudinal body musculature

(similar to filling a balloon with air). When threatened

by a predator, they release a forceps-like tail-retainer

under their abdomen. Blood instantly rushes into the

base of the tail, which is now free to expand, and the

tail pushes off the ground, catapulting the springtail

into the air. A springtail can jump 20 to 100 times its

body length.

Kilbertius and Vannier (1981) documented the most

impressive significance of these tiny creatures (1-5

mm) in the oak forests of Europe. Tomocerus

springtails (which also live in the oak forests of the

Pacific Northwest) are keystone species that regulate

Figure 2-A fungivorous springtail, Isotoma. Springtails have

3 pairs of legs and antennae. At the rear of the body is a

forked appendage normally retained under the abdomen (2b).

Blood pressure, kept extremely high during normal foraging,

can flood into the base of the tail and extend the Y-shaped

appendage. When this occurs the tail is slapped onto the

surface of the ground and the springtail is catapulted through

the air and away from potential enemies.

the rate of leaf decomposition. Oak leaves (living or

dead) are filled with a highly reactive chemical (i.e.,

phenol) which prevents most caterpillars from

damaging the living oak leaves. This chemical also

prevents arthropod shredders from crushing the dead

plant cells into units that expose the nutrients for

microbial decay. Tomocerus can overcome this

chemical deterrent by filling its digestive system with

inorganic clay particles, before ingesting the oak

leaves. The cation exchange capacity of the clay

prevents the phenols from polymerizing and impeding

digestion. After passing through the digestive system

of the Tomocerus, the phenol remains detoxified in the

feces. The detoxified feces are then available for

exploitation by all the rest of the diverse soil fauna,

which would otherwise be unable to recycle the

nutrients in the forest litter. This detoxified and

efficient nutrient recycling by Tomocerus permits oak

trees on clay soils to be tall and robust, but oaks on

sandy soil are forced to be shrubby and stunted.

Predators

Pseudoscorpions (Chelonethida; fig, 3) are abundant

predators in all soil habitats (Weygoldt 1969). They

have the general body form of a scorpion, to which

they are not related, but without the tail and its

terminal sting. The pseudoscorpion's sting is in the

thumb of their large pincers. Though generally only

1/10-1/4 of an inch in length, their venom is sufficient

to kill all but the largest insect prey. There are several

dozen species of pseudoscorpions in the Pacific

Northwest, some of which habitually move long

distances by hitching rides on winged insects. Though

diminutive, these efficient predators display complex

behavior associated with mating and rearing their

young.

35
Figure 3-A pseudoscorpion has 5 pair of legs, with the first

pair modified into lobster-shaped capture devices.

Pseudoscorpions usually grow to 0.1-0.4 inches in length.

Predacious mites (Gamasida; fig. 4) are the most

abundant arthropod predators. There are probably

several thousand gamasid mites in each square meter

of Northwest forest soil, less in the basin rangelands.

Gamasid mites (as well as psuedoscorpions) seem to be

particularly sensitive to human-caused disturbances.

Augmentation of gamasid and other predatory mites to

control agricultural pests has been a boom industry in

the past twenty years. Though most effective against

herbivorous mite species, individual predacious mites

are often capable of killing large caterpillars as well.

Most groups of predaceous mites are agile, fast moving

and bright red to reddish-tan in color.

Figure 4-Predaceous gamasid mites are capable of rapid

motion. They kill their prey with poison secreted in small

scissor-shaped appendages at the front of the body.

Ants (Formicidae; fig. 5), though primarily predators,

perform a wide variety of ecosystem functions. Their

composite effect on aeration and water infiltration of

compacted soils is unparalleled. During the period of

maximum brood rearing, their predatory vacuuming of

the environment may extend for 100m from the central

burrow. Many species supplement their diet with

flower nectar and aphid honeydew, which they obtain

by tapping their "aphid-cows" with their antennae. In

the Great Basin, the most conspicuous ants

(Pogonomyrmex) harvest the seeds of plants; the dense

abundance of ants and the rank preference order for

seeds of different species shapes the vegetative

community of annual plants.

Figure 5-The face of an ant. Most ants are efficient

predators.

Shredders

The shredders of the soil ecosystem initiate the process

of nutrient recycling. Shredders come in a variety of

sizes (macroshredders—millipedes; mesoshredders -

sciarid gnats; and microshredders—penknife turtlemites; fig. 6). Most shredders consume fallen leaves,

but the native sowbug (Ligidium gracile) consumes the

final stages of class V log decay (Moldenke personal

observation). By crushing intact plant cells, a millipede

has been documented to increase the surface area

availability for microbial enzyme attack 10,000 fold

(Camatini 1979). The cyanide-producing millipede

(Harpaphe haydeniana) is the most critical shredder on

the west side of the Cascade Mountains (Moldenke

1984); it is 1 % to 2 inches long and black with a

lateral pair of red to yellow stripes. The odor of roasted

almonds that it produces when captured by hand is

hydrogen cyanide gas, perhaps the most efficient

metabolic poison known; as a consequence, it is nearly

immune from predation (except for one species of

beetle, Promecognathus laevissimus; fig. 7). The

presence or absence of shredders in the soil can alter

the total overall rate of nutrient mineralization by tenfold, and the growth of tree seedlings by four-fold

(Anderson and others 1985).

36
Figure 6-The micro-shredder penknife-mite, Phthiracarus.

When threatened by a predator the mite can retract its head

and all 4 pair of legs into its posterior "shell," covering all the

tender structures with the flat plate seen over the head.

Figure 7-The macroshredder millipede, Harpaphe haydeniana

and its specialist-feeding predator, the groundbeetle

Promecognathus laevissimus.

SOIL FOOD WEBS

The Functioning of the Soil Food Web

The basic structure of all natural soil food webs

requires a minimum of four components; (1) the living

plant root, which is incapable of enzyme production,

but obtains nutrients for the plant passively through

water absorption; (2) the soil bacteria and fungi, which

are in most circumstances fuel (carbohydrate)-limited,

but are capable of a bewildering diversity of

degradative enzymatic decomposition processes; (3)

the fauna which graze upon the microbial biomass,

thereby releasing, as part of their waste products,

water-soluble ( and hence "useful") forms of plant

nutrients; and (4) the predator fauna which keep the

population of microbial grazers below levels which

might limit the microbes in their pivotal role as the

"biological sponge" which prevents loss of nutrients

from an ecosystem.

This cycle of growth and decomposition is driven by

photosynthesis. Plants send from 25 to 50% of the

carbon fixed by photosynthesis directly to the roots to

be excreted in order to facilitate microbial growth. This

fuel is shared with both specific symbiotic fungi

(mycorrhizae) and the general microbial population of

the rhizosphere (the volume of soil within 1-2 mm of a

living root tip). Bacteria are important as local foci of

intense metabolic activity on a scale of several

millionths of a pinch of soil. Fungi are capable of slower

metabolic activity, but because their basic body form,

similar to a railroad system, allows the simultaneous

exploitation of widely spaced, yet different, required

nutrients, activity can be maintained continually. Both

bacteria and fungi are relatively incapable of

movement; they require insects and other arthropods to

bring them to sites of new resources or to physically

mix the resources currently available. Arthropods which

graze upon the fungi are seldom able to consume them

entirely, resulting in stimulation of rapid regrowth (like

a cow in a pasture). The result of all of this microbial

growth is the immobilization of nutrients into living

biomass. The long-term health of any ecosystem is

intimately dependent upon minimizing the rate of

nutrient loss; the greater the biomass of soil microbes,

the slower the rate of nutrient loss. The decrease in soil

organic matter (both living and dead) is a major factor

leading to chemical pollution in waterways in the

United States.

The grazing by soil fauna leads to momentary

availability of soluble nutrients for two reasons. (1) The

ratio of carbon to nitrogen in the tissues of bacteria and

fungi is less than that of animals. As a consequence,

an organism feeding on a bacterium or fungus must

excrete any nitrogen in excess of its own requirements

(or become poisoned by it). (2) As an animal locomotes

in search of food, it burns (respires) carbohydrates for

energy. As CO2 leaves the body, other nutrients

(nitrogen, etc.) must leave it as well in order to

maintain the ratio within living cells. Any nutrient

release by the fauna is immediately recaptured by

microbes in their vicinity, unless the release is close

enough to a root to be swept into the plant by the

inward rush of water caused by transpiration from the

leaves. Since most faunal activity is within the

rhizosphere, faunal activity does result in nutrient

uptake (but only when the stomata are open).

Soil is physically and chemically diverse at the scale of

microbes and arthropods. The movement of arthropods

aerates the soil and transports microbial inocula.

Microbial diversity, though fundamental, is difficult to

measure. Arthropod diversity is far easier to monitor

and is likely to reflect microbial diversity. Soil

degradation is associated with a decrease in soil fauna

37
diversity. Some forms of degradation are more easily

measured structurally or physically, for instance the

depth of litter or the resistance to a penetrometer. But

complex chemical changes may best be monitored

through arthropod biodiversity. Any change in

invertebrate diversity should be interpreted not just in

isolation, but as an integral part of the basic

functioning of the soil foodweb, which supports all

plant life.

Management Considerations

A generalized foodweb of the diversity of soil- and

litter-associated invertebrates is represented in figure 8.

There are indistinct boundaries on the definition of soilassociated arthropods; the herbivorous fauna on the

grass, forb and sometimes shrubby vegetation is a

major component of the food resource of epigeic

macropredators. Though many invertebrate predators

(i.e., Calosoma sycophanta) routinely ascend tall trees

to hunt, they represent exceptions and are not

considered in this discussion. Far more important, are

all the species of truly arboreal herbivores, which

descend to the ground to pupate. Such "amphibious"

species not only provide food resources to soil-dwelling

predators, but they represent important additional

linkages between potential canopy health and

prescribed burning protocols.

The actual dynamics of the soil foodweb in any Pacific

Northwest community is far more complex than the

one outlined in figure 8. The foodweb structures among

differing community types (forest, desert, grassland)

are likely to differ significantly, both in terms of the

relative sizes of the boxes and the rates of energy flow

between them. No foodweb analysis of invertebrates

has ever been attempted in the Pacific Northwest on

other than a very generalized basis, figure 8 represents:

(1) a structure into which most communities can be

accommodated; and (2) a structure which is adapted

for invertebrate bio-indicator use.

The biomass of organisms high in the trophic pyramid

is relatively insensitive to shifts in species composition

of the lowest trophic levels and is theoretically

sensitive only to the total primary production of the

ecosystem. However, the species comprising the higher

levels of the trophic pyramid are very sensitive to

changes in management policies. To understand

whether management practices will affect community

structure, it is necessary to differentiate the producer

biomass into the separate components within column I

and to quantify how efficiently this biomass is

transformed into column II. The majority of the biomass

and species richness of any community is contained

within column II. Any (and perhaps, every) major

change in management policy will alter the percentages

of biomass between boxes within column I, and

consequently within column II as well. Such changes

are not necessarily "bad", but it is important to

recognize such a change (and quantify it) when it

occurs. The organisms within column II are the engines

that drive these 20 separate functions within an

ecosystem. In general, increases either in the biomass

of resources or their spatial heterogeneity will favor

increased species richness.

The farther to the right one proceeds in figure 8, the

larger the individual organisms become and the easier

they are to collect and monitor. However, the further to

the right one proceeds, the lower is the species

richness and the more generalized is the behavior (and

therefore the less useful it is at indicating basic

changes in energy functioning within the ecosystem).

There are far more than 10,000 species of arthropods

(though many of them do not even have scientific

names yet) within the Pacific Northwest. Though the

specific identity of the members of any one box in

figure 8 will vary between community types and

between sites, at the family-level of taxonomic

resolution there is considerable consistency. I present

the taxa in table 1, as a checklist for comparing

different sites. This list is certainly not all-inclusive,

but it attempts to encompass the great majority of

arthropod taxa found associated with the soils of the

Pacific Northwest. It is not yet possible to try to

enumerate all the species of arthropods at any one site,

but it is certainly logistically practical to monitor several

of the component groups when most of the species are

collected in a particular trapping system. For instance,

epigeic macropredators, macroherbivores and

macrodetritivores can be collected in pitfall traps and

compared between treatments. The trapped specimens

can be weighed to get an estimate of biomass, and

morphospecies analysis of the component groups can

indicate significant changes in species richness.

These taxa are listed in the table, not as an attempt to

facilitate another one-species-at-a-time "bologna"

approach to conservation/monitoring, but rather as an

educational device. Most ecosystem managers are

unaware of the functional roles played by invertebrates,

and even if they are aware of the roles, they have

never learned who does what. If we are serious about

managing our natural resources for the long-term, then

we have to ensure the persistence of all the functional

links in the community.

CRITICAL ENVIRONMENTAL

DETERMINANTS

I have grouped the critical environmental determinants

into three categories to fit with figure 8: epigeic

habitats, soil surface, and hypogeic habitats.

38
Figure 8-Structure of the invertebrates in a typical soil foodweb. The numbers in the boxes refer to the taxa in table 1. The

arrows represent the pattern of energy flow (food intake) within the community.

39
Table 1 —The usual components of the soil foodweb in nearly all ecosystems. Taxonomic names ending with -idae are family level groupings.

EPIGEIC

1-

2-

3-

4-

5-

6-

phloem suckers 1aAphidae +

xylem suckers

Cicadelloidea 2aFulgoroidea

cominutors 3aAcridoidea

Tettigonioidea

Grylloidea

Noctuoidea +

Mollusca

Sminthuroidea

seed harvestors

Pogonomyrmex

Pheidole

seed "parasitoids" 5aEurytomidae

Curculionidae

Bruchidae

granivores

Lygaeidae

Harpalinae+

SURFACE DWELLING

7-

8-

9-

muscivores

Endeostigmata (pt)

Byrridae

Tingidae (pt)

Boreidae

lichenivores

Polyxenes

Machilida

Oribatida (pt)

Coccinellidae

Nabidae

Strepsiptera (pt)

Trombidiidae

Ichneumonidae

Sphecoidea

Anconema (Mollusca)

Aeliothripidae +

Euryopis

Myrmeliontidae

algivores/cyanobacterivores

Endeostigmata (pt)

Liposcelidae

Nematoda (pt)

10.

11.-

12-

13-

14-

granivores

Lygaeidae

Tenebrionidae

Silvanidae

Ptinidae

Dermestidae

detritivores

Isoptera (pt)

Tenebrionidae

Ptyctimina

Diplopoda

Raphidiphorinae?

Isopoda

Camptosomata (pt

CWD burrowers

a) Isoptera

b) Scolytidae

c) Cerambycidae

Buprestidae

Kylophagidae

Anobiidae

Bostrichidae

d) Camponotus

necrophages

11a-Promecognathus

Zarhipis

Trombidiidae

Lampyridae?

)

1 2a-Stephaniidae

Orussidae

Cleridae

Cucujidae

Ostomidae

Histeridae (pt)

(Including 2o predators on

bacteria/nematodes/etc)

Histeridae

Silphidae

Hydrophilidae (pt)

Staphylinidae (pt)

Dermestidae

Tineidae

Muscoidea

Trogidae

Phoridae

coprophages

Scarabaeidae (pt)

Muscoidea

Scatophagidae

Silphidae (pt)

Leiodidae

Hydrophilidae (pt)

1 5- pollenivores

Collembola

Oribatida (pt)

Anthicidae?

HYPOGEIC

16- root suckers

Cicadidae

Cydnidae

Eriococcidae (pt)

Ortheziidae+

Tylenchida+

17- root chewers 

Grylloyalpidae 

Stenopelmatidae

Curculionidae

Scarabaeidae (pt)

Elateridae

Chrysomelidae

Hepialidae

Pyralidae

Symphyla (pt)

1 8- root detritivores

Curculionidae

Tipulidae

Elateridae

Tenebrionidae

Oligochaeta

19- pollen cachers 

Apoidea 

20- cadaver cachers

Sphecoidea

Pompilidae

40
21

- primary predators

Aranae (pt)

Formicidae

Vespidae

Geocoris

Nabidae

Reduviidae

Carabidae (pt)

Cantharidae

22- secondary predators

Araneae (most)

formicid colonies

vertebrates

23- higher level predators

vertebrates

24- primary predators

Formicidae

Carabidae

Staphylinidae

Sphecidae

Chilopoda

Araneae

Phalangida

Acarina (pt)

Panorpidae

25- secondary predators

Araneae (large spp)

Gnaphosidae

Antrodiaetidae

Agelenidae

Lycosidae

Lithobiomorpha

Scorpionida

Solifugida

vertebrates

Flicker

formicid colonies

26- higher level predators

vertebrates

shrews

lizards

birds

27- primary predators

Formicidae

Geophilimorpha

vertebrates

28- secondary predators

Formicidae

Geophilimorpha

Diplura-Japygidae

29- primary predators

Formicidae

Geophilimorpha

Diplura-Japygidae

30- secondary predators

Formicidae

Geophilimorpha

vertebrates

31- higher level predators

vertebrates

Talpidae

32- microconsumers

Protozoa

Amoebae

Ciliates

33- filter-feeding consumers

Rhabditoidea

Rotifera

34- puncture-feeders

Tylenchida +

Protozoa (pt)

35- comminuting consumers

Oribatida

Endeostigmatida (pt)

Collembola

36- meso-predators

Tardigrada

Mononchida

Gamasida

Endeostigmata (pt)

37- larger meso-pre

Gamasida

Endeostigmata 

Micryphantidae

Scydmaenidae

Pselaphidae

Pseudoscorpion

Geocoris

Rhaphidoptera

Staphylinidae

Aleocharinae +

Throscidae

Palpigrada

Notiophilus/Leis

38- larger comminu

Mycetophilidae

Sciaridae

Cecidomyiidae

Phoridae

Drosophilidae

Protura

Campodeidae?

Thysanura?

Thysanoptera

Staphylinidae (p

Ptiliidae

Scaphidiidae

Cisidae

Derodontidae

Nitidulidae

Endomychididae

39- macroarthropod 

Formicidae

Chilopoda

Vespidae

Pompilidae

Sphecidae

Reduviidae

Thysanoptera (p

Nabidae

Carabidae

Cicindelidae

Staphylinidae

Cantharidae

Mydaeidae

Phalangida 

41
Epigeic Habitats

Grazing —In general, any photosynthetic structures

that end up in the stomach of a vertebrate grazer are

subtracted from the resource base for invertebrates. A

reduction in grazing by mammals will also increase

habitat structure and refugia.

Drought—Precipitation is generally highly

unpredictable. Total annual precipitation coupled with

periodicity determines gross primary productivity in

large part.

Habitat Refugia —Every species requires refugia of

some sort; the more numerous and diverse the

potential refugia, the more individuals and species will

prosper. Litter depth and coarse woody debris are often

cited as examples of refugia.

Soil Surface Phenomena

Surface compaction/cementation—Any change

leading to a decreased ability of the fauna to pass

through the boundary layer or a decreased infiltration of

precipitation is generally detrimental. In many arid

environments a cyanobacterial crust (see Belknap, this

volume) forms on the surface of the soil. The

relationship that this crust has on the total community

is unclear; one would expect it to facilitate faunal

diversity due both to its nutrient biomass and ability to

absorb water.

Litter depth—In forest environments the litter layer is

apparently positively correlated with increased density

and diversity (Madson 1998). Litter serves as both a

food source and in a three-dimensional habitat

capacity. In scrublands a definitive litter layer is

generally absent, but individual detrital pieces are the

center of soil invertebrate activity, especially after they

are anchored in place by fungal hyphae and are no

longer subject to wind displacement. Early in the

morning vertical migration of the microfauna creates an

evanescent hot spot of activity within the litter

(Wallwork 1970).

Erosion —In forest environments erosion is promoted

principally by fire. However, on steep slopes within

semi-arid forests, the passage of mammals (cows,

deer, man) can also seriously disrupt the fragile litter

layer. Under circumstances in which nutrients are

limiting, the continual flux of recyclable litter and

nutrient-rich topsoil downslope is likely to pose a

serious stress to upland vegetation. In the scrublands,

erosion (both wind and surface rainwater) is

accelerated by any disturbance of the surface layer; the

most usual form of disturbance is by mammalian

passage.

Condensation/seasonality— For much of the year

many of these surface soil microenvironments are

severely limited by water; the only reliable source of

water is dew formation. Any physical or chemical

attributes of the soil surface that promote condensation

should be encouraged.

Habitat refugia —In any region, other than the more

mesic forests, a premium must exist on physical means

that promote vertical circadian migration patterns. In

forested systems, consolidated clay soils are frequent,

which limits vertical movement to those few taxa that

have the actual ability to burrow, or to those fauna that

can utilize the burrows created by other species. Very

few soil invertebrates are physiologically capable of

true burrowing; most of them, such as anecic

earthworms and geophilomorph centipedes are absent

from drier soils. Cicadas, millipedes, gryllotalpids and

stenopelmatids may exercise "ecosystem engineer"

keystone roles in these environments. In

unconsolidated desert soils, true burrowing is not

required and many/most taxa may be able to bury

themselves sufficiently to prevent dessication.

Fire, the ultimate consumer of resources—For

dominant "climax" organisms, most effects of fire must

be detrimental. Historically there were probably a

number of beneficial effects upon the subdominant and

early successional taxa, as there clearly are in the

grassland/savanna biomes worldwide. The

multitudinous roles of fire in Pacific Northwest forest

and desert communities have never been examined

with respect to their effects on upon non-pestiferous

invertebrates. Any features of a fire protocol that

increases heterogeneity and promotes islands of

unburned soil surface will minimize the detrimental

effects of fire.

Hypogeic Functions

Water-holding capacity—In most of these

environments anything that favors infiltration and deep

water-holding capacity is paramount. Most of these

variables are long-term soil genesis properties that are

liable to remain the same under different management

scenarios. Though coarse woody debris acts as a

sponge anywhere in the soil profile, it probably has its

greatest hydrating potential below the surface layers.

Stumps, with their vascular bundles oriented to

facilitate upwards water flow, may not be as useful in

retaining water below-ground as an equivalent amount

of coarse woody debris oriented parallel to the surface.

True deep-burrowing —Vertebrate burrows within the

mineral soil are probably an important resource (both

occupied and unoccupied burrows) in all ecosystems as

they are in the well-publicized Florida sand pinelands

(Myers 1985). There must be numerous inquiline and

42
accidental "guests", that require such borrows for

persistence in the community. The percentage of the

fauna dependent upon vertebrate burrows probably

increases with aridity. Classic studies by Ehrlich and

others, have shown the population density of the

Euphydryas butterfly to be largely controlled by the

populations of gophers, whose burrowing activity alters

host-plant phenology, permitting caterpillars to reach a

threshold body weight before aestivating. Communitywide effects of mammal burrowing are also

documented in many temperate communities in which

bumblebees are the major pollinators; bumblebees

require abandoned burrowing-mammal nest chambers in

which to establish their colonies in the spring.

Deep soil compaction (through overgrazing or

mechanized vehicle use)—This is probably the

ultimate threat in any environment, since as far as I

know it is biologically irreversible. It creates an abiotic

environment and potentially permanently immobilizes

resources within it.

limitations we need to work jointly in academia and

management to develop a continuing fiscal support

base that can: (1) generate local reference material in a

systematic way regionwide, (2) acquire and organize

the necessary literature and consultant references, (3)

train district personnel on a rotating basis, and (4)

enlist the public to participate in a comprehensive

monitoring program. Without ultimate hands-on

ownership by the public, long-term stewardship is not

attainable. Such stewardship is our challenge. The

diversity of both the speakers and the audience at this

symposium holds potential for a solution to eventual

stewardship.

I am developing a network of tools for the study of soil

arthropods: standardized methodologies for capture;

user-friendly identification techniques; a system for

archiving physical specimens and electronic data; and a

network of participants region-wide. It's easy; it's

cheap; it's fun; but most importantly its critically

needed. For more information contact the author.

Please join our network.

CONCLUSION

Many of a land manager's concerns are biologic - the

growth of trees, the yield of cattle. However, the

inorganic aspects of soil have always been seen as

critical determinants. But soil ecologists now realize

that the biologic aspects of soils are perhaps even more

important. This realization has been driven largely by

the perceived changes in soil properties of agricultural

soils. Maintenance of the diversity and integrity of a

complex soil foodweb appears to be critical, in many

ways that are currently without adequate scientific

explanation.

The biodiversity within our Northwest soils is

miraculous. I personally have witnessed degradation of

forest and rangeland soils in numerous parts of North

and South America, all instances associated with a

synchronous depletion of soil arthropod diversity as

well. Before we can fully understand the multitudinous

functions carried out by these diverse creatures, we

need to know more about our native fauna. Basic

descriptive natural history precedes analytic

interpretation. One purpose of this symposium is to

stimulate land managers region-wide to become

involved in this basic learning process.

The interpretation of basic descriptive data is enhanced

by the utilization of similar procedures and similar

sample designs region-wide. The development of userfriendly data archiving and interpretive literature is also

critical. Funding for the study of arthropods and their

roles in soil processes will never be more than a few

percent of the funds devoted to owls, salmon, and

other charismatic megafauna. Due to these funding

LITERATURE CITED

Amaranthus, M.P.; Perry, D.A. 1989. Interaction

effects of vegetation type and pacific madrone soil

inocula on survival, growth and mycorrhiza

formation of Douglas-fir. Canad. J. Forest Res.

19:550-556.

Anderson, J.M.; Huish, S.A.; Ineson, P.; and others.

1985. Interactions of invertebrates,

microorganisms and tree roots in nitrogen and

mineral element fluxes in deciduous woodland

soils. P. 377-392 In Fitter, A.H. (ed.). Ecological

Interactions in the soil Blackwell Scientific, Oxford.

Fichter, B.L. 1984. Arboreal arthropod predation on

early instar Douglas-fir tussock moth. PhD. thesis,

Oregon State University, Corvallis, OR.

Hopkin, S.P. 1997. Biology of the springtails. Oxford

Univ. Press, Oxford.

Kilbertius, G.; Vannier, G. 1981. Relations microfloremicrofaune dans le grotte Sainte-Catherine. Rev.

Ecol. Biol. Sol 17:543-557.

Leopold, A. 1987. A Sand County almanac, and

sketches here and there. Oxford Univ. Press, New

York City.

Madson, S.L. 1998. Correlation between structural

heterogeneity and arthropod biodiversity:

implications for management of Pacific Northwest

43
forests. MSci. Thesis, Oregon State University,

Corvallis, OR.

Mitchell, M.J.; Parkinson, D. 1976. Fungal feeding of

oribatid mites in an aspen woodland soil. Ecology

57:302-312.

Moldenke, A.R. 1989. Creatures of the forest soil.

Educational video, Oregon State University Media

Center, Corvallis, OR.

Moldenke, A.R. 1990. One hundred twenty thousand

little legs. Wings 15:11-14.

Moldenke, A.R.; Baumeister, N.; Ingham, E.; Rygiewicz,

P.; Ver Linden, C. 1998. Experiments in global

climate change: the critical roles of soil fauna and

microbes. Educational video, Oregon State

University Media Center, Corvallis, OR.

Moldenke, A.R.; Baumeister, N.; Estrada-Venegas, E.;

Wernz, J. 1994. Linkages between soil

biodiversity and aboveground plant performance.

Trans. 15th World Congress of Soil Science, Int.

Soc. Soil Sci. (July, 1994, Acapulco) Vol. 4A:186-

204.

western Cascade Range, OR V: an annotated list of

the insects and other arthropods. USDA-FS, Pacific

Northwest Forest and Range Experiment Station,

U.S. Gov. Printing Office, Washington, DC.

Petersen, H.; Luxton, M. 1982. A comparative

analysis of soil fauna populations and their role in

decomposition processes. Oikos 39:287-388.

Siepel, H.; de Ruiter-Dijkman, E.M. 1993. Feeding

guilds of oribatid mites based on their

carbohydrase activities. Soil Biol. Biochem.

25:1491-1497.

Southwood, T.R.E. 1978. Ecological methods with

particular reference to the study of insect

populations, 2nd ed., Chapman and Hall, London.

Stefaniak, 0.; Seniczak, S. 1978. The effect of

fungal diet on the development of Oppia nitens

and on the microflora of its alimentary tract.

Pedobiologia 21:202-210.

Tiedje, J.M. 1994. Microbial diversity: a value to

whom? American Society of Microbiology News

60:524-525.

Molina, R.; Hassicotte, H.; Trappe, J.M. 1982.

Specificity phenomena in mycorrhizal symbioses:

community ecological consequences and practical

implications. P. 357-423 In : Allen, M.F. (ed.).

Mycorrhizal Functioning. Chapman and Hall, New

York.

Myers, R.L. 1985. Fire and the dynamic relationship

between Florida sandhills and scrub-pine

vegetation. Bull. Torrey Bot. Club 112:241-252.

Newell, K. 1984. Interaction between two

decomposer basidiomycetes and a collembolan

under Sitka spruce. Soil Biol. Biochem. 16:227-

239.

Parsons, G.L.; Cassis, G.; Moldenke, A.R.; and others.

1991. Invertebrates of the HJ Andrews Forest,

Trappe, J.M. 1977. Selection of fungi for

ectomycorrhizal inoculation in nurseries. Ann. Rev.

Plant Phytopath. 15:203-222.

Visser, S. 1985. Role of soil invertebrates in

determining the composition of soil microbial

communities. P. 297-317 In: A.H. Fitter (ed.)

Ecological Interactions in the Soil, Blackwell

Scientific, Oxford.

Wallwork, J. 1970. The Ecology of Soil Animals.

McGraw Hill, London.

Weygoldt, P. 1969. The Biology of Pseudoscorpions.

Harvard Univ. Press, Cambridge, MA.

Wilson, E.O. 1992. The diversity of life. Belknap

Press, Cambridge, MA.

44
BIOLOGICAL DIVERSITY AND ECOSYSTEM FUNCTIONS OF FOREST SOIL FUNGI:

MANAGEMENT IMPLICATIONS

Randy Molina1, Thomas O'Dell2, Susie Dunham3, and David Pilz4

ABSTRACT

This chapter provides a brief overview of the ecosystem function and biodiversity of forest fungi, which are critical to

the health and sustainability of forest ecosystems. Species diversity of fungi is second only to arthropods but it is their

functional diversity that makes fungi so important in forest ecosystems. In addition to participating in the entire

nutrient cycling process, they also are fundamental parts of the forest food web, and many forest organisms depend on

them in mutualistic and other symbiotic associations. Recent efforts to conserve rare fungi and manage the commercial

harvest of edible forest mushrooms for sustainable production provide examples of integrating fungi into ecosystem

management schemes.

Keywords: mycology, mycorrhizae, biodiversity, survey and manage, conservation biology, population biology, edible

mushrooms, special forest products, nontimber forest products

INTRODUCTION

Ecologically sound management of natural resources

requires a broad understanding of how ecosystems

function. Likewise, conserving biodiversity requires

knowledge not only of the population biology of

individual organisms but also of the interactions and

interdependencies of organisms. Such understanding

provides a means to include the myriad of less visible

organisms along with the megafauna and megaflora in

our management schemes.

This integrated approach is essential for managing the

great diversity of species and functions of soil

microorganisms in forest ecosystems. Total species

richness is greatest in the soil, but most importantly, the

functions of soil microbes power the complex

biochemistry of nutrient cycling that yields soil fertility

and, ultimately, plant and animal productivity.

This paper briefly discusses the biological diversity of

forest soil fungi and emphasizes their functional roles in

forest ecosystems. Two current examples of managing

forest fungi to conserve rare species (survey and

manage species listed in the northwest forest plan,

USDA and USDI 1994) and to sustain their productivity

(commercially harvested edible mushrooms) provide a

basis for integrating fungi into ecosystem

management plans.

INTRINSIC VALUES OF FOREST FUNGI

Many fungi produce beautiful and elegant reproductive

structures known as mushrooms. These fruiting

bodies or sporocarps are collected regularly by

thousands of amateur and professional mycologists

who marvel at their complex forms, learn their

identifying characters, study their ecology, or simply

savor the edible species for their unique and

delectable tastes. Regional mushroom clubs and

societies are scattered across North America and other

continents. They organize collecting trips and

meetings to discuss their findings and share their

mycological experiences. Just as we recognize the

enthusiasm of bird watcher, native plant, and other

nature loving societies, so too must we recognize

society's long-time interest in, and fondness for,

gathering fungi as a valued recreational activity.

Perhaps the most obvious social value of fungi comes

from their use in medicine. Many molds are famous

for their production of antibiotics such as penicillin.

Ancient pharmacopoeia list numerous medicinal uses

of fungi. Yet, we have barely explored the medicinal

1Randy Molina is a research botanist and team leader, USDA Forest Service, Pacific Northwest Research Station, Forestry Sciences

Laboratory, 3200 Jefferson Way, Corvallis OR 97331.

2

Thomas O'Dell is the regional mycologist, USDA Forest Service, Pacific Northwest Research Station, Forestry Sciences Laboratory,

3200 Jefferson Way, Corvallis OR 97331.

3Susie Dunham is a graduate student in the Department of Forest Science, Oregon State University, Corvallis OR 97331.

4David Pilz is a botanist, USDA Forest Service, Pacific Northwest Research Station, Forestry Sciences Laboratory, 3200 Jefferson

Way, Corvallis OR 97331.

45
potential of fungi. Any conservation discussion that

emphasizes efforts to preserve plant species for as yet

unknown medicinal value must include the fungi.

Several species of wild, edible, forest mushrooms are

harvested commercially in Alaska, Oregon, Washington,

Idaho, and California, and a multimillion dollar wild

mushroom industry has developed (Molina and others

1993, Schlosser and Blatner 1995). This new industry

has created conflicts between traditional recreational

users of forest mushrooms and commercial mushroom

harvesters which has prompted resource managers to

develop new permit systems to protect and regulate this

resource. The monetary value of this new special forest

product and the regional conflict over its proper use has

brought attention to this previously unheralded group of

forest organisms. Such social and economic values

underscore the need to understand the biodiversity of

fungi in relation to both forest and human communities.

The management of special forest product fungi is

discussed in greater detail later.

ECOSYSTEM FUNCTIONS OF FUNGI

All fungi are heterotrophs and they obtain energy,

directly or indirectly, from organisms that can

photosynthesize. Fungi that decompose dead organisms

are called saprotrophs, those that get energy directly

from a living host are called biotrophs. Saprotrophs

include the various cultivated mushrooms, yeasts and

many other species. Biotrophs include those fungi that

form symbioses with living plants (such as mycorrhizal

or lichenized fungi) and many pathogens. Some

pathogenic fungi span the gradient from biotroph to

saprotroph, colonizing a living host and persisting to

decompose its remains.

Miller (1995) organizes the primary functions of fungi in

ecosystems into three general categories, each with

several subcategories:

Physiological and metabolic

Decomposition of organic matter; volatilization of

C, H, and 0; fragmentation

Elemental release and mineralization of N, P, K, S,

and other ions

Elemental storage: immobilization of elements

Accumulation of toxic materials

Synthesis of humic materials

Ecological

Facilitation of energy exchange between

aboveground and belowground systems

Promotion and alteration of niche development

Regulation of successional trajectory and velocity

Mediative and integrative

Facilitation of transport of essential elements and

water from soil to plant roots

Facilitation of plant-to-plant movement of essential

elements and carbohydrates

Regulation of water and ion movement through

plants

Regulation of photosynthetic rate of primary

producers

Regulation of C-allocation belowground

Increased survivability of seedlings

Protection from root pathogens

Modification of soil permeability and promotion of

aggregation

Modification of soil ion exchange and water

holding capacities

Detoxification of soil (degradation, volatilization, or

sequestration)

Participation in saprotrophic food chains

Instigation of parasitic and mutualistic symbioses

Production of environmental biochemicals

(antibiotics, enzymes, and immunosup

presants)

The first category encompasses the physiological and

metabolic attributes of fungi. Fungi are most well

known for their decomposition of organic matter. In

forest ecosystems they play a significant role in the

depolymerization of cellulose, hemicellulose, and

lignin, the primary constituents of wood (Cooke and

Whipps 1993, Miller 1995). A primary outcome from

the complex biochemical degradation processes is the

release of elements and mineralization particularly of

N, P, K, S and other ions. Of course, these become

nutrients available for uptake by plants. Another

critical function in this category is the storage of

nutrients within the vast body of fungi in the soil.

Fungi are well known as nutrient accumulators, and

some of their mycelial structures, such as

rhizomorphs, tenaciously maintain elements in organic

form, thus preventing their loss from the ecosystem.

Fungi also participate in many ecological processes

that reverberate through the ecosystem (Trappe and

Luoma 1992). Related to the elemental storage

capacity noted above is the direct movement of

energy in the form of photosynthetic carbon (sugars)

between aboveground and belowground ecosystem

components. For example, mycorrhizal fungi obtain

their primary carbon from photosynthate directed

towards the fine roots; as much a 50 percent of the

carbon fixed in photosynthesis can move to

mycorrhizal fungi (Smith and Read 1997). This

significant source of carbon fuels many of the

rhizosphere organisms and their processes in the soil

(Molina and Amaranthus 1991). As noted by

Moldenke (this volume), for example, this nutrient rich

fungal mycelium is a food source for the majority of

soil microarthropods.

46
At larger spatial scales, fungi also contribute to

development of critical niches for myriad forest

organisms (Miller 1995). For example, although many

soil fungal pathogens directly kill or weaken forest

plants, the dead snags or the softened tissue of infected

trees allow for cavity construction by birds and other

wildlife (Parks and Shaw 1997, Schowalter and others

1997). At a forest gap or stand scale, root pathogens

such as Phellinus weirii (Murr.) Gilbertson or Armillaria

species create openings in the forest that allow for

successional development of understory, thus enhancing

biodiversity and providing resources such as browse for

large mammals (see Thies, this volume). Although it is

important that we not exacerbate diseases with our

management practices, it is important to realize that

pathogens are a natural component of forest

ecosystems and contribute to landscape and biological

diversity. Schowalter and others (1997) provide a

framework for integrating the ecological roles of

pathogens and mycorrhizae in managed forests.

Miller (1995) defines the third functional category as

mediative and integrative and emphasizes numerous

ways other organisms are affected by fungal presence

and processes. These range from facilitating the

movement of water and nutrients from soil to roots to

providing protection against some root pathogens and

increasing seedling survival. From a holistic ecosystem

perspective, fungi can mediate restoration of severely

disturbed sites and thus act as key resiliency factors

(Amaranthus, this volume, Molina and Amaranthus

1991, Perry and others 1989a). Nowhere is this

integrative function exemplified better than with the

mycorrhizal symbioses.

MYCORRHIZAL FUNGI OF FOREST

PLANTS

Mycorrhiza literally translates as "fungus-root" and

defines the common association of specialized soil fungi

with the fine roots of nearly all forest plants.

Mycorrhizal associations represent one of the more

widespread forms of mutualistic symbioses in terrestrial

ecosystems. Indeed, these plant-fungus associations

have coevolved over the eons such that each partner

depends on the other for survival.

The mycorrhizal fungus basically serves as an extension

of the plant root system, exploring soil far beyond the

reach of the roots and transporting water and nutrients

to the roots. The uptake of phosphorus and nitrogen are

especially critical functions of mycorrhizal fungi, which

can release bound forms of these nutrients otherwise

unavailable to the roots. In return, the plant is the

primary energy source for the fungus, providing simple

sugars and vitamins produced in photosynthesis and

transporting them to the roots and then the fungus.

Mycorrhizal fungi have limited saprophytic ability; that

is, they possess fewer enzymes needed to efficiently

break down complex organic matter in the soil than do

true saprotrophs . Hence, their survival in natural

ecosystems depends on the supply of sugar

(photosynthate) from their associated host plants.

Likewise, many plants strongly depend on their

mycorrhizal fungi for nutrient and water uptake. This

coevolution in dependency reinforces the need to view

organisms through their relations within the forest.

There are several classes of mycorrhizae, but only

ectomycorrhiza will be discussed (see Smith and Read

1997 for a complete review on mycorrhizae).

Ectomycorrhiza is the most common type on forest

trees in the Western United States and Canada.

Ectomycorrhizal hosts in our region include all species

in the Pinaceae (Abies, Larix, Picea, Pinus,

Pseudotsuga, and Tsuga), Fagaceae (Castinopsis,

Lithocarpus, Quercus), Salicaceae (Populus, Salix),

and Betulaceae (Alnus, Betula, Cory/us). Additionally,

genera such as Arbutus, Arctostaphylos, Cercocarpus,

and Eucalyptus form ectomycorrhizae. The dominant

nature of these species in western forest ecosystems,

especially Pinaceae, contributes to the pervasiveness

of ectomycorrhizae in western forests.

Ectomycorrhizae develop on the short, feeder roots.

The fungus forms a sheath or mantle of fungal tissue

around the feeder root. The mantle serves as a

storage tissue for nutrients received from mycelium in

the soil and physically protects the fine roots from

some pathogens and desiccation. The fungus also

penetrates between the root's cortical cells to form a

network of fungus tissue called the Hartig net.

Nutrient exchange occurs within this extensive,

intimate contact zone: sugars and vitamins move from

plant to fungus, and water and nutrients move from

fungus to plant. Many ectomycorrhizal fungi produce

plant hormones that cause extensive branching of

colonized feeder roots. This branching greatly

increases the root surface area and provides an

extensive contact zone between the fungus and root.

The fungus grows from the colonized roots into the

surrounding soil. Mycelial colonization of the soil

differs among ectomycorrhizal fungi; some may grow

only a few centimeters into the soil and others can

grow several meters from the ectomycorrhiza. Some

fungi produce dense, hyphal mats that strongly bind

the soil and organic matter. If the mycelium is white

or brightly colored, these extensive mats may be

readily visible when a bit of the upper organic layer is

removed. Other fungi produce colorless or dark

mycelia that are more difficult to see with the unaided

eye, but their growth into the soil is likewise

extensive.

Thousands of ectomycorrhizal fungus species occur in

our region; Douglas-fir (Pseudotsuga menziesii (Mirb.)

47
Franco) alone associates with nearly 2,000 species

(Trappe 1977). Most ectomycorrhizal fungi are

basidiomycetes and ascomycetes, and many produce

mushroom fruiting bodies that abound in western forests

during the moist times of the year. Many

ectomycorrhizal mushroom-forming fungi are especially

prized for their edibility. Well-known edible species

occur in the genera Cantharellus (chanterelles), Boletus

(boletes), Lactarius (milky caps), Tricholoma

(matsutake), Hydnum (tooth fungi), and Ramaria (coral

fungi) (see Molina and others 1993 for a complete listing

of ectomycorrhizal fungus genera). Another large and

diverse group of less well-known ectomycorrhizal fungi

are the truffles, which produce fruiting bodies beneath

the duff or soil surface. Only a couple of truffle species

are harvested commercially, but as will be discussed

later, truffle fungi are a central part of the forest food

web.

The habitat requirements of ectomycorrhizal fungi and

their interactions with particular plant species are poorly

understood. Some fungi are more abundant in certain

age classes of forests (O'Dell and others 1992b). As

plant-species composition changes during forest

succession, the fungus communities similarly undergo

change. This fungal succession is in response to

changes in tree composition, tree age, and soil qualities

such as accumulation of organic matter. The ecological

requirements of mycorrhizal fungi, particularly their

relation to forest community succession and disturbance

events, represent large knowledge gaps needing

research attention.

Research on the biodiversity of ectomycorrhizal fungi in

the Pacific Northwest has focused on the relations and

interdependencies among mycorrhizal plants and fungi

rather than on total number of fungal species. It is

within this context of relations that the functional

characteristics of ectomycorrhizal fungi emerge. Many

ectomycorrhizal fungi, for example, have broad host

ranges; others are host specific, typically restricted to a

single genus of host (Molina and others 1992). In the

Pacific Northwest, Douglas-fir forms ectomycorrhizae

with numerous host-specific fungi; Pinus species

similarly associate with a wide array of Pinus-specific

fungi. At the same time, both Douglas-fir and Pinus

species are compatible with many broad-host-ranging

fungi. Thus, a forest composed of different

ectomycorrhizal host species will exhibit a spectrum of

ectomycorrhizal relations with diverse fungi. Molina and

others (1992) hypothesize that host-specific fungi may

provide different benefits to their specific hosts than do

more generalist, broad-host-ranging fungi. Perry and

others (1989b) found that host-specific fungi can reduce

competition among different tree species compared to

broad-host-ranging fungi.

The ability of different host species to form

ectomycorrhizae with mutually compatible fungi also

influences forest succession and community

development (Molina and others 1992). Read and

others (1985) demonstrated that forest trees can be

connected by common mycorrhizal fungi and that

photosynthate, minerals, and water pass among plants

via the fungal linkage. Understory plants, especially

seedlings, may benefit from this linkage if

photosynthate passes from the overstory tree to a

shaded understory plant (Simard and others 1997).

Or, a plant entering in the later stages of forest

succession can grow into a community of

ectomycorrhizal fungi that has been nourished and

maintained by early successional plant species. Perry

and others (1987) hypothesize that these functional

connections in time between successional stages as

mediated by compatible mycorrhizal fungi contribute

significantly to forest resiliency and recovery following

disturbance. Thus, as an ecological management tool,

resource managers may chose to maintain a

compatible stocking level of commercially nonvaluable

pioneering and understory plants following tree

harvest. In this approach, the beneficial mycorrhizal

fungi are maintained on the site by pioneering

vegetation for use by later tree crops (see

Amaranthus, this volume).

Ectomycorrhizal fungi are most well known for

enhancing nutrient uptake and growth of the host.

Consequently, many mycorrhiza researchers have

developed methods to inoculate forest tree seedlings

with selected, beneficial mycorrhizal fungi (Castellano

and Molina 1989, O'Dell and others 1992a, Perry and

others 1987). Ectomycorrhizal fungi also perform

other critical ecosystem functions, however, and it is

from understanding this diversity of functions that

additional ecological management tools can be

developed. Management practices that encourage the

healthy functioning and recovery of forest fungi and

other soil microbes following disturbance will provide

positive benefits to ecosystem productivity

(Amaranthus and others 1990, Molina and

Amaranthus 1991).

TREES, TRUFFLES, AND BEASTS: AN

EXAMPLE LINKING ABOVEGROUND

AND BELOWGROUND ECOLOGIES

Ecological theory recognizes that organisms do not

live in isolation from other organisms. The literature is

rich with examples of interdependencies among

plants, among animals, and between plants and

animals. But rarely do these examples include the

"invisible" flora, fauna or mycota in ecosystems. The

following example of ecological connections among

mycorrhizal fungi, old-growth forests, and a

controversial, endangered bird, the northern spotted

owl (Strix occidentalis), in the Pacific Northwest

48
illustrates the value and necessity of integrating

aboveground and belowground ecologies.

Northern spotted owls live predominantly in old-growth

forests. A primary prey of the northern spotted owl is

the northern flying squirrel (Glaucomys sabrinus Shaw).

During certain times of the year, northern flying squirrels

feed on a wide variety of subterranean mycorrhizal

fungus sporocarps called truffles (Maser and others

1985). As truffles mature, the interior tissue fills with

spores and emits a distinctive odor. Small mammals,

such as the northern flying squirrel, forage for these

truffles by smell and then excavate and devour them.

The spores pass through the mammals unharmed and

are dispersed during defecation. The spores then

re-enter the soil where they germinate and form new

mycorrhizae with tree roots. This dispersal is important

for young tree growth in areas where the previous forest

has been destroyed by natural catastrophic events (fire,

wind) or timber harvests. Hundreds of truffle fungi

occur in western forests and most forest mammals

consume them. Thus, in this example of

tree-fungus-prey-predator connections, resource

managers can view the northern spotted owl within the

context of its ecosystem linkages rather than through a

simple dependency on old-growth forests. Such holistic

perspectives enable development of ecosystem

management tools and avoid the often divisive nature of

single-species conservation approaches.

BIODIVERSITY OF SOIL FUNGI

Soil harbors tremendous fungal diversity; a single gram

may contains thousands of species. It is difficult to

describe, let alone understand, this diversity. Most soil

fungi produce only microscopic fruiting structures and

require sterile culture methods for their isolation and

identification. Macrofungi (species that produce

macroscopic fruiting bodies), although having their

cryptic aspects, can be observed without having to be

cultured. Furthermore, a century of observation

provides information regarding their rarity, distributions,

and habitat preferences. Macrofungi include important

commercially harvested species and the majority of all

species covered under the survey and manage provisions

of the Northwest Forest Plan (USDA and USDI (1994).

Our discussion will focus on macrofungi, but many of

the concepts apply equally to all fungi.

The concept of diversity encompasses many species,

and relative abundancies, and genetic variation among

individuals of a species. Maintenance of ecosystem

function and protection of biological diversity require

conservation of species and within-species variation.

Diverse assemblages of species contribute to ecosystem

resiliency (ability to recover after disturbance) (Tilman

and Downing 1994), and of course different species

may serve different functions in the ecosystem. The

variation among individuals of a species is grist for the

mill of natural selection and this genetic variation

increases the chance for a species to survive in the

face of a changing environment.

Species diversity and the intraspecific genetic diversity

among individuals of a species both contribute to

functional diversity in the ecosystem. Although

different species may do drastically different things

(for example, decompose wood vs. form mycorrhizae),

individuals of a species differ more subtly, (for

example, decomposing the same substrate most

efficiently at different temperatures).

Two concepts, communities and guilds, are used to

discuss species with similar characteristics. A

community is a group of organisms sharing a piece of

ground; for example, the fungi in a particular

Douglas-fir stand. Communities of species commonly

encountered together are often given a name: the

Pseudotsuga menziesii-Tsuga heterophyllalPolystichum

munitum plant association; fungi are too poorly known

to have many named communities. A guild is a group

of species that share a common function, such as

fungi that form ectomycorrhizae. Increased

understanding of communities and guilds will facilitate

better management of fungi. If a rare fungus often

occurs in a community and that community has an

identified habitat requirement, we can move away

from single-species management toward managing for

the habitat.

Communities are not restricted to a particular spatial

scale; a community could be identified that occurs on

large-diameter Douglas-fir logs. Species in such a

community are at risk from management strategies

that fail to provide a long-term source of such logs.

Succession describes how communities change over

time. Successional change often results from

feedback between organisms in the community and

their environment. Immediately following a

disturbance, for example, nutrient availability

increases, and pioneer species tend to tolerate and

make full use of high levels of soil N and P. As the

nutrients become immobilized in biomass, late-seral

species capable of using organic forms of nutrient and

competing effectively for scarce resources tend to

replace the pioneers. Thus, the different species that

comprise early and late serai communities contribute

to long-term ecosystem productivity.

49
POPULATION BIOLOGY OF FUNGI

What Can Genetic Studies of Fungal Populations

Contribute to Management of Forest Fungi?

Genetic diversity is recognized by the World

Conservation Union as one of three levels of biodiversity

important to conserve (McNeely and others 1990), a

task ultimately falling to land managers. A growing

literature on the possibilities of linking genetics and

ecology in conservation biology (Avise 1994, Avise and

Hamrick 1995, Hastings and Harrison 1994, Milligan

and others 1 994) suggests that spatial analysis of

genetic patterns can provide guidance for defining the

landscape scale at which ecosystem management

activities should occur to best conserve genetic variation

within species (Moritz 1994). Two long-term goals of

our population genetic research on ectomycorrhizal fungi

are to gain some understanding of how past and current

land management practices in the Pacific Northwest

have altered the dynamics of fungal populations, and to

use this information to predict how future land

management scenarios may affect persistence of fungal

populations. Achieving these goals requires an

understanding of the spatial and temporal scales at

which evolutionary significant processes (such as

natural selection, gene flow, and genetic drift) act on

fungal populations. We also need to understand which

processes are most salient at spatial scales relevant to

management, and how variability in landscape structure

affects these processes (Wiens 1989, 1997).

Difficulties With Studying Fungal Populations

Several ecological characteristics of fungi impede our

ability to select scales appropriate for quantifying the

evolutionary processes important to fungal populations.

Selecting an appropriate scale for population genetic

studies requires an understanding of how individuals are

distributed across the landscape. This presents a

difficult obstacle with fungi because the mass of

threadlike mycelium (thallus) that gives rise to

fruit-bodies grows throughout the soil and cannot easily

be measured or counted. Attempts to characterize the

size and distribution of discrete thalli have identified

individuals encompassing hectares (Smith and others

1992), square meters (Baar and others 1994), and

square centimeters (Gryta and others 1997). Dahlberg

and Stenlid (1990) also show that ecological factors

such as forest age can influence the size of individuals

and alter the mechanism by which new thalli are

established. The range of individual size and distribution

unique to each fungal species should dictate the scale at

which population boundaries are encountered and set

the minimum size limit for study area selection.

When fungi are sampled in the field, it is impossible to

tell what proportion of the sporocarps collected

represent genetically identical individuals produced by

the same mycelium, and genetically distinct mycelium

masses may produce intermingled sporocarps (Molina

and others 1993). Discrete thalli can be established

by either sexual reproduction (genets) or clonal

vegetative propagation (Todd and Rayner 1980), and

our understanding of populations hinges on correctly

estimating the contribution of each reproductive type

to the genetic diversity observed within populations

(Avise 1 994)

Determining the boundaries of fungal populations and

identifying landscape features that impede dispersal

requires the development of sampling techniques that

allow us to estimate how much gene flow occurs in

forest stands, in watersheds, and across regions.

These sampling techniques also must account for

variability in vegetation cover, disturbance type, and

disturbance intensity at each scale. The ultimate goal

of such a multiscale sampling regime is to ensure that

the scale of our measurements and the response of

fungal populations to some variable of interest (for

example, disturbance) fall within the same spatial

domain or area affected by the same ecological

processes. Variability across fungal species in

individual size and distribution, mating system, and

dispersal capability makes it impossible to apply a

single sampling protocol to all species. Advancement

in understanding and maintaining genetic variability in

fungal populations requires basic research on how

species with vastly different life histories partition

their variation at different spatial scales. This

fundamental research will lay a general foundation for

predicting how land management strategies may

affect rare species depending on their life histories.

Multiscale genetic studies also will help define scales

potentially appropriate for future research on

demographics or habitat requirements.

Current Research Example: Distribution of

Genetic Variation Across Different Spatial Scales

in Cantharellus

Building a strong knowledge base requires that basic

genetic research be done on model species that will

yield large sample sizes and are predictable in their

fruiting from year to year. Research on such common

species will allow initial studies to have the statistical

power sufficient to correctly identify evolutionary

important pattern-process relations (Dizon and others

1995). Pacific golden chanterelles (Cantharellus

formosus Corner) is an economically important species

that forms mycorrhizal associations with a wide range

of host species over a broad geographic range (Molina

and others 1993). Chanterelles also fruit predictably

and abundantly and provide an ideal model species for

initiating base line genetic studies.

50
Current genetic research on chanterelles is being

conducted in collaboration with an integrated research

project, the Young Stand Thinning and Diversity Study,

sponsored by the Willamette National Forest, Oregon

State University, and the Cascade Center for Ecosystem

Management. It is located in the central Oregon

Cascade Range east of Eugene, Oregon. The overall

goal of the study is to determine how stand-thinning

techniques and leave-tree densities affect site conditions

and ecological responses. Our initial mycological goal

was to determine how chanterelle fruiting respond to

stand thinning (Pilz and others 1996). Our genetic

population study attempts to determine the size and

distribution of chanterelle genets in forest stands that

differ in intensity of disturbances associated with tree

harvest, and to quantify patterns of genetic variability at

fine spatial scales. Results of this research will aid in

the design of a sampling method that will detect genetic

variation at spatial scales relevant to management.

Future research efforts will examine how heterogeneity

in landscape features (for example, forest stand age or

patterns created by different disturbance regimes)

affects rates of gene flow among chanterelle patches.

Using this line of research to develop a better

understanding of how fungi move across landscapes will

facilitate identification of population boundaries and

geographic areas important to the conservation of

genetic diversity.

SURVEY AND MANAGE FUNGI IN THE

NORTHWEST FOREST PLAN: NEW

CHALLENGES IN CONSERVATION

BIOLOGY

A challenge within the Forest Ecosystem Management

Team (FEMAT) (1993) process was protecting biological

diversity for all species within the master plan for

protecting the northern spotted owl. Protecting all

species and their ecological functions is indeed a primary

tenet of the new paradigm of ecosystem management.

Taxa specialist panels were convened to examine

mammals, amphibians, fungi, lichens, bryophytes,

vascular plants, mollusks, and arthropods with the task

of listing rare taxa within the range of the northern

spotted owl and taxa that required late successional

forest habitat. These taxa were then run through risk

assessment screens in conjunction with the various

options being considered for implementation. Under

option 9, the risk assessment yielded 409 species and 4

arthropod communities whose viability was not

protected. The survey and manage guidelines were then

developed as part of the record of decision (ROD)

(USDA and USDI 1994) and serve as the basic structure

for ensuring the viability of these rare taxa.

The survey and manage guidelines are organized around

four strategies, each having different objectives for

protecting species viability and gaining new

information for use in adaptive management. The

taxa were placed into these various strategies in the

C-3 list (USDA and USDI 1994: C49-C61). The four

strategies and their relevance to fungi are described

below.

Strategy 1—Manage Known Sites

Taxa in this category receive the highest priority for

protection. The strategy calls for gathering

information on all extant sites, developing a

Geographical Information system (GIS) data base for

all species and using this information for preparing

management recommendations. Management

activities implemented in 1995 and later must include

provisions for these known sites. In most cases, this

will involve protection of relatively small sites. "For

rare and endemic fungus species, areas of 160 acres

should be temporarily withdrawn from

ground-disturbing activities around known sites until

those sites can be thoroughly surveyed and

site-specific measures prescribed. For one fungus

species, Oxyporus nobilissimus, there are only six

known sites and two of these do not currently have

protected status. Management areas of all usable

habitat up to 600 acres are to be established around

these two sites for the protection of those individuals

until the sites can be thoroughly surveyed and

site-specific measures prescribed. The actions to

protect Oxyporus must be undertaken immediately"

(USDA and USDI 1994: C-5). There are 147 fungal

species listed under strategy 1.

Strategy 2 —Survey Prior to Ground-Disturbing

Activities

This strategy was designed to address the high-risk

species for which survey protocols exist or could be

developed reasonably soon, and for which adequate

searches could be conducted in a single year. Efforts

to design protocols and implement surveys would be

started immediately. The information gained from the

surveys would be used to develop management

strategies for species protection. For fungi, these

surveys must be completed prior to ground disturbing

activities that will be implemented in 1999 or later.

Only two fungal species are listed under strategy 2,

one being Oxyporus nobilissimus.

Strategy 3 —Extensive Surveys

The objective of this strategy is to conduct surveys to

find high-priority sites for species management. "It is

recommended primarily for species whose

characteristics make site and time-specific surveys

51
difficult. For example, some fungi only produce fruiting

bodies under specific climatic conditions, so finding their

location may take several to many years" (USDA and

USDI 1994: C5-C6). Surveys under this strategy had to

be underway by 1996. These surveys are likely to take

several years to complete and to identify high-priority

sites for protection. All strategy 1 fungal taxa also are

listed under strategy 3; 87 additional fungal species are

designated as strategy 3 species for a total of 234

species in this strategy.

Strategy 4—General Regional Surveys

The objective is to survey for the species to acquire

additional information to determine necessary levels of

protection. Although these species are not considered

rare, there is only limited information regarding their

abundance and distribution. The information gathered

from these surveys would be used in adaptive

management to refine the guidelines to better protect

these species. These surveys were expected to be both

extensive and expensive. They had to begin in 1996

and be completed within 10 years. There are 20 fungal

species listed in this strategy.

Current Work in Progress

A primary task in implementing the survey and manage

guidelines was to gather all available information on

extant collections of strategy 1 taxa. For fungi,

collection records and ecological information were

obtained from several mycological herbaria, literature

searches, and field botanists throughout the Pacific

Northwest; about 1800 collections were discovered and

reverified. These are now in a database and available to

field managers. The database is managed and updated

at the Regional Ecosystem Office in Portland, Oregon.

From the known site database and literature searches,

management recommendations were drafted for use by

field personnel. The management recommendations

include information on species descriptions,

nomenclature, known sites and geographic range,

ecology, reproductive biology, habitat requirements,

impacts of past disturbances, specific recommendations

to maintain species presence at known sites, and

research and monitoring needs. These documents will

undergo formal scientific and managerial reviews before

publication.

Future Directions and Challenges

Designing efficient survey protocols and conducting the

extensive surveys called for under strategies 3 and 4

remain the greatest challenges in implementing the

survey and manage guidelines. As part of a national

effort to develop inventory and monitoring protocols for

all fungal groups, members of Pacific Northwest

Research Station Forest Mycology Team at Corvallis

have written protocols for mushroom and truffle fungi.

These protocols will be published in an upcoming book

on international mycological survey protocols by the

Smithsonian Institution and will serve as the

foundation for developing specific regional protocols

for survey and manage fungi.

A high priority is determining which survey and

manage fungal species are at highest risk and thus

should be surveyed immediately. Another priority is to

build simple habitat-based models that will guide

decisions on where to conduct surveys. For most

known site records, little site-specific habitat

information is recorded. Revisiting known sites to

determine whether the taxon is still present (many of

the known sites have been disturbed since the original

collections) and gathering specific habitat information

are urgent tasks prerequisite to building habitat

models.

Complimentary mycological research is needed in two

new areas for building a comprehensive conservation

strategy for fungi. First, because the overall goal is

protecting species viability, we must develop an

understanding of population dynamics for forest fungi.

This includes information on geographic distribution,

genetic structure, dispersal capabilities, reproductive

biology, and habitat fragmentation effects on

population structure. As noted in the previous

section, population biology is poorly understood for

fungi in general, largely because of the difficulty in

defining an individual and how individuals interact

genetically to yield distinct populations. Many

traditional approaches for defining and measuring

populations used for plants and animals simply do not

work for fungi. This research will require use of both

field biological sampling and molecular genetic

approaches (see previous section, "Population Biology

of Fungi").

A second important research need is habitat modeling

at the landscape and regional scale. Given the large

number of survey and manage fungal species, it will

be difficult to assess viability for all species across the

range of the northern spotted owl. Also, because

known sites occur on various public and private land

holdings, it will be necessary to integrate how all

available habitat will contribute to maintaining species

viability. To do so will require development of

predictive GIS and statistical models that examine the

current and future forest conditions affecting fungal

species persistence. The end product would be a

working model that resource managers could use to

examine how their management practices will affect

viability of survey and manage species. This effort

will take considerable cooperation among mycologists,

statisticians, and land managers to develop a usable

habitat model for decision making.

52
Two conditions are necessary for the long-term success

of survey and manage strategies. First, communication

and coordination with field units are needed to develop

management recommendations and survey protocols

that are attainable and efficient. Field managers are key

members in the adaptive management process. They

must be involved in gathering new information from the

surveys and in deciding how best to use the information

to refine management strategies. Second, monitoring

the results of our recommendations and management

decisions is essential to determining species viability.

These monitoring efforts must be carefully coordinated

with other Northwest Forest Plan effectiveness

monitoring activities across the region.

COMMERCIALLY HARVESTED EDIBLE

FOREST MUSHROOMS: OPPORTUNITIES

FOR ECOSYSTEM MANAGEMENT

Edible sporocarps of many ectomycorrhizal fungus

species are highly prized by different cultures and local

markets for their commercial sale are common around

the world. However, the decline of favorite edible

species in Europe (Arnolds 1991, 1995) and Japan

(Hosford and others 1997) has created a market demand

for wild fungi from regions where the species remain

abundant and a modern infrastructure (forest roads and

nearby airports) provides quick access to and export of

the fresh product. These economic forces have led to

development of a multimillion dollar industry of wild

mushroom harvest from the extensive forests of western

North America (Amaranthus and Pilz 1996, Molina and

others 1993, Schlosser and Blatner 1995).

Most of the valuable commercial species are mycorrhizal

fungi. The primary tree hosts include the Pinaceae but

also Fagaceae and Betulaceae. Although we have a

general understanding on geographic range, host

associations, and mycorrhizal characteristics of

important edibles such as Cantharellus (Danell 1994),

we poorly understand their ecology and productivity in

natural forests, an understanding essential to conserving

and managing the use of these fungi. The following

sections discuss ecological and management

implications of this commercial harvest on forest fungi

and focus on current research and monitoring

approaches aimed at developing management guidelines

for protecting the mushroom resource and ensuring a

sustainable harvest.

Commercial Mushroom Harvest in the Pacific

Northwest

In the early 1980s in Oregon and Washington, a small

entrepreneurial wild mushroom industry began to harvest

small volumes of mushrooms and ship them to gourmet

restaurants across the United States. Then, with the

decline of chanterelle productivity in Europe, a new

export market developed: harvested chanterelles were

packed in brine and shipped to canneries in Germany.

This new market vastly increased the volume of

mushrooms harvested and led to development of a

system of harvesters, buyers, processors, and brokers

for the expanding industry. Later in the decade, the

industry returned primarily to the fresh export market.

Schlosser and Blatner (1995) extensively surveyed the

wild mushroom industry in Oregon, Washington, and

Idaho for 1992. They report that the industry

employed about 11,000 people and contributed $41.1

million to the regional economy; nearly 1,800,000

kilograms of wild mushrooms were harvested.

Twenty five species are harvested commercially but

the most economically important are chanterelles

(mostly Cantharellus formosus but also C. subalbidus

Smith & Morse), pine mushroom or American

matsutake (Tricholoma magnivelare (Peck) Redhead),

king bolete (Boletus edulis BulkFr.) and several morel

(Morchella) species. Twenty-eight percent of all

exports went to Japan and 25 percent to European

markets (primarily France and Germany).

The most abundantly harvested ectomycorrhizal fungi

are the chanterelles and American matsutake, with

454,000 and 375,000 kilograms respectively,

harvested in 1992. Nearly half of all harvested

chanterelles were air-shipped to Europe; chanterelles

are popular throughout Europe, but productivity has

declined in some regions (Arnolds 1991). For similar

reasons, 70 to 90 percent of harvested American

matsutake went to Japan, where they supplement the

highly revered Japanese matsutake (Tricholoma

matsutake (S. Ito et Imae) Sing.). Over the last 50,

years productivity of T. matsutake in Japan has

declined dramatically to 10 percent of previous yields

as a result of insect infestations of host trees and

changes in forest management (Hosford and others

1997).

Biological and Management Concerns

The magnitude of the commercial mushroom harvest

in the Pacific Northwest has raised controversy about

conservation of the mushroom resource, particularly

regarding harvest effects on species viability and

ecosystem function. As part of a comprehensive

research program for managing commercial mushroom

harvests, we have conducted several public

workshops to gather information on the primary

concerns of resource managers, the mushroom

industry, and general public. Those concerns fall into

four categories; we present them as a series of

questions so readers can envision the scale and

complexity of needed mycological research.

53
Production and distribution-How many fruiting

bodies are being produced? How are they distributed

across the landscape or within a given area? How does

production vary during a season and among years?

What is the actual or potential commercial productivity

of a given area? What proportion of forest habitat is

available and accessible for harvesting? What factors

determine productivity and how might they be

managed?

Mushroom harvesting-How can the sustainability of

mushroom harvesting be assured? What proportion of

the crop can be harvested without unacceptable impacts

on the fungus itself or other resources? What

techniques will mitigate those impacts; does mushroom

harvesting increase or decrease subsequent production?

Is spore dispersal reduced by removal of immature

mushrooms, and if so, does it impair reproductive

success? Are fungal mycelium and subsequent

mushroom production affected by search and harvest

techniques such as raking, moving woody debris, and

digging? Are mushrooms harmed by numerous

harvesters trampling the forest floor? How important

are commercially valuable species as food for wildlife,

and is human competition for the resource significant?

Land management-How do various timber harvesting

methods (clearcutting, thinning to various densities, host

species selection) affect subsequent mushroom

production over time? What is the impact of soil

compaction or disturbance from logging activities?

What are the relations between fire and subsequent

mushroom production, especially for morels? How does

the intensity and timing of fire influence edible

mushroom production? What impacts do grazing,

fertilization, or pesticide application have on production?

Can mushroom production be improved through habitat

manipulation; for example, planting tree seedlings

inoculated with specific fungi, thinning understory brush

for sunlight and rainfall penetration, prescribed burns,

and irrigation? Can production be increased across the

landscape by managing forests to attain tree age class,

structure, and composition optimal for fruiting?

Biology and ecology-What are the important

reproductive events in the life cycle of a particular

fungus species? How are new colonies or populations

established and maintained? What causes them to

diminish or perish? How important is spore dispersal to

reproductive success, population maintenance, genetic

diversity, or adaptability to unique microhabitats? How

much genetic diversity exists within and among

populations? Are there endemic, narrowly adapted, or

unusual populations of otherwise common species?

What are the growth rates of fungal colonies in soil and

degree of mycorrhizal development by specific fungi on

root systems? To what degree do other mycorrhizal or

saprophytic fungi compete with desired fungi for

colonization sites on host roots or space in the forest

soil?

Answering these questions will require scientific

investigations on an array of ecological scales (soil

microniches to regional landscapes) and using a

variety of investigative methods.

Monitoring Needs and Methodological

Considerations

Numerous biotic and abiotic factors influence fruiting

of ectomycorrhizal fungi, and sporocarps of different

species differ in abundance and distribution from year

to year (Vogt and others 1992). Sporocarps are

typically clustered and associated with particular

substrates. Fruiting is generally nonuniform and

frequently occurs in arcs or patches of a few to

numerous sporocarps. The distribution of patches

also can range widely from a few scattered groups to

concentrated clusters. Commercially harvested

species such as American matsutake, Cantharellus

species, and Boletus edulis also associate with a broad

range of host tree genera in the Pacific Northwest

(Abies, Arbutus, Lithocarpus, Pinus, Pseudotsuga,

Quercus, Picea, and Tsuga) and occur over a range of

forest habitats and tree age classes. This broad

spatial and temporal variability of sporocarp

occurrence, combined with insufficient ecological

knowledge of the mycorrhizal fungi, presents a

formidable challenge for mycological research.

The first step to meeting this challenge is establishing

long-term monitoring of the fungal resource over the

regional land base. Three types of monitoring are

required: detection, evaluation, and research. Each

type addresses different concerns and provides data

to meet specific management objectives.

Detection monitoring addresses questions of

productivity and distribution. It documents the

current production of commercial fungi and variability

with space and time. This baseline information is

essential to assess whether production is declining,

increasing, or stable. Because annual fruiting typically

occurs over a short period and varies from year to

year, sampling must be repeated at least weekly

throughout the fruiting period and over several years

to assess seasonal production. Total seasonal

production must be tracked for years to detect trends

among high levels of annual variation. Numbers of

mushroom and their weights must be recorded as well

as commercial quality and current market prices, so

that total commercial value can be determined.

Detection monitoring also requires establishment of

control areas where mushroom harvest is restricted.

The scattered distribution of fungal sporocarps in

forest stands and the extensive land area needing to

54
be sampled present the greatest difficulty for

establishing a regional monitoring program. Most

mushroom-detection sampling designs use either

transects (long, narrow strip plots) or quadrats (square

or circular plots); each approach has strengths and

weaknesses (Vogt and others 1992). We use a variety

of transect methods. Some transects serve as

permanent plots and are repeatedly sampled throughout

the fruiting season for several years. Others are

temporary, with either randomly chosen locations, or

locations chosen to systematically sample a given land

area; for example, numerous parallel transects. They

allow search of extensive areas for sporocarps. Once

located, sporocarp distributions can be mapped for later

ecological study of the fungal colonies. Compared to

square or circular quadrat sampling plots, transects

increase the likelihood of detecting mushrooms in areas

where fungal fruiting is patchy or is responding to

fine-scale changes in abiotic or biotic conditions within a

stand. Narrow transects 1 to 2 meters wide reduce

trampling effects and human error in detection, because

the harvester can easily see and reach into the plot

without disturbing the soil. We are testing transects of

variable width and length to determine which best

assess mushroom productivity and distribution in

different forest habitats and topographic conditions.

Evaluation monitoring addresses questions about the

effects of mushroom harvest and forest management

practices. It relies on trends in data from detection

monitoring. For example, if detection monitoring

indicates declining levels of mushroom harvest,

additional evaluation monitoring is needed to determine

the extent and cause of the impact. Evaluation

monitoring focuses on specific concerns such as

mushroom harvest level, type and intensity of site

disturbance and changes in forest structure and

composition. Forest managers can use results from

evaluation monitoring to assess impacts of these

activities and then modify management approaches

accordingly. Evaluation monitoring also can include

studies that evaluate strategies for increasing production

of commercial mushrooms.

As an example, to evaluate the effects of various

mushroom harvest techniques on future mushroom

productivity of American matsutake, we are conducting

evaluation monitoring studies in four areas in Oregon.

Harvesters sometimes disturb soil layers looking for the

more valuable immature matsutake, which causes

concern for future fruiting. Specific treatments in order

of increasing soil and mycelium disturbance include (1)

control plot - no mushroom harvest; (2) mushroom

harvest with minimal disturbance of soil and fungus

mycelium - individual mushrooms are located and gently

pulled from the soil; (3) after a matsutake colony is

located, the litter and humus are removed over the

entire colony by raking the surface organic material to

the depth of the mycelium and pulling it to the side of

the colony - the mushrooms are then pulled from the

soil and the litter and duff replaced over the colony;

(4) same procedure as treatment three, but the litter

and duff are not replaced over the colony; (5) same

procedure as treatment four, but deep raking occurs

throughout the mycelium, thus completely disrupting

the roots and fungal network - disturbed layers are

replaced; and (6) deep raking but soil layers are not

replaced. In another study, 5 liters of water per

square meter of fruiting area was added at 2-week

intervals throughout the fruiting season as a fruiting

enhancement technique.

Research monitoring examines basic biological and

ecological phenomena and presents a myriad of

investigative opportunities for mycological research.

In addition to specific studies on their mycorrhizal

ecology, research would examine the functional roles

of these fungi in forest ecosystems. For example,

how do these fungi contribute to nutrient cycling, soil

aggregation, and overall site productivity? These

investigations can be conducted concurrently within

detection and evaluation monitoring sites where

environmental and forest stand data are available.

Modern methods and techniques, such as molecular

DNA tools and in vivo root observation chambers, can

be used to great advantage in these studies to better

understand mycorrhiza dynamics.

Future Challenges

The difficulty in assessing cause and effect for the

decline for some ectomycorrhizal fungi in Europe

emphasizes the need for establishing long-term fungal

inventory and monitoring programs. The intense

commercial harvest of edible ectomycorrhizal fungus

sporocarps heightens this need in the Pacific

Northwest. Long-term monitoring is essential to

assess the "carrying capacity" of the land and to

develop ecological models to predict future mushroom

productivity. Changes noted in mushroom

productivity also can serve as biological indicators of

large-scale changes in forest health due to

management practices, pollution, or global climate

change.

The greatest challenge in this endeavor for

mycologists and forest managers is dealing with broad

landscape, regional-scale, and long-term questions.

Most field studies of mushroom production and fungal

species diversity focus at the stand level and are of

short duration (Pilz and Molina 1996, Vogt and others

1992). Few fungal community studies address

questions of habitat disturbance or effects of

mushroom harvest on future sporocarp production,

and available data thus relate little to managing

widespread commercial harvesting. To meet these

challenges and provide the essential information

needed to wisely manage fungal resources,

55
mycologists must work closely with other scientists,

forest managers, and interested publics to conduct

cooperative studies within an ecosystem framework.

Several such integrated studies are underway in the

Pacific Northwest (see Pilz and Molina 1996).

QUESTIONS FROM THE SYMPOSIUM

FLOOR

Q: Can we ever hope to characterize system health

from the fungal diversity that would be expected at

different times over succession of a forest?

A: Yes, I think this is a realistic goal and we are

conducting research on that topic. We have found that

several species of fungi occur during different periods of

forest community development. Some are only found in

old, mature forests and often associate with legacy

structures such as buried wood from fallen trees or deep

organic matter. The bright yellow ectomycorrhizal

fungus, Piloderma croceum, is one such fungus. We

believe the presence of this fungus during later stages of

forest community development could serve as an

indicator of a healthy, functioning forest ecosystem.

One of our goals over the next several years is to build

habitat based models of fungal diversity. Such models

would provide specific habitat characteristics that

predict presence and function of particular fungal

species or species guilds. If used at a broad landscape

scale, such models could provide valuable information to

managers and enable them to select management

options that maintain or create critical fungal habitat.

Q: What is the health and functionality of mycorrhizal

mats in moderate to heavily compacted skid trails? If

these mats are still functionally, active, does subsoiling

negatively impact these mats?

A: Ectomycorrhizal fungi are strongly aerobic so soil

compaction can have a negative effect on presence and

function of ectomycorrhizal fungi. We have studied the

production of ectomycorrhizal fungus sporocarps in

compacted vs. noncompacted soils and production is

less in compacted soils. Also, fungal diversity differs

between these two soil types and is typically lower in

compacted soils. Those fewer species found in

compacted soils are likely adapted to such disturbed

conditions. Although we have not conducted

quantitative studies on ectomycorrhizal root tip

abundance in compacted soils, other studies have

shown diminished rooting of tree seedlings in compacted

soil.

It follows that ectomycorrhizae would also be diminished

and subsequently the activity of mycorrhizal fungus

hyphae. We have not seen typical production of

ectomycorrhizal fungal mats in compacted soil as is

common in most forest soils. Subsoiling could have a

direct negative effect on fungi that are trying to grow

in compacted soil. But more likely, subsoiling would

be beneficial in the long run by providing better soil

aeration. Specific studies are needed to address longterm effects of soil compaction and subsoiling

treatments on mycorrhizal functioning.

Q: What is the relationship, if any, between the

decay of coarse woody debris, the formation of

humus, and the health of mycorrhizal fungi?

A: The presence of coarse woody debris and humus

formation is particularly critical to the health and

function of ectomycorrhizal fungi. Many

ectomycorrhizal fungi (and their ectomycorrhizae)

develop exclusively or fruit extensively in well-rotted,

buried wood or deep humus. More importantly, work

in Idaho and southern Oregon (see chapters by Al

Harvey and Mike Amaranthus, this proceedings)

shows that in arid forest ecosystems, ectomycorrhizae

are most abundant in buried wood compared to

mineral soil. Such substrates have strong moisture

holding capacity so they may serve as critical water

resources during times of drought. Also, recent

studies show that ectomycorrhizal fungi have greater

enzymatic capabilities to degrade organic nitrogen and

carbon than was previously thought. Thus, coarse

woody debris and humus material provide a direct

food source for mycorrhizal fungi and a nutrient

source for their associated host trees. Maintaining

such coarse woody debris is critical for sustaining the

healthy function of the diverse forest mycoflora.

LITERATURE CITED

Amaranthus, M.P.; Molina, R.; Trappe, J. 1990.

Long-term productivity and the living soil. In:

Perry, D.A.; Meurisse, R.; Thomas, B.; and others

(eds.). Maintaining long-term productivity of Pacific

Northwest ecosystems. Portland, OR: Timber

Press: 36-52.

Amaranthus, Michael; Pilz, David. 1996. Productivity

and sustainable harvest of wild mushrooms. In

Pilz, D; Molina R. (eds.). Managing forest

ecosystems to conserve fungus diversity and

sustain wild mushroom harvests. Gen. Tech. Rep.

PNW-GTR-371. Portland, OR: U.S. Department of

Agriculture, Forest Service, Pacific Northwest

Research Station: 42-61.

Arnolds, Eef. 1991. Decline of ectomycorrhizal fungi

in Europe. Agriculture, Ecosystems and

Environment. 35:209-244.

56
Arnolds, Eef. 1995. Conservation and management of

natural population of edible fungi. Canadian Journal

of Botany. 73 (Supplement 1): S987-S998.

Avise, John C. 1994. Molecular markers, natural

history, and evolution. New York: Chapman and Hall.

Avise, John C; Hamrick, James L, eds. 1995.

Conservation genetics. New York: Chapman and

Hall.

Baar, J.; Ozinga, W.A.; Kuyper, T.H.W. 1994. Spatial

distribution of Laccaria bicolor genets reflected by

sporocarps after removal of litter and humus layers in

a Pinus sylvestris forest. Mycological Research. 98:

726-728.

Castellano, M.A.; Molina, R. 1989. Mycorrhizae. In:

Landis, T. D.; Tinus, R.W.; McDonald. S.E.; Barnett,

J.P. (eds.). The container tree nursery manual,

volume 5. Agric. Hndbk. 674. Washington, DC: U.S.

Department of Agriculture, Forest Service: 101-167.

Cooke, R.C.; Whipps, J.M. 1993. Ecophysiology of

saprotrophic fungi. London: Blackman Scientific.

Dahlberg, Anders; Stenlid, Jan. 1990. Population

structure and dynamics in Suillus bovinus as

indicated by spatial distribution of fungal clones.

New Phytologist. 115: 487-493.

Danell, Eric 1 994. Cantharellus cibarius: mycorrhiza

formation and ecology. Uppsala, Sweden: University

of Uppsala. Ph.D. Thesis.

Dizon, Andrew E.; Taylor, Barbara L; O'Corry-Crowe,

Gregory M. 1995. Why statistical power is

necessary to link analyses of molecular variation to

decisions about population structure. In. Nielsen,

J.L.; Powers, D.A. (eds.). Evolution and the aquatic

ecosystem: defining unique units in population

conservation. Bethesda, MD: American Fisheries

Society.

Forest Ecosystem Management Assessment Team.

1993. Forest ecosystem management: an ecological,

economic, and social assessment. Portland, OR: U.S.

Department of Agriculture; U.S. Department of the

Interior [and others]. [Irregular pagination].

Gryta, H.; Debaud, J.-C; Effosse, A.; Gay, G.;

Marmeisse, R. 1997. Fine-scale structure of

populations of the ectomycorrhizal fungus Hebeloma

cylindrosporum in coastal sand dune forest

ecosystems. Molecular Ecology. 6: 353-364.

Hastings, Alan; Harrison, Susan. 1994. Metapopulation

dynamics and genetics. Annual Review of Ecological

Systematics. 25: 167-188.

Hosford, David; Pilz, David; Molina, Randy; Amaranthus,

Michael. 1997. Ecology and management of the

commercially harvested American matsutake

mushroom. Gen. Tech. Rep. PNW-GTR-412.

Portland, OR: U.S. Department of Agriculture,

Forest Service, Pacific Northwest Research Station.

68 p.

Maser, Zane; Maser, Chris; Trappe, James. 1985.

Food habits of the northern flying squirrel

(Glaucomys sabrinus) in Oregon. Canadian Journal

of Zoology. 63:1084-1088.

McNeely, Jeffrey A.; Miller, Kenton R.; Reid, Walter

V.[and others]. 1990. Conserving the world's

biological diversity. Gland, Switzerland:

International Union for the Conservation of Nature.

Miller, Steven L. 1995. Functional diversity in fungi.

Canadian Journal of Botany. 73 (Supplement 1):

S50-S58.

Milligan, B.G.; Leebens-Mack, J.; Strand, A.E. 1994.

Conservation genetics: beyond the maintenance of

marker diversity. Molecular Ecology. 3:423-435.

Molina, Randy; Amaranthus, Michael. 1991.

Rhizosphere biology: ecological linkages between

soil processes, plant growth, and community

dynamics. In: Proceedings, Management and

productivity of western-montane forest soils; April

10-12; Boise, Idaho. Gen. Tech. Rep. INT-280.

Ogden,UT. U.S. Department of Agriculture, Forest

Service: 51-58.

Molina, Randy; Massicotte, Hugues; Trappe, James

M. 1992. Specificity phenomena in mycorrhizal

symbioses: community-ecological consequences

and practical implications. In: Allen, Michael (ed.).

Mycorrhizal functioning, an integrative plant-fungal

process. New York: Chapman and Hall: 357-423.

Molina, Randy; O'Dell, Thomas; Luoma, Daniel; and

others. 1993. Biology, ecology, and social

aspects of wild edible mushrooms in the forests of

the Pacific Northwest: a preface to managing

commercial harvest. Gen. Tech. Rep.

PNW-GTR-309. Portland, OR: U.S. Department of

Agriculture, Forest Service, Pacific Northwest

Research Station. 42 p.

Moritz, Craig 1994. Applications of mitochondrial

DNA analysis in conservation: a critical review.

Molecular Ecology. 3:401-411.

O'Dell, Thomas E; Castellano, Michael A.; Trappe,

James M. 1992a. Biology and application of

ectomycorrhizal fungi. In: Metting, Jr., F. Blaine

(eds.). Soil microbial ecology: applications in

agricultural and environmental management. New

York: Marcel Dekker, Inc.: 379-416.

O'Dell, Thomas E.; Luoma, Daniel L.; Molina,

Randolph J. 1992b. Ectomycorrhizal fungal

57
communities in young, managed, and old-growth

Douglas-fir stands. Northwest Environmental

Journal. 8:166-168.

Parks, C.G.; Shaw, D.C. 1997. Death and decay: a

vital part of living canopies. Northwest Science

Special Issue: 46-53.

Perry, D.A.; Amaranthus, M.P.; Borchers, J.G.; and

others. 1989a. Bootstrapping in ecosystems.

BioScience. 39:230-237

Perry, D.A.; Margolis, H.; Choquette, C; and others.

1 989b. Ectomycorrhizal mediation of competition

between coniferous tree species. New Phytologist.

112:501-511.

Perry, David A.; Molina, Randy; Amaranthus, Michael P.

1987. Mycorrhizae, mycorrhizospheres, and

reforestation: current knowledge and research needs.

Canadian Journal of Forest Research. 17:929-940.

Pilz, D.; Molina, R. (eds.). 1996. Managing forest

ecosystems to conserve fungus diversity and sustain

wild mushroom harvests. Gen. Tech. Rep.

PNW-GTR-371. Portland, OR: U.S. Department of

Agriculture, Forest Service, Pacific Northwest

Research Station. 104 p.

Pilz, David; Molina, Randy; Mayo, Jim. 1996.

Chanterelle production responses to stand thinning.

In Pilz, D.; Molina, R. (eds.). Managing forest

ecosystems to conserve fungus diversity and sustain

wild mushroom harvests. Gen. Tech. Rep.

PNW-GTR-371. Portland, OR: U.S. Department of

Agriculture, Forest Service, Pacific Northwest

Research Station: 81-82.

Read, D.J.; Francis, R.; Finlay, R.D. 1985. Mycorrhizal

mycelia and nutrient cycling in plant communities.

In: Fitter, A.H.; Atkinson, D.; Read, D.J.; Usher, M.B.

(eds.). Ecological interactions in soils. Oxford:

Blackwell Scientific Publications: 193-218.

Schlosser, W.E.; Blatner, K.A. 1995. The wild edible

mushroom industry of Washington, Oregon and

Idaho: a 1992 survey. Journal of Forestry. 93:

31-36.

Schowalter, Timothy; Hansen, Everett; Molina, Randy;

Zhang, Yanli. 1997. Integrating the ecological roles

of phytophagous insects, plant pathogens, and

mycorrhizae in managed forests. In: Kohm, Kathryn

A.; Franklin, Jerry F., eds. Creating a forestry for the

21st century; the science of ecosystem

management. Washington, DC: Island Press:

171-189.

Simard, Suzanne W.; Perry, David A.; Jones, Melanie

D.; and others. 1997. Net transfer of carbon

between ectomycorrhizal tree species in the field.

Nature. 388: 579-582.

Smith, Myron L.; Bruhn, Johann N.; Anderson, James

B. 1992. The fungus Armillaria bulbosa is among

the largest and oldest living organisms. Nature

356:428-431.

Smith, S.E.; Read, D.J. 1997. Mycorrhizal symbiosis.

London: Academic Press. 605 p.

Tilman, David; Downing, John A. 1994. Biodiversity

and stability in grasslands. Nature. 367:363-365.

Todd, N.K.; Rayner, A.D.M. 1980. Fungal

individualism. Science Progress (Oxford).

331-354.

66:

Trappe, J. 1977. Selection of fungi for

ectomycorrhizal inoculation in nurseries. Annual

Review of Phytopathology. 15:203-222.

Trappe, James M.; Luoma, Daniel L. 1992. The ties

that bind: fungi in ecosystems. In: Carroll, George

C; Wicklow, Donald T. (eds.). The fungal

community, its organization and role in the

ecosystem. New York: Marcel Dekker, Inc.:

17-27.

U. S. Department of Agriculture, Forest Service; U. S.

Department of the Interior, Bureau of Land

Management. 1994b. Record of decision for

amendments to Forest Service and Bureau of Land

Management planning documents within the range

of the northern spotted owl. [Place of publication

unknown]. 74 p.

Vogt, Kristina A.; Bloomfield, Janine; Ammirati,

Joseph F.; Ammirati, Susan R. 1992. Sporocarp

production by basidiomycetes, with emphasis on

forest ecosystems. In: Carroll, George C; Wicklow,

Donald T. (eds.). The fungal community, its

organization and role in the ecosystem. New York:

Marcel Dekker, Inc.: 563-581

Wiens, John A. 1989. Spatial scaling in ecology.

Functional Ecology 3:385-397.

Wiens, John A. 1997. Metapopulation dynamics and

landscape ecology. In Hanski, Ilkka A; Gilpin,

Michael E. (eds.). Metapopulation biology: ecology,

genetics, and evolution. London: Academic Press.

58
SOIL BACTERIA: A DYNAMIC POOL OF SOIL ORGANIC MATTER AND CATALYSTS OF

KEY BELOWGROUND PROCESSES

Jeanette M. Norton1

ABSTRACT

Bacteria are key components of the soil ecosystem forming the basis for the soil food web and mediating important

biogeochemical reactions. Bacteria are the most numerous soil organisms in terms of numbers of individuals, but in

most wildland ecosystems bacteria comprise a smaller portion of the microbial biomass than the fungi. The bacterial

biomass is a portion of the soil organic matter that has a relatively rapid turnover time and therefore has the potential

for releasing quantitatively important plant available nutrients during turnover (death). The active soil bacteria can be

distinguished from inactive by their ability to respire and process carbon, but microscopic methods for this

differentiation are too labor intensive or expensive for most ecosystem level (field) studies. The active bacteria have an

incredible metabolic diversity and their metabolic types may be classified on the basis of their sources of energy, carbon

and reducing power. The types of bacterial metabolism present in the soil are dependent on the oxygen status (redox

potential) in the soil environment. Specific bacterial metabolic products are indicators of anaerobic (oxygen deprived)

soil conditions. The presence of active bacteria with a specific metabolic potential combined with the appropriate

environmental conditions results in the activity of bacteria as catalysts for key soil processes. The chemolithotrophic

bacteria are especially important for their major role in the biogeochemistry of nitrogen, sulfur and iron.

Keywords: soil bacteria, bacterial biomass, microbial biomass, diversity, redox potential, biogeochemistry

INTRODUCTION

Bacteria are the most numerous of the living organisms

on the surface of the earth and are the most diverse.

In a single gram of surface soil there are generally in

excess of a billion individuals (Metting 1993),

potentially representing more than 1000 different types

or species (Torsvik and others 1990). Bacteria and

fungi together comprise the soil microflora and are the

foundation for the soil food web (Bloem and others

1997). The products of primary production, both above

and below-ground plant debris, must pass through the

soil microflora before entering the soil organic matter.

Jenkinson (1977) described the microbial biomass as

the "eye of the needle through which all organic

material that enters the soil must pass." The soil

microbial biomass carbon comprises approximately 2-5

percent of the soil organic carbon in a wide range of

surface soils (Smith and Paul 1990).

The soil bacteria are incredibly diverse both in their

types and their metabolic capabilities. The bacteria can

use a wider range of compounds for the production of

energy and biomass than any other group of organisms.

Soil bacteria can be active and mediate the

transformations of carbon and nutrient elements under

environmental conditions that restrict the activities of

other organisms. Therefore, the bacteria are both a

significant portion of the soil microbial biomass and the

catalysts of key belowground processes in carbon and

nutrient cycling.

BACTERIA AS A POOL OF SOIL

ORGANIC MATTER

The soil organic matter (S0M) is comprised of both

living and non-living components and varies widely in

amount and quality, depending on the soil type and

location. On a broad scale, the amount of soil

microbial biomass correlates to the amount of SOM,

with the microbial biomass carbon accounting for 2-5

percent of the soil organic C (Smith and Paul 1 990,

Martens 1995). Bacteria comprise only a portion of this

microbial biomass and the dominance of bacteria

versus fungi varies with soil and ecosystem type,

location and depth and previous management history.

Quantifying Soil Bacteria

The numbers of bacteria in the soil varies widely in

different ecosystems but is grossly correlated with the

level of organic C in the soil. Numerous attempts to

compare total microscopic counts of bacteria to viable

counts on media plates have shown that most of the

1

Jeanette M. Norton is an Assistant Professor of soil microbiology in the Department of Plants, Soils and Biometeorology at Utah

State University, Logan UT 84322-4820.

59
bacteria in soils are not recoverable on standard

microbiological media (Torsvik and others 1990,

Bakken 1997). Therefore direct microscopy is the

preferred method for determining total numbers of

bacteria in any given soil, and plate counts are of very

limited value. Epifluorescence microscopy has the

advantage of allowing the visualization of bacteria

against the opaque background of soil particles and

allowing the visualization of bacteria in the smaller

(<0.1 μm

3

) size classes. The currently recommended

fluorescent stains and procedures have been

summarized (Bottomley 1994, Bloem and others1995).

The numbers of bacteria in defined size classes can be

used to convert bacterial numbers to biovolume and

biomass estimates. Recent advances in image analysis

and confocal microscopy may make direct counting and

size analysis of fluorescently stained bacteria more

efficient and less subject to microscopist bias (Bloem

and others 1995).

Techniques are available for differentiating active from

inactive bacterial cells in soils (Norton and Firestone

1991, Bottomley 1994, Hartmann and others 1997).

Using these methods one can estimate the portion of

the bacterial biomass that is considered active using an

operational definition. Typical values vary around 50

percent but are dependent on the soil type, organic

matter availability and proximity to roots (for example

see Norton and Firestone 1991). The complexity of the

methods available and the high spatial and temporal

variability of active bacteria means that determination

of the quantity of active bacteria is currently too

laborious and expensive for most field studies.

As an alternative to microscopic methods,

physiological methods have been used to examine the

quantity and activity of soil bacteria. In the widely used

substrate induced respiration (SIR) method, the

respiratory response of soil microorganisms to glucose

substrate addition, in combination with selective

antibiotics, is measured and related to the contribution

of bacteria and fungi to soil respiration (Anderson and

Domsch 1973, 1975). SIR values are also typically

converted to microbial biomass C and used to derive

fungal to bacterial ratios. Discrepancies between

values determined by microscopy and SIR techniques

(Bardgett and others 1996, Velvis 1997) and the

numerous assumptions typical of both microscopic and

SIR determinations has made cross site comparisons

and interpretations of active bacteria values extremely

difficult. Selected values for bacterial biomass C as

determined by fluorescent microscopy and the SIR

method are shown in table 1.

Turnover of the Bacterial Biomass Releases

Nutrients

The death and decomposition of microbial cells and

their incorporation into new microorganisms or into the

SOM is known as microbial turnover. The nutrient

content of living microorganisms is generally higher

than that of the total SOM, and as these

microorganisms die the nutrients released become

available for plant uptake. Conversely, the growth of

the microbial biomass can incorporate or immobilize

nutrients, making these unavailable for plant uptake.

Estimates of the size of the microbial biomass and its

nutrient content, together with calculated turnover

times, reveal that the release of nutrients from

microbial turnover is large enough to meet plant

demands in many ecosystems (Smith and Paul 1990).

The microbial biomass is estimated to turnover

approximately 2-5 times per year (Smith and Paul

1990) although the bacterial turnover rate is thought to

be significantly more rapid. For example, in an

agricultural and a forest soil the bacterial population

was calculated to turnover between 14-28 times

annually (Baith 1990). These turnover values

emphasize that the bacteria in soil are a significant

source and sink for plant nutrient elements.

Turnover of bacterial cells in soil is caused by both

biotic and abiotic factors. Grazing or predation by soil

fauna such as protozoans and nematodes plays an

important role in releasing the nutrients found in the

microbial biomass (Bloem and others 1997). When

grazers and predators consume the relatively nutrient

rich bacteria, excess nutrients are released in plant

available forms, such as ammonium and phosphate.

Abiotic factors, such as wetting and drying cycles and

physical soil disturbances (for example tillage, timber

harvest, fire), cause various levels of microbial

turnover. Soil wetting and drying events have been

shown to affect bacterial numbers and biomass values,

with moist soils favoring bacteria growth (Clarholm and

Rosswall 1980, Lundgren and Soderstrom 1983, van

Veen and others 1985). For example, in the humus

layer of a pine forest, a peak in bacterial numbers

followed within 3 days of wetting rainfall events

(Clarholm and Rosswall 1980). Freeze-thaw cycles may

also release some of the plant available nutrients

contained in bacterial cells. Soil disturbance which

incorporates plant residues such as tillage or those

accompanying timber harvesting may cause rapid

growth of the soil microorganisms and then subsequent

death and release of nutrients to the soil.

60
Table 1 —Bacterial biomass and fungal/bacterial (F/B) ratios in surface soils.

a

 Methods: direct microscopy (DM) or substrate induced respiration (SIR).

b

 Staben and others (1997).

c

 Norton (1987) unpublished data.

d

 Griffiths and others (1997).

e

 Berg and others (1998).

f

 Not available.

9

 Vishnevetsky and Steinberger (1997).

BACTERIA AS CATALYSTS OF KEY SOIL

PROCESSES

Energy and elemental cycling through the microbial

biomass is the basis for plant residue decomposition

and biogeochemical nutrient cycling. The portion of the

microbial biomass active in the transformation of a

substrate can be described as the catalyst for the

specific process. In enzyme kinetic theory the

dependence of the rate of substrate depletion on both

the substrate concentration and the microbial catalyst

would be described as a second-order reaction

according to the following equation:

where k1 is the apparent rate constant and the reaction

is a pseudo-first-order reaction.

In the soil environment, the microbial catalyst may be

considered constant only over relatively short time

periods, and annual cycles in the size of the microbial

catalyst population should be considered for reliable

estimates and models (Smith and Paul 1990). The

reactions catalyzed by the soil bacteria include a wide

diversity of processes which cycle carbon from plant

primary production into the SOM and transform

nutrient elements. While many bacteria are generalists

and are capable of varied processes, some bacterial

groups have specific metabolisms and a restricted

range of transformation which they mediate. The next

61

System type Treatment or soil type

Agriculturalb

conservation reserve

wheat fallow

Mixed Conifer Forestc

old growth H-layer

clearcut mineral soil

Pacific NW Ripariand (summer)

mixed forest

pasture

Scots Pine Foreste (seasonal mean)

L-layer

F-layer

H-layer

Desertg

wet season

dry season

Location

WA, USA

CA, USA

OR, USA

Wekerom, NL

Negev, Israel

Methoda

DM

DM

DM

DM

SIR

Bacterial

biomass

(dry wt)

mg kg

 -1

 soil

29

25

590

80

1200

800

302

308

252

200

30

F/B

biomass

2.9

2.1

2.2

7.9

8.8

3.3

NAf

0.6

1.2

concentration. If the microbial biomass concentration

is constant over time then the above equation reduces

to:

where is the rate of change of the substrate (S), k

is the rate constant and B is the microbial catalyst
section classifies the metabolic types found in the soil

bacteria.

Classification of the Metabolic Diversity of

Bacteria

The systematic classification of metabolic types of

organisms is based on three basic biochemical

requirements: 1) an energy source, 2) a source of

carbon and 3) a source of reducing equivalents. This

classification scheme is summarized in table 2. Each of

these categories has a specific prefix that is amended

to the root, troph (from the Greek trope meaning

nourishment) to yield combinatorial names for the

specific metabolic types. For example, many common

soil bacteria use organic materials as a source of

energy, reducing equivalents and cell carbon. These

organisms are then classified as chemoorganoheterotrophs which is often shortened to simply heterotrophs.

The cyanobacteria important in the formation of

microbiotic crusts are generally photolithoautotrophs,

using photosynthesis and C02 fixation to colonize their

harsh environment. Some important bacteria in biogeochemical cycling are chemolithotrophs which use

chemical oxidation as the source of energy and

inorganic compounds as the source of reducing

equivalents. Some representative genera of bacteria

important in the cycling of nitrogen, iron, sulfur and

methane are listed in table 3. Many of these chemolithotrophs are autotrophs and capable of fixing CO2 for

their C source. The chemolithoautotrophic bacteria

include the ammonia oxidizing bacteria such as

Nitrosomonas important in the process of nitrification.

The chemolithotrophic bacteria are of crucial

importance to ecosystem processes and control the

abundance and form of major nutrient elements, such

as N, S, Fe and Mn through their oxidation and

reduction reactions (Haack 1996). Examples of these

bacterial processes are given in the following sections.

Table 2 —Classification of Organisms based on Energy

Source, Carbon Source and Source of Reducing Equivalents

Energy Source:

Light = phototroph

Oxidation of organic or inorganic compounds

chemotroph

Carbon Source:

Inorganic = autotroph

Organic = heterotroph

Source of Reducing Equivalents:

Inorganic = lithotroph

Organic = organotroph

Bacteria may also be categorized based upon their

requirement for oxygen for growth and the production

of ATP. Bacteria which require oxygen are aerobes and

have respiratory metabolisms with O2 as the terminal

electron acceptor. Anaerobic respiration uses other

inorganic oxidized compounds such as nitrate or sulfate

as alternative electron acceptors, while fermentation

uses internal electron transfer between organic

compounds. The facultative anaerobes can grow with

or without oxygen and may be particularly important in

controlling the oxygen availability in soil environments.

While the metabolic classification of bacteria has

proven useful, it is important to remember that some

organisms in nature do not respect these rigid

categories and switch from one type of metabolism to

another, dependent on environmental conditions.

Environmental Constraints on Bacterial

Metabolism

The types of bacterial metabolism that are active in a

soil are highly dependent upon the availability of

oxygen, which is primarily controlled by the water

status of the soil. As O2 is used by microbial

metabolism it is replaced by diffusion from the

atmosphere a process which is extremely slow through

water filled pores. The intensity of a soil's reduction or

oxidation is known as the redox potential and is a

measure of the availability of electrons. Aerated soils

have positive redox potentials varying from around

+ 400 to +700 mV, while waterlogged reduced soils

have negative redox potentials typically as low as -200

to -300 mV. The redox potential of soils interacts with

soil pH and sets limits on the redox and pH values

found in soil systems (fig. 1, Bohn and others 1985).

The redox potential decreases as microbial metabolism

shifts from aerobic to anaerobic respiration (that is

denitrification, sulfate respiration) and eventually to

fermentation. A representative series of microbial

reactions and their reference redox potentials is shown

in table 4. The bacterial groups responsible for a wide

variety of transformations in soil are active in different

ranges of redox potential (Eh) and pH as shown in

figure 2. Some of these processes are discussed

individually in the following section. The abiotic

controlling factors of temperature and water potential

on the transformation of organic matter and on

microbial metabolism in general are outside of the

scope of this report but are summarized by Paul and

Clarke (1989).

Examples of Key Soil Processes Unique to

Bacterial Metabolism and their Relevance to

Ecosystem Management

The following examples have been chosen as examples

of processes mediated by bacteria that may be of key

62
Table 3 —Examples of chemolithotrophic bacteria and their functions in the soil.a

Figure 1 —The range of redox potentials and pH values in

soils. From Bohn and others 1985, Soil Chemistry based on

Bass Becking and others 1960, J. Geology 68: 243, 1960.

importance to ecosystem function and management.

The role of bacteria as generalized decomposers of

plant residues is not discussed here, but see Bloem and

others (1997) for a discussion of the soil food web.

Several of the processes described are involved in the

N cycle. Since N is the most often limiting nutrient to

plant growth, bacterial N transformations are of

particular importance.

Chemolithotrophic nitrification —Nitrification is

defined as the biological conversion of reduced nitrogen

in the form of ammonia (NH3) or ammonium (NH4

 + ) to

oxidized N in the form of nitrite (NO2

.

) or nitrate (NH3

.

).

The conversion of a cation, NH4

+

, which is retained by

the negatively charged soil matrix to an anion (NO2 or

NO3), which is mobile in the soil, strongly influences

the fate of N in the ecosystem. In most soil

environments nitrification is mediated primarily by

chemolithotrophic bacteria which gain energy from the

oxidation of N. The oxidation of N takes place in two

steps: the ammonia oxidizing bacteria such as

Nitrosomonas and Nitrosospira convert NH3 to N02

according to:

NH3 + O2 + 2H+ + 2e. NH2 0H + H20 N02

-

 +5H++4e.

While the nitrite oxidizing bacteria such as Nitrobacter

convert N02 to NO3 :

NO2 + H2O NO3 + 2H+ + 2e.

Note that both reactions require O2 and produce acid in

the soil environment. The population of nitrifying

bacteria and their activity is generally limited by the

availability of their substrates, particularly NH4

+

 but

also O2. In systems where vegetation has been

removed (for example clearcuts, overgrazed range),

excess NH4 which was formerly taken up by plants or

heterotrophic microorganisms may become available for

nitrification. This stimulation of nitrification in

conjunction with diminished N03

-

 use can lead to

accumulations of NO3

-

 in soils (Vitousek and others

1982, Stark and Hart 1997). The potential for N loss

from the system by leaching and denitrification (see

below) increases under conditions of high NO3

-

availability and may have detrimental effects on water

quality and atmospheric chemistry. Maintaining the N

63

Element

N

Fe

S

C

Active under aerobic (oxidizing) conditions

Nitrifiers

Ammonia oxidizers: Nitrosomonas, Nitrospira

Nitrite Oxidizers: Nitrobacter

Iron Oxidizers

Gallionella, Sulfobacillus

Thiobacillus ferrooxidans

Sulfur Oxidizers

Thiobacillus, Thiothrix

Methane Oxidizers (C1 only)

Methylococcus, Methylocystis, Methylosinus

Active under anaerobic (reducing) conditions

Denitrifiers

Paracoccus denitrificans,

Thiobacillus denitrificans

Iron Reducers

Geobacter

Sulfur Reducers

Desulfovibrio, Desulfobacter

Methanoaens and Acetogens

Methanobacteriaceae, Acetobacterium

8

 from Haack 1996.
Table 4 — Thermodynamic sequence for reduction of inorganic substances at pH

7.0 and reference redox potentials.8

a

From Stevenson (1986).

Figure 2—Environmental limits of Eh and pH for some

bacterial groups. From Ehrlich 1990, Geomicrobiology based

on Bass Becking and others, J. Geology 68: 243, 1960.

fertility of a wildland site over the long-term will also

require management to retain N on the site or to

replace it by fertilization. In agricultural systems,

nitrification inhibitors such as nityrapyrin (trade name

N-serve, Dow Chemical Co.) are used to inhibit

nitrification for a relatively short time period

(approximately 4 to 8 weeks) and retain added fertilizer

N in the NH4

+

 form (Paul and Clark 1989). In wildland

systems, it is reasonable to control the accumulation of

excess N03

-

 by reducing the NH4

+

 available as

substrate or by encouraging the use of NO3

-

.

Management practices which maintain a high microbial

and plant demand for inorganic N (NH4

+

 and N03

-

)

include: 1) leaving woody residues on the soil surface ,

2) leaving adequate vegetation which can assimilate N,

and 3) managing grazing to prevent waste

accumulation in sensitive areas such as riparian zones

and to maintain plant cover. The goal of ecosystem N

management is to keep N onsite, in the NH4

+

 or organic

N forms, where it can be available for future plant

uptake.

Denitrification —Denitrification is defined as the

bacterial respiratory reduction of nitrate (N03

-

) or nitrite

(NO2) to the gaseous end products nitrous oxide (N2O)

and dinitrogen (N2). The denitrification process results

in the net loss of N from the ecosystem and is the

primary route by which N is returned to the atmosphere

from terrestrial systems. Denitrification occurs when

soils containing NO3 become depleted in O2 primarily

due to excess moisture. Under these conditions,

certain bacteria have the ability to use NO3

-

 as an

alternative electron acceptor for anaerobic respiration.

The capacity for denitrification has been found in at

least 23 genera of bacteria, most of which are

chemoheterotrophs (for example Pseudomonas,

Alcaligenes) although there are several

chemolithotrophs (for example Thiobacillus

denitrificans) which also have this capacity (Firestone

1982, Haack 1996). The general requirements for

rapid denitrification are 1) lack of O2, 2) supply of NO3,

3) available organic substrates (electron donors), 4)

biological temperature range, and 4) presence of

denitrifying bacteria. The management of denitrification

in order to maintain N in the soil system should focus

64

Reaction

Disappearance of O2 (aerobic respiration)

O2 + 4H+ + 4e - 2H2O

Reduction of NO3

-

 (1st step of denitrification)

NO3

-

 + 2H+ + 2e - NO2

-

 + H2

Reduction of Mn(IV) to Mn (II)

MnO2 + 4H+ + 2e- - Mn2 + 2H2O

Reduction of Fe(lll) to Fe(ll)

Fe(OH)3 + 3H+ + e- - Fe2+ + 3H2O

Reduction of sulfate to H2S gas

S04

2

 + 10H+ + 8e - H2S + 4H2O

Formation of CH4 (methanogenesis)

C02 + 8H+ + 8e- - CH4 + 2H2O

Eh (Volts)

0.816

0.421

0.396

-0.182

-0.215

-0.244
on maintaining the O2 supply (aerobic soil) and limiting

the accumulation of NO3 (see section of nitrification).

Management decisions should be focused on avoiding

soil compaction and excess water which will promote

the anaerobic conditions required for denitrification. It

should be noted that in situations where there is excess

N in the system (for example wastes, sewage)

denitrification is often used as a way to eliminate

excess N and return it to the atmosphere.

Sulfur oxidation and reduction-A summary of the

microbial cycling of sulfur in soils is shown in figure 3.

The oxidation and reductions of sulfur in soil are highly

dependent on O2 availability. In a transformation similar

to nitrification, chemolithotrophic bacteria can use the

oxidation of reduced sulfur compounds such as FeS

and FeS2 (iron sulfides and pyrite) and elemental S as a

source of reducing equivalents and energy. This

reaction generally requires available O2 and produces an

extremely acidic end product, H2SO4 (sulfuric acid) and

frequently associated with heavy metal contamination.

Bacteria in the genus Thiobacillus are examples of

bacteria which mediate and accelerate this process.

Sulfur oxidation can be devastating to the environment

when reduced S containing ores or sediments are

exposed to O2 by being brought to the surface, causing

acid drainage. Some examples include the exposure of

sulfide containing sediment layers in road cuts and the

activities associated with mining. Management

implications are to ensure that mine spoils containing

pyrite and others metal sulfides are properly managed

in designed waste dumps to prevent exposure to air

and acid drainage into surface waters. The BLM and

the U.S. Forest Service require that companies assess

the potential for acid mine drainage as part of the

permitting process (White and Jeffers 1994). The

other situation where sediments are rich in reduced S is

found in coastal regions with sediments previously

exposed to salt or brackish water. If these sediments

are drained and exposed to aerobic conditions, the

sulfide will oxidize and acid will be produced. Acid

drainage often has an orange color typical of the

oxidized iron compounds formed.

Under anaerobic conditions the opposing process of

sulfate reduction may become significant. Certain

bacteria (for example Desulfovibrio) can use sulfate

(SO4

-2) as an alternative electron acceptor when O2 and

N03 have become depleted (table 4). The reduction of

SO4

-2

 produces the sulfide ion (S-2), which may be

released from the soil as hydrogen sulfide gas (H2S) or

combine with Fe to form FeS. The presence of the

"rotten egg" smell of H2S and the black color

associated with FeS are both strong indicators of

anaerobic soil conditions. These are useful indicators of

the hydric soils typical of wetlands (see following

section also).

Figure 3 —Microbial sulfur transformations in the soil

environment.

Fe and Mn oxidation and reduction —The soil redox

chemistry of Fe and Mn are complex with many

reactions capable of occurring without microbial

intervention. However, these same reactions often

occur more quickly when the microbial catalyst is

active (Loveley 1995). Certain chemolithotrophic

bacteria have been implicated in the oxidation of Fe (II)

to Fe (III); an example of a chemolithotrophic aerobic Fe

oxidizing bacteria is Thiobacillus ferroxidans.

Manganese oxidizing bacteria have been implicated in

the formation of desert varnish, the dark coloring on

desert rocks containing the brown MnO2. Conversely,

bacteria are also involved in the reduction of Fe and Mn

under anaerobic conditions (table 4). Both Fe (III) and

Mn (IV) are capable of acting as terminal electron

acceptors in anaerobic respiration (Lovely 1995). The

reduced products Fe (II) and Mn (II) give anaerobic soils

their dark, low chroma or gleyed appearance. These

reduced forms are also more mobile in soils and may be

translocated, forming the redox depletions and

concentrations that are indicators for hydric soils in

wetland delineation.

SUMMARY

Bacteria have a crucial role in ecosystem processes

both as a source and sink of carbon and nutrients and

as catalysts for key ecosystem processes. Their

incredible metabolic diversity allows them to transform

carbon and nutrients under a wide variety of

environmental conditions. In contrasts to plants and

animals, the true extent of the diversity of bacteria in

the soil environment is only beginning to be explored.

The varied roles of bacteria in biogeochemistry are

integral to the understanding and management of

nutrient cycling, wetland geochemistry and C cycling in

terrestrial ecosystems.

65
ACKNOWLEDGMENTS

The author was supported during the preparation of

this manuscript by the Utah Agricultural Experiment

Station, Utah State University, Logan Utah 84322-

4810. Approved as journal paper # 7091.

LITERATURE CITED

Anderson, J.P.E.; Domsch, K.A. 1973. Quantification

of bacterial and fungal contributions to soil

respiration. Archiv for Microbiolgie. 93:113-127.

Anderson, J.P.E.; Domsch, K.A. 1975. Measurements

of bacterial and fungal contribution to respiration to

selected agricultural and forest soils. Canadian

Journal of Microbiology.21:314-322.

Baath, E. 1990. Thymidine incorporation into soil

bacteria. Soil Biology Biochemistry. 22(6):805-810.

Bakken, L. R. Culturable and nonculturaable bacteria in

soil. 1997. In: van Elsas, J.D.; Trevors, J.T.;

Wellington, Elizabeth M.H. (eds.). Modern Soil

Microbiology. Marcel Dekker, Inc. New York.

Bardgett, R.D.; Hobbs, P.J.; Frostegard, A. 1996.

Changes in soil fungal: bacterial biomass ratios

following reductions in the intensity of management

of an upland grassland. Biology and Fertility of Soils.

22:261-264.

Bass Becking, L.G. M.; Kaplan, I.R. Moore, D. 1960.

Limits of the natural environment in terms of pH and

oxidation-reduction potentials. The Journal of

Geology. 68(3):243-285.

Berg, M.L.; Kniese, J.P.; Verhoef, Mar, H.A. 1998.

Dynamics and stratification of bacteria and fungi in

the organic layers of a scots pine forest soil. Biology

and Fertility of Soils. 26:(4)313-322.

Bloem, J.; de Ruiter, P.; Bouwman, L. 1997. Soil food

webs and nutrient cycling in agroecosystems. In:

van Elsas, J.D.; Trevors, J.T.; Wellington, E.M.H.

(eds.). Modern Soil Microbiology. Marcel Dekker,

Inc. New York.

Bloem, J.; Veninga, M.; Shepherd, J. 1995. Fully

automatic determination of soil bacterium numbers,

cell volumes, and frequencies of dividing cells by

confocal laser scanning microscopy and Image

Analysis. Applied and Environmental Microbiology.

61:926-936.

Bohn, H.L.; McNeal, B.L.; O'Connor, G.A. 1985. Soil

Chemistry, 2nd Ed. John Wiley & Sons. New York.

Bottomley, P. J. 1994. Light microscopic methods for

studying soil microorganisms. In: Weaver, R.W.

(chair, Edit. Committee). Methods of Soil Analysis,

Part 2, p. 81-105. Soil Science Society of America,

Madison, Wl.

Clarholm, M.; Rosswall, T. 1980. Biomass and

turnover of bacteria in a forest soil and a peat. Soil

Biology Biochemistry. 12:49-57.

Ehrlich, H.L. 1990. Geomicrobiology, 2nd Ed. Marcel

Dekker, Inc. New York.

Firestone, M.K. 1982. Biological denitrification.

pp.289-318 In: Stevenson, F.J. (ed.). Nitrogen in

Agricultural Soils. American Society of Agronomy,

Madison.

Griffiths, R.P.; Entry, J.A.; Ingham, E.R.; Emmingham,

W.H. 1997. Chemistry and microbial activity of

forest and pasture riparian zone soils along three

pacific northwest streams. Plant and Soil. 190(1):

169-178.

Haack, S.K. 1996. Chemolithotrophic Bacteria. In:

Hall, G.S. (ed.). Methods for the Examination of

Organismal Diversity in Soils and Sediments. CAB

International, Oxford UK.

Hartmann, A.; Abmus, B; Kirchhof, G.; Schloter, M.

1997. Direct approaches for studying soil microbes.

In: van Elsas, J.D.; Trevors, J.T.; Wellington,

E.M.H. (eds.). Modern Soil Microbiology. Marcel

Dekker, Inc. New York.

Jenkinson, D.S. 1977. The soil biomass. New Zealand

Soil News. 25:213-218.

Lovely, D.R. 1995. Microbial reduction of iron,

manganese and other metals.pp. 175-231 In:

Sparks, D.L. (ed.). Advances in Agronomy.

Academic Press, San Diego.

Lundgren, B.; Soderstrom, B. 1983. Bacterial numbers

in a pine forest soil in relation to environmental

factors. Soil Biology Biochemistry. 15 (6):625-630.

Martens, R. 1995. Current methods for measuring

microbial biomass C in soil: potentials and

limitations. Biology and Fertility of Soils. 19:87-99.

Metting, F. Blaine Jr. Structure and physiological

ecology of soil microbial communities. In: Metting,

F. Blaine Jr. (ed.). Soil Microbial Ecology. 1993.

Marcel Dekker, Inc. New York.

Norton, J.M.; Firestone, M.K. 1991. Metabolic status

of bacteria and fungi in the rhizosphere of

66
Ponderosa pine seedlings. Applied Environmental

Microbiology. 57:1161-1167.

Paul, E.A.; Clark, F.E. 1989. Soil Microbiology and

Biochemistry, Academic Press, Inc. San Diego.

Smith, J.L.; Paul, E.A. 1990. The significance of soil

microbial biomass estimations. In: Bollag, JeanMarc; Stotzky, G. (eds.). Soil Biochemistry. Marcel

Dekker, Inc. New York. (6):357-396.

Staben, M.L.; Bezdicek, D.F.; Smith, J.L.; Fauci, M.F.

1997. Assessment of soil quality in conservation

reserve program and wheat fallow soils. Soil

Science Society of America Journal. 61(1):124-130.

Stark, J.M.; Hart, S.C. 1997. High rates of

nitrification and nitrate turnover in undisturbed

coniferous forests. Nature 385:61-64.

Stevenson, F.J. 1986. Cycles of Soil, John Wiley &

Sons. New York.

Torsvik, V.; Goksoyr, J.; Daae, F.L. 1990. High

diversity in DNA of soil bacteria. Applied and

Environmental Microbiology. 56:782-787.

van Veen, J.A.; Ladd, J.N.; Amato, M. 1985.

Turnover of carbon and nitrogen through the

microbial biomass in a sandy loam and a clay soil

incubated with [14C(U)]glucose and [15N](NH4)2SO4

under different moisture regimes. Soil Biology

Biochemistry. 17 (6):747-756.

Velvis, H. 1997. Evaluation of the selective

respiratory inhibition method for measuring the ratio

of fungal: bacterial activity in acid agricultural soils.

Biology and Fertility of Soils. 25 (4):354-360.

Vishnevetsky, S.; Steinberger, Y. 1997. Bacterial and

fungal dynamics and their contribution to microbial

biomass in desert soil. Journal of Arid Environments.

37 (1):83-90.

Vitousek, P., Gosz, J.R., Grier, C.C, Melillo, J.M.;

Reiners, W.A. 1982. A comparative analysis of

potential nitrification and nitrate mobility in forest

ecosystems. Ecological Monographs. 52 (2):155-

177.

White, W.W.; Jeffers, T.H. 1994. Chemical predictive

modeling of acid mine drainage from metallic

sulfide-bearing waste rock In: Alpers, C.N; Blowes,

D.W. (eds.). Environmental Geochemistry of Sulfide

Oxidation. American Chemical Society, Washington

D.C.

67
DECOMPOSITION OF WOODY MATERIAL: NUTRIENT DYNAMICS, INVERTEBRATE/

FUNGI RELATIONSHIPS AND MANAGEMENT IN NORTHWEST FORESTS

Robert L. Edmonds1 and James L. Marra2

ABSTRACT

Coarse woody debris (CWD) is being actively managed in many Pacific Northwest forest ecosystems because of its

ecological importance. This paper discusses: (1) the source, amount and type of CWD present in ecosystems relative

to their fire regime, (2) decomposition rates of CWD, (3) relationships among invertebrates, fungi and fire in the

decomposition process, (4) nutrient dynamics in CWD, and (5) management of CWD. Amounts of CWD have

decreased as a result of forest management west of the Cascade crest, but increased in the drier areas east of the

Cascade crest as a result of fire suppression, defoliator and bark beetle attacks and diseases. Logs decay slowly, but

this depends on their species and size and the environmental conditions. Coarse woody debris provides a considerable

amount of soil organic matter, but lower amounts of nutrients. However, it can be both a source and sink for nitrogen.

Invertebrates and fungi are intimately related in the decomposition process and CWD is a source of ecosystem

biodiversity. Understanding the natural ecology of CWD in various ecosystems is important in determining how much

CWD to leave and recruit. Long-term field experiments and simulation modeling efforts are being conducted to

determine the role of CWD in ecosystem sustainability and in maintaining long-term productivity.

Keywords: logs, snags, coarse woody debris, decomposition, nutrients, invertebrates, fungi

INTRODUCTION

Until relatively recently coarse woody debris (CWD)

was considered to be a nuisance for forest managers,

interfering with plans for planting seedlings and regenerating sites. It was commonly yarded to logging landings, a practice known as YUM (yarding unmerchantable material), where it was usually burned. If it could

not be yarded, it was piled and burned, or broadcast

burned. However, since the 1980s scientists and

forest managers have become increasingly aware of the

need to retain intact CWD in ecosystems because of its

ecological importance. Coarse woody debris is an

important component of the concept of ecosystem

management (Kohm and Franklin 1997). In terrestrial

ecosystems CWD provides habitats for plants, small

mammals, birds, bats, invertebrates, mycorrhizal and

decomposer fungi and bacteria; improves slope stability

and prevents erosion; stores water, carbon and nutrients; is a site for N fixation; helps maintain site productivity; and is an important site for biodiversity (Maser

and Trappe 1984, Harmon and others 1986, Jurgensen

and others 1987, Carpenter and others 1988, Franklin

and others 1989, Harmon 1992, Caza 1993, Griffith

and others 1993, Harmon and others 1994, Harmon

and Sexton 1995, Bull and others 1997). It is also

important in forest streams where it creates pools for

fish and stabilizes stream banks (Maser and Trappe

1984). Some researchers, however, are still skeptical

about the role of CWD, particularly with respect to its

role in maintaining forest productivity. For example,

Johnson and Todd (1998) suggested that dead logs in

southeastern U.S. oak forests appear to be an ephemeral and rather unimportant ecosystem component.

Respiration from logs is also considerable (Marra and

Edmonds 1994), and this contribution to atmospheric

CO2 and global warming has been widely debated

(Harmon and others 1990; Johnson and Todd 1998).

Harmon and Chen (1991) and Harmon and Sexton

(1996) defined CWD as material greater than about 10

cm in diameter at the large end and 1.5 m in length.

This is roughly comparable to the 7.6 cm diameter (the

100-hour fuel size break) used for fuel inventories.

Coarse woody debris is now being actively managed in

many forest ecosystems and sound, fresh logs as well

as decaying logs and snags are left on site after harvesting. Fresh snags are created by cutting tree tops

with chain saws or blowing them off with explosives

and green trees are left as future sources of snags and

eventually logs. These are considered to be biological

legacies for the next forest and most state forest practices acts now include provisions for managing CWD.

In Washington typical state provisions require managers

to leave 5 logs greater than 30 cm in diameter and 12

green trees/ha (Washington Forest Practices Board

1995), but this varies from organization to organization. The Federal Northwest Forest Plan requires 15%

1Robert L. Edmonds is a professor of soil microbiology and forest pathology. University of Washington, Seattle, WA 89195

2

James L. Marra is a research associate in soil microbiology, University of Washington, Seattle, WA 89195

68
retention of the area in green trees where harvesting is

allowed with tree densities ranging from 15-60 trees/ha

(USDA Forest Service and USDI Bureau of Land Management 1994). Guidelines are also given for CWD; for

most of western Oregon and Washington leave 180

linear meters of logs/ha greater than 51 cm in diameter,

but logs less than 6 m in length are not credited towards the total. In eastern Oregon and Washington

and southwestern Oregon a minimum of 90 linear meters of logs per ha greater than or equal to 41 cm in

diameter and 5 m long should be retained. But should

these recommendations for CWD apply to all forest

ecosystems? We would argue no; CWD should be

managed relative to its natural ecology. The Northwest

Forest plan, however, does distinguish between moist

and dry ecosystems.

We still have a lot to learn about the ecology and management of CWD in Pacific Northwest forest ecosystems. Coarse woody debris does not behave the same

in all ecosystems and the functioning of CWD is largely

controlled by the climate, tree species and fire regime

occurring in a given ecosystem. Fire regimes are extremely important in the Pacific Northwest so they will

provide the focus for this discussion. Agee (1993)

delineated three distinct fire regimes defined by temperature and moisture conditions (fig. 1): (1) high intensity, low frequency fires that occur in western hemlock

(Tsuga heterophylla (Raf.) Sarg.) /Douglas-fir (Pseudotsuga menziessi (Mirb.) Franco), Pacific silver fir (Abies

amabilis (Dougl.) Forbes) and subalpine ecosystems, (2)

moderate intensity, frequency fires that occur in dry

Douglas-fir, mixed evergreen, lodgepole pine (Pinus

contorta var. murrayana (Grev. and Balf.) and red fir

(Abies magnifica var. shastensis Lemm.) ecosystems,

and (3) low intensity, high frequency fires that occur in

mixed conifer, ponderosa pine (Pinus ponderosa Dougl.

ex Laws) and oak woodland ecosystems.

The objectives of this paper are to discuss: (1) the

source, amount and type of CWD present in ecosystems relative to their fire regime, (2) decomposition

rates of CWD in representative ecosystems, (3) relationships among invertebrates, fungi and fire in the

decomposition process, (4) nutrient dynamics in CWD,

and (5) management of CWD.

SOURCE, AMOUNT AND TYPE OF CWD

Unlike fine litter, such as needles, leaves and small

branches, inputs of CWD are pulsed and related to the

events listed in table 1, such as fire, wind, insect attacks and management activities, that occur infrequently. Large wind storms such as the 1921 blow

and the 1962 Columbus day storm in coastal

Washington produced huge quantities of CWD. Coarse

Figure 1 —Natural fire regimes in the Pacific Northwest

(adapted from Agee 1993).

woody debris is also unevenly distributed spatially. A

good example of this is provided by Agee and Huff

(1987) who presented detailed maps of the position of

large logs and snags in plots on the Olympic Peninsula

of Washington. Logs may cover a considerable portion

of the forest floor.

Table 1 —Factors influencing inputs of CWD into forest ecosystems.

Fire

Insect attacks (especially defoliators and bark beetles)

Disease

Wind

Interactions of disease and wind, fire and insects and fire,

insects and diseases

Volcanic eruptions

Landslides and erosion

Floods

Avalanches

Competition among trees

Human activity - clearcutting, thinning, etc.

In the western Olympic Peninsula Edmonds and Lebo

(in press) and Graham and Cromack (1982) reported

that the average surface area occupied by dead boles

was 9.8 and 11.6 percent, respectively. Franklin and

Waring (1980) reported an even higher value of 16

percent for the Oregon and Washington Cascades.

There is also a considerable amount of buried wood in

the soil in cool, moist ecosystems. In drier east slope

environments the surface area covered is generally

smaller and buried wood is less common, especially in

ecosystems with high fire frequency. The soil surface

area occupied by logs is important because many tree

species, such as western hemlock and Sitka spruce

69
(Picea sitchensis (Bong.) Carr.) regenerate almost solely

on rotted wood. Rotted wood can also act as refugia

for roots and mycorrhizal fungi, as well as invertebrates

and other animals.

Coarse woody debris is an extremely dynamic component of forest ecosystems. In the Douglas-fir/western

hemlock ecosystems in the Cascades of Oregon and

Washington large high intensity stand replacement fires

naturally kill large numbers of trees and produce large

amounts of CWD as shown in figure 2. This CWD

decomposes, but as the stand develops new inputs

from competition mortality, insects, diseases and wind

tend to increase the biomass of CWD until it is ready to

burn again in the next high intensity fire 500 years

after the initial fire, repeating the cycle.

Figure 2 —Amount of coarse woody debris (CWD) in Douglasfir/western hemlock ecosystems in the Oregon and

Washington Cascade Mountains relative to time after a natural

catastrophic stand replacement fire (adapted from Spies and

others 1988).

Large wind storms such as the 1921 blow and the

1962 Columbus Day storm on the Olympic Peninsula of

Washington also produce large areas of downed trees.

Smaller gaps may develop with less intense storms.

For example, on December 31, 1996 a storm created

many gaps that were 1 ha or smaller in the Hoh rain

forest in Olympic National Park. On closer examination

we found that many of the blowdown trees in these

gaps were suffering from Armillaria root disease. Interestingly, all of the trees were western hemlocks ranging in age from about 40 to 110 years. Larger and

older Douglas-fir and western redcedars in the area

were not blowdown. Wind and root diseases are intimately related in creating CWD in this case.

Table 2 shows the amount of CWD occurring in oldgrowth forests on the Olympic Peninsula with total

amounts of logs and snags ranging from around 100

Mg/ha to more than 500 Mg/ha. Forest management,

with YUM yarding, slash burning and 60-year rotations

has short-circuited this process and as a result amounts

of CWD in most managed west-side stands are now

considerably lower and of smaller diameter than those

in natural stands. Edmonds (1991) estimated that log

biomass was <10 Mg/ha in third rotation Douglas-fir

stands in the Puget Sound region while stumps contributed only 3 Mg/ha.

Table 2 —Biomass (Mg/ha) of logs and snags (coarse woody

debris) in contrasting cool, moist western hemlock/Douglas-fir

ecosystems on the Olympic Peninsula of Washington and hot,

dry mixed conifer ecosystems in the Sierra-Nevada Mountains

of California.

In contrast to the large amount of CWD in the moist

forests, the amount of CWD in the low intensity fire

mixed conifer forests of California is much less (< 100

Mg/ha and more usually near 50 Mg/ha) (table 2).

However, in many dry forest types in the inland west,

the amount of CWD is higher than natural levels, due

to fire suppression, selective species logging, increases

in stand density, bark beetles, and diseases such as

white pine blister rust and Armillaria root disease. This

has given forest managers a problem; should they

remove this additional CWD by prescribed burning or

salvage beetle killed trees to prevent future large fires?

Many environmentalists strongly oppose salvage logging, but in some cases it may be justified. The season

of burning controls the amount of CWD removed.

Prescribed burns in mixed conifer forests in California in

late fall can essentially consume all the CWD, especially rotten logs (table 3). Cooler fires in spring typically do not reduce the amount of CWD to the same

extent and in some cases can actually increase CWD

biomass due to tree mortality; for example, at the Challenge Experimental Forest (table 3). In some cases

CWD may need to be protected during burning if there

is not enough on the site.

DECOMPOSITION RATES OF CWD IN

DIFFERENT ECOSYSTEMS

Decomposition rates of CWD vary according to species,

diameter and stage of decay. Substrate chemistry is a

major factor controlling decomposition. Extractives, or

non-cell wall chemicals that convey decay resistance

70

Western hem

Logs

442a/

94-121c/

lock/Douglas-fir

Snags

95

26-40

Mixed conifer

Logs Snags

20.2 - 43.3b/ -

28.0-48.74d/ 0.7-52

aAgee and Huff (1987)

bKauffman and Martin (1989)

cGraham and Cromack (1982)

dHarmon and others (1986)
strongly control wood decomposition rates. Species

with colored heartwood, such as Douglas-fir and western redcedar (Thuja plicata D. Don) are generally more

decay resistant than white wood species such as western hemlock and true firs. Environmental factors such

as temperature and moisture also control decomposition rates.

Table 3 —Percent consumption by prescribed fire of greater

than 7.6 cm diameter logs in Sierra Nevada mixed conifer

forests in California (from Kauffman and Martin 1989).

Figure 3 shows the typical course of mass loss of

Douglas-fir logs in old-growth forests in the Cascades

of Washington and Oregon in relation to time. Four

phases are involved. Phase 1 is a lag period before

mass loss can be detected lasting 5 or so years. Following the lag period there is a period of fairly rapid

decay involving fragmentation, leaching and microbial

mineralization (Phase 2). Phase 3 is slightly slower and

involves cellulose loss. The final stable phase involves

very slow decomposition and lignin degradation. This

is typical for brown rotted logs. The difference between brown and white rots is discussed in the next

section.

Log decomposition is typically expressed as a series of

negative exponential curves. A 3-exponential model

fits the data in figure 3 representing phases 2, 3 and 4.

Schowalter and others (1998) suggested that a 3-exponential model might also be suitable for oak log decomposition. The first phase represents an initial rapid loss

from bark substrates followed by slower sapwood

decay (phase 2). Phase 3 is the very slow heartwood

decay. Logs, however, do not always decompose from

the outside in. Insects are important in the initial

stages of decay and wood borers may introduce decay

fungi into logs allowing them to decompose rapidly

from the inside (Edmonds and Eglitis 1989).

Table 4 shows CWD decomposition rates, expressed as

half-lives, for both snags and logs of different tree

species. Half-lives, or time to 50% exponential mass

Figure 3 —Phases of decomposition of old-growth Douglas-fir

logs in the Oregon and Washington Cascades. Data from

Sollins and others (1987)..

loss, for both mineralization and fragmentation are

shown. Snags tend to decompose faster than logs,

although some large Douglas-fir snags may slowly

mineralize and may have half-lives as long as 230

years. Half-lives for snags vary from 5-55 years for

fragmentation and 18-230 years for mineralization.

Some standing dead trees simply fall after a period of

time. In lodgepole pine ecosystems in Oregon, beetle

killed trees mostly fell in the first 7 years (Mitchell and

Preisler 1998).

Table 4—Decomposition rates (half-lives) of snags and logs of

different tree species (from Harmon and others 1 986)

Half-lives for logs vary from 36-150 years for fragmentation and 14-172 years for mineralization. Note that

both Douglas-fir snags and logs decompose less rapidly

than western hemlock snags and logs (table 4). Logs

of white fir (Abies concolor (Gord & Glend.) Lindl. ex

Hildebr.), a white wood species, and even lodgepole

pine (Busse 1994), decompose very rapidly despite the

dry environments in which they occur. Insects play an

71

Season

Early fall

Early spring

Location

Blodgett Experimental Forest

Challenge Experimental Forest

Quincy Experimental Forest

Average

Blodgett Experimental Forest

Challenge Experimental Forest

Quincy Experimental Forest

Average

Log type

Sound

40.3

78.6

83.7

67.5

37.3

+ 35.6

48.0

16.6

Rotten

96.1

100.0

100.0

8.7

33.4

+ 35.2

81.5

26.6

Species

Ponderosa pine

lodgepole pine

Douglas-fir

western hemlock

Ponderosa pine

lodgepole pine

white fir

Douglas-fir

western hemlock

Half-lives (years)

Snags

Fragmentation

5-18

4-37

6-55

12

mineralization

-

-

25-230

18-43

Logs

-

-

36

166

-

-

57

14

98-172

29-88
important role in the decomposition of white fir logs.

In the dry mixed-conifer forests in Sequoia National

Park in California, wood volume excavated by insects

increased with the age of white fir logs (Harmon and

others 1987). Logs less than 6-years-old were rarely

attacked by carpenter ants (Camponotus spp.) or termites (Zootermopsis nevadensis), but in older logs

insects excavated wood volume at a rate of about 0.7

percent per year. Relationships between insects, fungi

and fire are discussed further in the next section.

Unfortunately, we could find no data for decomposition

rates of large ponderosa pine logs. Erickson and others

(1985), however, examined the decomposition rates of

small 8-12 cm diameter sections of ponderosa pine in

eastern Washington and found them to have an average half life of 53 years on the ground and 58 years

when suspended above ground (table 5). Decomposition rates of the ponderosa pine sections were considerably slower than those for the Douglas-fir and western hemlock sections in western Washington. Table 5

also indicates that decomposition rates for the Douglasfir, western hemlock and ponderosa pine sections were

generally greater when the wood was in contact with

the ground surface. Logs do not decompose uniformly

and sections suspended above ground usually decompose slower than those in contact with the forest floor

and soil because of more favorable moisture conditions.

Interestingly, the slowest decomposition rates occurred

in Pacific silver fir sections which had a half-life of 78

years whether suspended or not. In cold, moist, high

elevation Pacific silver fir ecosystems, deep, long-lasting snow packs keep moisture conditions above ground

similar to those on the soil surface for most of the year.

Table 5 —Half-lives of 8-12 cm diameter wood sections situated on or suspended above the soil surface in different ecosystems in Washington (from Erickson and others 1985).

Species of wood

and location

Pacific silver fir

Findley Lake

Ponderosa pine

Eastern Washington

Douglas-fir

Pack Forest

Western hemlock

Olympic Peninsula

Position

On soil

Suspended

On soil

Suspended

On soil

Suspended

On soil

Suspended

Half-lives

(years)

77

77

58

53

19

43

19

29

RELATIONSHIPS BETWEEN INVERTEBRATES, FUNGI AND FIRE

Invertebrates, fungi and fire typically interact in the

decomposition of CWD, although in cool, moist ecosystems fire plays a smaller role because of the long time

periods between fires and the small average annual

area burned. In these areas invertebrates and fungi

dominate. In dry ecosystems fire and insects may

dominate the decomposition process. Table 6 shows

fire data for different forest types in Oregon, including

areas in each type, fire severity, fire periodicity, average annual area burned and the percentage of area

burned annually. As much as 6.7 percent of the ponderosa type burned historically every year, while less

than 1 percent of the Douglas-fir, cedar/spruce/hemlock

and subalpine types burned.

Table 6 —Fire periodicity, severity, area, and annual historic area and percentage of area typically burned

in Oregon (data courtesy of James Agee, College of Forest Resources, University of Washington).

Forest type

Ponderosa pine

Oak woodland

Mixed conifer

Lodgepole pine

Douglas-fir

Cedar/spruce/

hemlock

Subalpine

Other

Fire periodicity

(years)

15

25

30

80

150

400

800

-

Severity

Low

Low

Low

Moderate

Moderate-High

High

High

--

Area in type

(thousands

of ha)

3,142

1,002

399

757

4,446

292

1,076

2,398

Average annual

historic area burned

(thousands of ha)

209.4

40.1

13.3

9.4

29.6

0.7

1.3

18.0

Percentage of

area burned

6.7

4.0

3.3

1.2

0.7

0.3

0.1

0.8

72
An excellent example of interactions between fire,

insects and fungi is provided in the moderate fire regime lodgepole pine forests in central Oregon (Gara and

others 1985). In their study area they determined that

a new lodgepole pine forest was established after a

stand replacement fire in 1839. As the forest developed competition for light and moisture caused considerable suppression mortality. Over the next 60 years

these small dead trees provided fuels that lay on top of

the more decayed logs remaining from the 1839 fire.

In 1989 a slow-moving-smoldering fire slowly burned

the logs away to ash, killing trees and wounding major

lateral roots and boles of residual trees where they

contacted the burning logs. Ips spp. beetles attacked

the killed trees and a succession of fungi entered the

wounded trees, especially Poria asiatica. In 1970, the

largest trees with advanced butt rot were preferentially

attacked by the mountain pine beetle (Dendroctonus

ponderosae). In subsequent years the largest of the

remaining trees were attacked setting up the forest for

another stand replacement fire 1 50 years after the

initial fire.

Bark beetles and wood borers are among the earliest

colonizers of wood (Edmonds and Eglitis 1987; Harmon

and others 1987; Carpenter and others 1988).

Edmonds and Eglitis (1987) found that the Douglas-fir

beetle quickly attacked freshly downed Douglas-fir logs

in the Cedar River Watershed 60 km southeast of

Seattle, but had little influence on log decomposition.

On the other hand, the wood borer Monochamus

scutellatus introduced fungi deep into wood speeding

up the decomposition process. The two most common

fungi observed were Oligoporus placentus and

Fomitopsis pinicola, both brown rotters. Logs that

were screened to prevent insect attack had much less

decomposition. Log diameter was an important variable; small diameter logs had less insect attack, perhaps because they could not establish in the bark or

phloem. As a result large diameter logs had a faster

decomposition rate than small diameter logs, which is

counter to conventional wisdom.

Wood borers and bark beetles are not the only invertebrates associated with logs, and as decomposition

proceeds a wide variety of invertebrates are found.

Mites, Collembola, beetles and fly larvae commonly

occurred in 10-30 cm diameter decay class 3 Douglasfir and western hemlock logs in 50-70 year-old stands

on the Olympic Peninsula (table 7). Logs are typically

classified in 5 decay class from 1 (least decayed) to 5

(most decayed) (Maser and others 1979). Decay class

3 logs are the most common logs on the forest floor

(Harmon and others 1986), and this is clearly illustrated

in figure 4 for an old-growth forest in the Hoh River

Valley in Washington for both log biomass and numbers

per ha. Populations of mites and Collembola were

much higher than those for beetles and fly larvae in

both brown and white rotted wood (table 7). However,

brown rotted wood had significantly higher populations

of both mites and beetles than white rotted wood.

Collembola were also higher, but not significantly.

Brown rot is produced by fungi that decay only the

cellulose, leaving a lignin residue that remains for hundreds of years contributing to soil organic matter. On

the other hand, white rotted wood is produced by fungi

that decompose both cellulose and lignin and thus has

a much shorter residence time. White rotted wood

may less favorable for invertebrates because it is

thought to have a higher moisture content than brown

rotted wood, although this is not always the case

(table 8).

Table 7 —Invertebrate density in brown and white rotted wood

in 10-30 cm diameter decay Class 2-3 Douglas-fir/western

hemlock logs on the Olympic Peninsula, Washington.

Invertebrate

Mites

Collembola

Beetles

Fly larvae

Density (No.

Brown rot

177,231

80,801

2,141

1,959

m

-3)

White rot

1 1 2,71 6a

54,1 17

889a

3,854

significantly different p<0.05.

Figure 4 —Log biomass and number of western hemlock and

Sitka spruce logs per ha in various decay classes in an oldgrowth forest in the Hoh River Valley, Olympic National Park

(from Edmonds and Lebo in press).

Decaying logs provide a considerable amount of

biodiversity in forest ecosystems and to date we have

found an average of 142 species of invertebrates in

old-growth logs on the Olympic Peninsula and 112

species in logs in clearcuts. Interestingly, we have

found a species of beetle in the Family Throscidae in

73
logs young-growth stands that did not occur in logs in

old-growth stands (Marra 1995). In general beetles

appear to respond to changes in specific characteristics

of CWD such as log species, diameter and species of

fungal sporocarps more than other invertebrate groups,

suggesting that they may be good indicators of ecosystem change. Carpenter ants and termites are also

heavily involved in the wood decomposition process,

but are more important in drier environments.

Table 8 —Percentage of 10-30 cm diameter decay class 2-3

Douglas-fir/western hemlock logs on the Olympic Peninsula,

Washington with brown and white rot and related moisture

contents in spring 1995 and 1996.

Sampling

Season

Spring 95

Spring 96

Spring 95

Spring 96

Percentage

Brown

52

26

Brown

169

168

White

32

42

of decay type

Both

9

14

Percent moisture

White

157

206

Both

153

260

No decay

17

18

No decay

102

137

Carpenter and others (1986) examined the early pattern

of heterotroph activity in four species of logs in a 200

year log decomposition study established in 1985 in

the H.J. Andrews forest in the Oregon Cascades and

described in Harmon (1992). The foodweb during the

early stages of decomposition includes nitrogen-fixing

and other bacteria, fungi, protozoa, nematodes and

arthropods. Fungi play an extremely important role in

the wood decomposition process. Harmon and others

(1994) found 13 species of fungi fruiting on class 1

logs in the first 7 years with maximum production of

fruiting bodies at age 5. None of the species were

mycorrhizal fungi. Edmonds and Lebo (in press) noted

considerably more taxa (82, including 18 mycorrhizal

species) in decayed logs on the Olympic Peninsula

(table 9). Some species found by Harmon and others

(1994) were not observed by Edmonds and Lebo (in

press). Species common to both sites included

Naematoloma capnoides and Fomitopsis pinicola.

Naematoloma capnoides was the most common species on class 1 logs in Oregon. Logs are a considerable

source of ecosystem biodiversity for both invertebrates

and fungi in forest ecosystems.

Large diameter CWD in old-growth forests west of the

Cascade crest appears to be largely decayed by brown

rot fungi, particularly by Fomitopsis pinicola. Is the

smaller diameter CWD associated with managed stands

decayed mostly by brown or white rot fungi? We

hypothesized that white rot fungi might predominate

because of the higher sapwood component of small

diameter logs. However, table 8 shows no consistent

trend for small diameter Douglas-fir and western hemlock logs on the Olympic Peninsula. Samples of 30

decay class 2-3 logs revealed a predominance of brown

rot in Spring 1995, but the opposite trend in Spring

1996. More data need to be obtained before we can

come to a conclusion on this subject.

Table 9 —Numbers of fungal taxa occurring on logs in an oldgrowth western hemlock/Sitka spruce forest in the Hoh River

valley, Washington (from Edmonds and Lebo in press).

Fire, invertebrates and fungi all operate in the log decomposition process. However, we still need to learn

more about their roles and relationships, especially in

dry environments. We know very little about fungi and

invertebrates associated with charred wood and the

decomposition rate of charred wood. Hadfield and

Magelssen (1998), however, recently examined

changes in trees killed by the 1994 fires in eastern

Washington and found considerable stain, decay and

insect development in many trees after only 3 years,

depending on the tree species. The pouch fungus

(Cryptoporus volvatus) was most commonly observed.

NUTRIENT DYNAMICS IN CWD

Coarse woody debris is regarded as a source of organic

matter and nutrients for forest soils leading to the

argument that loss of CWD debris from ecosystems

will result in a loss of productivity. Johnson and Todd

(1998) found that removing oak CWD on a harvested

site in Tennessee had no effect on the C contents of

either vegetation or soils after 15 years harvesting. The

connection between CWD and forest productivity,

however, still needs more research. Coarse woody

debris contributes a large amount of organic matter to

the soil in moist ecosystems in the Pacific Northwest,

but only a small amount of nutrients. In the Hoh River

Valley of Washington, the biomass of logs in all decay

classes totals 205,000 kg/ha or 205 Mg/ha (table 10).

The nutrient content of these logs, however, is relatively small (180 kg/ha for N, 20 kg/ha for P and 64

kg/ha for K). Fine litter contributes considerably more

nutrients to soil than CWD (Keenan and others 1993).

74

Type of fungus

Decomposers/pathogens

Mycorrhizal

Total

Number of taxa

64

18

82
Because of the high C/N ratio of logs (> 500:1) it has

been suggested that logs are sinks for N during the

decomposition process and Grier (1978) found strong

immobilization of N in western hemlock logs in coastal

Oregon during the first 24 years of decomposition.

Nitrogen contents of logs can increase as a result of

oatmospheric inputs, fungal hyphae transporting N

from the surrounding litter, and internal N fixation by

bacteria. Decomposing logs could prevent nitrate

leaching losses from ecosystems by immobilizing N.

Nitrogen losses from logs occur through fire, leaching,

invertebrates and the production of fungal fruiting

bodies. Even during the early stages of decomposition

some N might be released from logs through leaching

and the production of fungal fruiting bodies. Harmon

and others (1994) found that from 0.9-2.9 % of the

initial N, 1.9-6.6% of initial P and 1.8-4.5% of the

initial K was released from logs in the first 7 years of

decomposition in the Oregon Cascades. However, as

decomposition proceeds, the amount of N, P and K

released in fungal fruiting bodies seems to get smaller

and in decay class 3 logs Edmonds and Lebo (in press)

found that only 0.009 percent of the log N capital,

0.023 percent of the log P capital and 0.029 percent of

the log K content was released annually. Both these

studies illustrate that logs not only function as sinks for

N but also as sources.

Table 10 —Log biomass and N, P and K content, annual

sporocarp production and percentage of log nutrient capital in

an old-growth western hemlock/Sitka spruce forest in Hoh

River valley, Washington (from Edmonds and Lebo in press).

MANAGEMENT OF CWD

There is strong evidence that CWD plays important

structural and functional roles in many forest ecosystems. However, the exact connection between the

amount of CWD, expressed either as numbers of logs

or biomass per ha, and forest productivity is not

known. The diameter of CWD, contact with the soil

surface, decay class, species, orientation on slopes,

climatic regime and whether the CWD is located in

upland or riparian forests or in a high, moderate or low

fire severity regime are variables that have to be taken

into account.

Very general guidelines have been given to forest managers with respect to leaving logs and snags and green

trees from which new snags and logs can be recruited.

How can we better refine these guidelines and make

them more specific for each forest type? Several approaches have been taken; long-term experiments and

modeling. There are a number of examples of longterm experiments including: (1) Mark Harmon's 200

year study at the H.J. Andrews Forest in Oregon

(Harmon 1992), (2) the Long-Term Ecosystem Productivity (LTEP) Study of the U.S. Forest Service, and (3)

the Habitat Development Project of the U.S. Forest

Service on the Olympic Peninsula. The 200-year study

involves 4 species (Douglas-fir, western hemlock, white

fir and western redcedar) and 530 logs 5 m in length

and 50 cm in diameter placed in 6 replicate locations

on the H.J. Andrews. The LTEP Study is located at 6

sites; Sappho, on Washington State Department of

Resources land and on the following national forests;

Willamette, Umatilla, Siskiyou and Siuslaw in Oregon

and Wenatchee in Washington. The study involves the

following treatments; untreated controls in 50-100

year-old forests, 3 levels of CWD (low, moderate, and

high) and 3 successional stages (early - clearcut with

Douglas-fir and red alder, mid - clearcut with Douglasfir, and late - thinned to promote late successional

species). Treatment blocks are 6 ha with 3 replicates

of each treatment. More details on the LTEP study are

available on the following world wide web site -

http://www.cof.orst.edu/research/ltep.

The Habitat Development Project is being conducted in

Adaptive Management areas on Olympic National Forest as part of the Northwest Forest Plan. One objective of this study is to determine if piling small diameter

CWD into structures resembling large logs provides

better habitat for small mammals in thinned 50-70

year-old stands than scattered individual logs. As a

component of this study we have been examining

invertebrates and fungi and the proportion of brown

and white rot in small diameter logs (tables 8 and 9). If

there is a very high proportion of white rot then the

logs will not last long and they will not function like

large logs which tend to last longer because they are

mostly brown rotted. One solution to this potential

problem may be to inoculate fresh logs with brown rot

fungi.

There are a number of computer models incorporating

CWD that can be used to simulate removal or addition

of CWD on forest productivity. An example is FORECAST (FORestry and Environmental Change Simulator)

which is an ecosystem-level model designed to analyze

the consequences of forest management on ecosystem

function, sustainability and harvestable productivity

(Kimmins 1997). Chapter 17 in Kimmins (1997) provides an analysis of models and their role in ecology

and resource management.

75

Logs (kg/ha)

Annual Sporocarp

production (g/ha)

Percentage of log

nutrient capital

Biomass

205,049

6.3

0.0002

N

180

0.21

0.009

P

20

0.06

0.023

K

64

0.24

0.029
While we wait for the results of long-term experiments

and simulation runs to be completed, the best advice

we can give is: 1. Inventory the CWD you have on a

site and see if it falls within the natural range for that

particular ecosystem. If not, then a decision needs to

be made to add new logs or snags or even reduce

CWD. Coarse woody debris levels should be related to

habitat type and east- and west-side ecosystems

should be managed differently. In moist ecosystems, it

should be recognized that a large number of snags and

logs would have to be added to many ecosystems to

reach historical levels. If there were typically 50 or

more logs per ha then adding 5 would make little difference to soil organic matter, but they might provide

habitat for small mammals and salamanders. In drier

ecosystems burning might be conducted to reduce

CWD to historical levels. Salvage logging should be

considered carefully with respect to potentials for future beetle epidemics and large wildfires. Maintaining a

supply of suitable snags may be more difficult to accomplish than logs. Experiments are currently being

conducted where a variety of fungi have been inoculated into green trees to prepare them for cavity nesting birds and other wildlife (Bull and others 1997).

2. Management objectives are very important to take

into account. Is the site to be managed primarily for

timber production or for other values, such as wildlife

habitat and how does the site fit within the objectives

for landscape? Understanding the ecology of CWD in a

particular ecosystem is very important.

CONCLUSIONS

1. CWD plays an important role in the functioning of

ecosystems. Its functional role in stream ecosystems has been well established and many stream

restoration projects are underway. Its role in terrestrial ecosystems is still not completely understood. The sources, amounts, decomposition rates

and perhaps the roles of CWD vary from one ecosystem to another. The amount of CWD varies by

an order of magnitude from the cooler, wetter

ecosystems to the hot, dry ecosystems in the Pacific Northwest.

2. Fire, fungi and invertebrates are all heavily involved

in the creation and decomposition of CWD. Wind

and fungi commonly function together to create

CWD.

3. Amounts of CWD have decreased as a result of

forest management in the cooler moister forests of

the west slopes of the Cascades, Olympics and

Coast ranges. However, it has increased in the

drier, hotter east slopes of the Cascades as a result

of fire suppression and bark beetle attacks.

4. Inputs of CWD are pulsed, which makes its management difficult. We still don't know exactly how

much CWD to keep and add in terrestrial ecosystems and how to arrange it. Interesting questions

still need to be answered, such as, what species is

best to leave, should small logs be in bundles or

left singly and how much brown rot versus white

rot do we need? Do small log lengths (1.5 m)

functions as well as long lengths (>6 m)?

5. Coarse woody debris is both a source and sink of

carbon and nutrients in ecosystems even in the

early stages of decomposition.

6. Coarse woody debris is a source of biodiversity in

ecosystems, especially fungi and invertebrates.

7. We have a better understanding of the role and

functioning of CWD in moist ecosystems than dry

ecosystems. More research needs to be done in

dry ecosystems.

8. We still don't know the exact relationship between

CWD and the sustainability and long-term productivity of Pacific Northwest forests. The long-term

log decomposition study at the H.J. Andrews and

the LTEP study of the U.S. Forest Service are addressing this study, but results will be a long time

in coming. The best short-term solution for examining management strategies is to be conservative

and use simulation models which include CWD as a

component.

QUESTIONS FROM THE FLOOR

Q: Could you speculate on why invertebrate numbers are

higher in brown rotted logs than in white rotted logs?

A: We think that the moisture environment in brown

rotted wood may be more favorable than that in white

rotted wood. Typically white rotted wood is stringy and

wetter than brown rotted wood. During certain stages of

the decay process you can actually wring water out of

white rotted wood. If the environment gets too low in

oxygen then invertebrates would probably leave. There

are probably other reasons to do with the food chain in

decaying wood and the food source for invertebrates.

Q: If fire is the main decomposition agent in dry systems

what difference does it make to have a fall burn? Isn't that

when fires naturally happened?

A: Yes, most fires in dry ecosystems in the west naturally

occur in the fall. What we were trying to point out was

that you get more complete combustion of CWD in the

fall than the spring. So if you wanted to reduce the fire

76
danger from fuel accumulation, but not reduce the CWD

completely, then spring would be time to burn, even

though this may not have happened naturally.

Q: I don't understand how logs tied together replicate a

natural CWD system? Can you explain?

A: The rationale behind this is that a large proportion of

the CWD in old-growth forests is of large diameter. We

obviously can't grow trees of this size in managed forests,

so why not tie smaller diameter logs together in a bundle

and see if they function like large logs. Small diameter

logs are readily available in managed stands. The emphasis of the studies in the Forest Service Adaptive Management Areas in Olympic National Forest is to see if wildlife, particularly small mammals and salamanders, find

bundled logs to be suitable habitat.We are looking to see

if the decay in these small logs, i.e., the proportion of

brown to white rot, is similar to large logs. If the decay is

predominantly white rot then they will not last long and

won't provide suitable habitat for small mammals and

probably invertebrates.

LITERATURE CITED

Agee, James K. 1993. Fire ecology of Pacific Northwest Forests. Washington, DC: Island Press.

493 p.

Agee, James K.; Huff, Mark H. 1987. Fuel succession

in a western hemlock/Douglas-fir forest. Canadian

Journal of Forest Research 17:697-704.

Bull, Evelyn L; Parks, Catherine G; Torgeson, Torolof

R. 1997. Trees and logs important to wildlife in

the interior Columbia River basin. Gen. Tech. Rep.

PNW-GTR-391. Portland, OR: U.S. Department of

Agriculture, Forest Service, Pacific Northwest Research Station. 55 p.

Busse, M.D. 1994. Downed bole-wood decomposition

of lodgepole pine forests of central Oregon. Soil

Science Society of America Journal 58:221-227.

Carpenter, Steven E.; Harmon, Mark E.; Ingham, Elaine

R.; Kelsey, Rick G.; Lattin John D.; Schowalter,

Timothy D. 1988. Early patterns of heterotroph

activity in conifer logs. Proceedings of the Royal

Society of Edinburgh 94B:33-43.

Caza, C.L. 1993. Woody debris in the forests of British Columbia: a review of the literature and current

research. Victoria, British Columbia, Research

Branch, Ministry of Forestry. 99 p.

Edmonds, Robert L. 1991. Organic matter decomposition in western United States forests. In: Harvey,

Alan E.; Neuenschwander, Leon F. (compilers).

Management and productivity of western montane

soils. Gen. Tech. Rep. INT-280, Ogden, UT: U.S.

Department of Agriculture, Forest Service,

Intermountain Research Station: 118-128.

Edmonds, Robert L; Eglitis, Andris. 1989. The role of

the Douglas-fir beetle and wood borers in the decomposition of and nutrient release from Douglasfir logs. Canadian Journal of Forest Research 19:

853-859.

Edmonds, Robert L; Lebo, David S. [in press]. Diversity, production and nutrient dynamics of fungal

sporocarps on logs in an old-growth temperate rain

forest, Olympic National Park, Washington. Canadian Journal of Forest Research.

Erickson, Heather E.; Edmonds R.L.; Peterson, C.E.

1985. Decomposition of logging residues in

Douglas-fir, western hemlock, Pacific silver fir, and

ponderosa pine ecosystems. Canadian Journal of

Forest Research 15:914-921.

Franklin, J.F.; Waring, R.H. 1980. Distinctive features

of the northwestern coniferous forest: development, structure and function. In: Waring, Richard

H. (ed.). Forests: fresh perspectives from ecosystem analysis. Proceedings of the 40th annual biology colloquium; 1979. Corvallis, OR: Oregon State

University Press: 59-86.

Franklin, J.F.; Perry, D.A.; Schowalter, T.D.; Harmon,

M.E.; McKee, A.; Spies, T.A. 1989. Importance

of ecological diversity in maintaining long-term site

productivity. In: Perry, D.A.; Meurisse, R.; Thomas,

B.; Miller, R.; Boyle, J.; Means, J.; Perry, C.R.;

Powers, R.F. (eds.). Maintaining the long-term

productivity of Pacific Northwest forest ecosystems. Portland, OR: Timber Press: 82-97.

Gara, R.I.; Littke, W.R.; Agee, J.K.; Geiszler, D.R.;

Stuart, J.D.; Driver, C.H. 1985. Influence of fires,

fungi and mountain pine beetles on development of

a lodgepole pine forest in south-central Oregon. In:

Baumgartner, David M; Krebill, Richard G.; Arnott;

James, T.; Weetman, Gordon, F. (eds.). Lodgepole

pine, the species and its management: Symposium

proceedings. Pullman: Cooperative Extension,

Washington State University: 153-162.

Graham, Robin Lambert; Cromack, Kermit, Jr. 1982.

Mass, nutrient content, and decay rate of dead

boles in rain forests of Olympic National Park.

Canadian Journal ofForest Research 12:511-521.

77
Grier, Charles C. 1978. A Tsuga heterophylla-Picea

sitchensis ecosystem of coastal Oregon: decomposition and nutrient balances of fallen logs. Canadian Journal of Forest Research 8:198-206.

Griffiths, Robert P.; Harmon, Mark. E; Caldwell, Bruce

A.; Carpenter, Steve E. 1993. Acetylene reduction in conifer logs during early stages of decomposition. Plant and Soil 148:53-61.

Hadfield, James S.; Magelssen, Roy W. 1998. Wood

changes in fire-killed eastern Washington tree

species-year three progress report. Wenatchee,

WA: U. S. Department of Agriculture, Forest Service, Pacific Northwest Region, Wenatchee Field

Office, Wenatchee National Forest. 37 p.

Harmon, Mark E. 1992. Long-term experiments on log

decomposition at the H.J. Andrews Experimental

Forest. Gen. Tech. Rep. PNW-GTR-280. Portland,

OR: U.S. Department of Agriculture, Forest Service, Pacific Northwest Research Station. 28 p.

Harmon, Mark E.; Chen, Hua. 1991. Coarse woody

debris dynamics in two old-growth ecosystems:

comparing a deciduous forest in China and a conifer forest in Oregon. Bioscience 41:604-610.

Harmon, M.E.; Franklin, J.F.; Swanson, F.J.; Sollins,

P.; Gregory, S.V.; Lattin, J.D.; Anderson, N.H.;

Cline, S.P.; Aumen, N.G.; Sedell, J.R.;

Lienkaemper, G.W.; Cromack, K. Jr.; Cummins,

K.W. 1986. Ecology of coarse woody debris in

temperate ecosystems. Advances in Ecological

Research 15:133-302.

Harmon, Mark E.; Sexton, Jay; Caldwell, Bruce A.;

Carpenter, Steve E. 1994. Fungal sporocarp-mediated losses of Ca, Fe, K, Mg, Mn, N, P, and Zn

from conifer logs in the early stages of decomposition. Canadian Journal of Forest Research 24:

1883-1893.

Harmon, Mark E; Sexton, Jay. 1996. Guidelines for

measurements of woody in forest ecosystems.

Seattle, WA: Publication No. 20. U.S. LTER Network Office: University of Washington. 73 p.

Harmon, Mark E.; Sexton, Jay. 1995. Water balance

of conifer logs in early stages of decomposition.

Plant and Soil 172:141-152.

Harmon, Mark E.; Kermit Cromack, Jr.; Bradley G.

Smith. 1987. Coarse woody debris in mixed-conifer forests, Sequoia National Park, California. Canadian Journal of Forest Research 17:1265-1272.

Harmon, Mark E.; Ferrell, William K.; Franklin, Jerry F.

1990. Effects on carbon storage of conversion of

old-growth forests to young forests. Science 247-

699-702.

Johnson, Dale W.; Todd, Donald E., Jr. 1998. Effects

of harvesting intensity on forest productivity and

soil carbon storage in a mixed oak forest. P. 351-

363. In Lal, R.; Kimble, J.M.; Follett, R.F.; Stewart,

B.A. Management of carbon sequestration in soil.

CRC Press, Boca Raton and New York. 457 p.

Jurgensen, M.F.; Larsen, M.J.; Graham R.T.; Harvey,

A.E. 1987. Nitrogen fixation in woody residue of

northern Rocky Mountain conifer forests. Canadian

Journal of Forest Research 17:1283-1288.

Kauffman, J. Boone; Martin. R.E. 1989. Fire behavior,

fuel consumption, and forest-floor changes following prescribed understory fires in Sierrra Nevada

mixed conifer forests. Canadian Journal of Forest

Research 19:455-462.

Keenan, Rodney J.; Prescott, Cindy E.; Kimmins, J.P.

(Hamish). 1993. Mass and nutrient content of

woody debris and forest floor in western red cedar

and western hemlock forests on northern Vancouver Island. Canadian Journal of Forest Research

23:1052-1059.

Kohm, Kathryn A.; Franklin, Jerry F. (eds.). 1997.

Creating a forestry for the 21st century. Washington, DC: Island Press. 475 p.

Kimmins, J.P. 1997. Forest ecology. Upper Saddle

River, NJ: Prentice Hall. 596 p.

Marra, James L. 1995. Coarse woody debris respiration and invertebrate diversity on old-growth and

clear-cut sites on the Olympic Peninsula. Seattle,

WA: University of Washington. Ph. D. dissertation. 149 p.

Marra, James L.; Edmonds, Robert L. 1994. Coarse

woody debris and soil respiration in an old-growth

coniferous forest on the Olympic Peninsula, Washington, USA. Canadian Journal of Forest Research

24:1811-1817.

Maser, Chris; Anderson, Ralph G.; Cromack, Kermit,

Jr.; Williams, Jerry T.; Martin, Robert E. 1979.

Dead and down woody material. In: Thomas,

Jack Ward, (tech. ed.). Wildlife habitat in managed forests: the Blue Mountains of Oregon and

Washington. U.S. Department of Agriculture

Handb. 553: 78-95.

Maser, Chris; Trappe, James M. 1984. The seen and

unseen world of the fallen tree. Gen. Tech. Rep.

PNW-164. Portland, OR: U.S. Department of Agri78
culture, Forest Service, Pacific Northwest Research

Station. 56 p.

Mitchell, Russel G.; Preisler, Haiganoush K. 1998. Fall

rate of lodgepole pine killed by the mountain pine

beetle in central Oregon. Western Journal of Applied Forestry 13(1):23-26.

Schowalter, T.D.; Zhang, Y.L.; Sabib, T.E. 1998.

Decomposition and nutrient dynamics of oak

Quercus spp. logs after five years of decomposition. Ecography 21:3-10.

Sollins, Phillip; Cline, Steven P.; Verhoeven, Thomas;

Sachs, Donald; Spycher, Gody. 1987. Patterns of

log decay in old-growth Douglas-fir forests. Canadian Journal of Forest Research 17:1585-1595.

Spies, Thomas A.; Franklin; J.F.; Thomas, Ted B.

1988. Coarse woody debris in Douglas-fir forests

of western Oregon and Washington. Ecology 69:

1689-1702.

USDA Forest Service; USDI Bureau of Land Management. 1994. Record of Decision for Amendments

to Forest Service and Bureau of Land Management

Planning Documents Within the Range of the

Northern Spotted Owl: Standards and Guidelines

for Management of Habitat for Late-Successional

and Old-Growth Related Species within the Range

of the Northern Spotted Owl. U.S. Government

Printing Office.

Washington Forest Practices Board. 1995. Washington forest practices - Rules WAC 222. Olympia,

WA: Department of Natural Resources.

79
INVERTEBRATE BIODIVERSITY, FORESTRY AND EMULATION OF NATURAL

DISTURBANCE: A DOWN-TO-EARTH PERSPECTIVE

John R. Spence1, Christopher M. Buddie2, Kamal J. K. Gandhi3, David W. Langor4, W. Jan A. Volney5, H. E. James

Hammond6 and Greg P. Pohl7

ABSTRACT

A fire-based conceptual model of forest spatial dynamics is widely embraced as the way to make boreal forest

management more ecologically friendly and to conserve biodiversity. However, use of the term "ecosystem based

management" does not automatically confer either wisdom or naturalness to resource management tactics. We

examine the logical structure of recently invoked arguments and summarize recent work to determine whether this

conceptual model is necessary, sufficient and adequately developed to generate specific tactics for management of the

western boreal forest. In particular, we show that biotic elements involved in both the early phases of natural

succession and in the breakdown of coarse woody material will be at risk if only landscape level processes like the

large-scale spatial dynamics of wildfire are considered in mangement. Development of complex stand structure

associated with gap dynamic processes that are characteristic of post-rotational age forest should be accommodated in

extensive forest management to conserve the biota and the processes that ensure long-term maintenance of site

productivity. We preview developing efforts under the Canadian EMEND project to understand how the situation might

be improved.

Keywords: biodiversity, invertebrates, boreal, natural disturbance, wildfire, spatial patterns

INTRODUCTION

Humans generally are impressed by big things, and

trees are the biggest living things that most of us

encounter day to day. Likewise large stands of trees,

comprising primeval forests, intrigue us and provide

comfort and solace to many. The Canadian boreal

forest, as it presently exists in northern Alberta, is

among the last of the Earth's large, wild forests, and

this ecosystem is tightly linked to the identity of many

Canadians. However, the western boreal forest is

being cut at an unprecedented and accelerating rate

(Pratt and Urqhart 1994). By the end of a projected 80

year rotation period started in the present decade,

much of the remaining primeval timber will have fallen

to the saw in Alberta. Never-cut forests of

merchantable volume will remain only on inaccessible

land or in a few scattered parks and reserves, but large

tracts of regenerated forested land will remain to

possibly satisfy our manifold non-timber interests.

Nonetheless, natural dynamics on many scales, which

have heretofore patterned the Canadian boreal

landscape, will be profoundly altered by human activity.

Will it matter? How will we know? Should we do

anything about it? Can we do anything to guarantee a

more favourable outcome? Although we are not in a

position to answer these questions confidently with

respect to many components of unmanaged forests,

including their invertebrate faunas, ongoing research

suggests answers that can be built into adaptive forest

management.

Western societies have started to feel uneasy about the

depth of our imprint on the planet's forests. In

developed countries we increasingly consider forests as

more than a collection of harvestable trees — first as

habitats for charismatic wildlife, and now after the Rio

Summit, as repositories for 'biological diversity'. In

fact, conservation of 'biodiversity' is now widely, and

perhaps somewhat uncritcally, accepted as

representing whole forest values (fig. 1). Land-use

policy makers are increasingly adopting biodiversity

'John Spence, Professor, Department of Biological Sciences, University of Alberta, Edmonton, Alberta, Canada T6G 2E9.

2Chris Buddie, graduate student, Department of Biological Sciences, University of Alberta, Edmonton, Alberta, Canada T6G 2E9.

3Kamal Gandhi, graduate students. Department of Biological Sciences, University of Alberta, Edmonton, Alberta, Canada T6G 2E9.

4David Langor, Research Scientist, Canadian Forest Service, Northern Forestry Centre, 5320-122nd St., Edmonton, Alberta,

Canada T6H 3S5.

5

Jan Volney, Research Scientist, Canadian Forest Service, Northern Forestry Centre, 5320-122nd St., Edmonton, Alberta, Canada

T6H 3S5.

6

James Hammond, Biologist, Canadian Forest Service, Northern Forestry Centre, 5320-1 22nd St., Edmonton, Alberta, Canada T6H

3S5.

7Greg Pohl, Biologist, Canadian Forest Service, Northern Forestry Centre, 5320-1 22nd St., Edmonton, Alberta, Canada T6H 3S5.

80
Figure 1-Forces that guide development of forest

management. Since the Rio Convention on Biodiversity, both

long- and short-term effects on biodiversity have entered the

forest planning domain.

measures to indicate whether ecosystems are being

maintained in a 'functionally integrated state'. In many

countries, some version of the biodiversity criterion is

being explored as a possible addition to those of

regeneration standards and future inventory projection

that have been used to support a sustained yield

approach to allocating harvests (Probst and Crow

1991, Franklin 1993, Angelstam 1997). The new suite

of evolving protocols proposed to regulate harvest and

evaluate forest management plans is termed

"sustainable forest management". Incorporation of the

biodiversity criterion broadens the focus beyond fiber

yield.

The general rationale for the approach adopted under

sustainable forest management is that "nature knows

best". Simply put, forest communities with biodiversity

measures in the range of natural variation, as

established from undisturbed areas used for baseline

studies, are judged to be "healthy", whether or not we

understand the ecological functions of the component

species. This approach to managment of complex

ecosystems acknowledges our ignorance. It prompts

us to tread more gently with respect for the whole

system, rather than insist on solid functional or

economic evidence for retaining any particular

component. Thus, more than a half century after it

was penned, land use policies are being developed in

the context of Aldo Leopold's famous, often-quoted

metaphor: "If the biota, in the course of aeons, has

built something we like but do not understand, then

who but a fool would discard seemingly useless parts?

To keep every cog and wheel is the first precaution of

intelligent tinkering" (Leopold 1953). In this context,

knowledge about non-pest forest invertebrate species

becomes a necessary component of forest

management.

FOREST MANAGEMENT AND

INVERTEBRATE POPULATIONS

Is there any evidence that invertebrates are being lost

as a result of forestry activity? Such losses are

generally accepted as being huge in the tropics (Erwin

1991, Wilson 1992), but undesirable side effects of

forest management also have been revealed by

research on invertebrate communities in northern

forests (Mikkola 1991, Nilsson and others 1995,

Niemela 1997). For example, carabid beetle species

abundant in primeval montane pine forests of western

Canada have not re-established populations in the

oldest available regenerated sites (i.e., up to c. 30

years after harvesting) (Niemela and others 1993,

Spence and others 1996), suggesting that old-growth

specialists will be disadvantaged in harvested

landscapes. In a landmark paper, Siitonen and

Martikainen (1994) compared the saproxylic beetles of

areas in eastern Finland that have been subjected to

intensive logging with those of similar areas of Russian

Karelia exposed to minimal human disturbance for the

past 40-50 years. With comparable effort, they

collected 186 individuals of 1 5 red-listed species from

aspen snags in Russia but found only one individual

representing this same set of species in Finland.

The important point of the recent work on northern

forest invertebrates is that large-scale changes in

empirical patterns of species diversity and composition

have apparently resulted from the way that humans

have repeatedly altered the land through intensive

forest management. Most current forestry practices

homogenize the landscape, leaving regenerated stands

ill-suited for maintenance of the full complement of

species that once inhabited them (Angelstam 1997,

Niemela 1997). Large scale patterns of unplanned

forest fragmentation and the resulting isolation of

species with particular habitat requirements may also

prevent recolonization of regenerating forests, even

when suitable microhabitats develop over time. The

recolonization process depends on spatial and temporal

scales characterizing the metapopulation dynamics of

particular taxa, and thus landscape level planning based

primarily on requirements of vertebrates may be at the

wrong scale to accomodate arthropods species with

different dispersal ability. However, even for

arthropods with excellent ability to seek out and

colonize appropriate microhabitats, it remains critical to

understand and optimize direct, stand-level habitat

effects. If appropriate microhabitats are missing, the

creatures that depend on them and the ecostystem

services that these creatures provide are surely at risk.

Unfortunately, with respect to development of forest

management prescriptions in western Canada, we are

woefully ignorant of much of the biodiversity that we

have agreed to protect as signatories of the Rio

81
Convention. Thus, we are caught on the two horns of

a dilemma: we must conserve this biodiversity, while

forests are being cut at an ever increasing rate. The

rosy ideal of 'sustainable forest management'

represents what we want, but do we have a rational

strategy for attaining this state? Furthermore, we

urgently need to define how such a strategy connects

appropriately to operational measures of biodiversity.

In the balance of this paper we discuss how knowledge

about invertebrates can contribute to adoption of

modern forest management models, and give examples

of research about soil and litter-dwelling arthropods

that can be applied toward this end.

NATURAL DISTURBANCE AS A

MANAGEMENT PARADIGM?

The currently advocated approach to sustainable

management of the boreal forest promotes linkage

between management regimes and natural disturbance.

The boreal forest, like many ecosystems, is reset in

fractal mosaics to earlier successional stages and has

its successional courses altered and diversified by

various natural disturbances (Haila and others 1994).

In the forestry arena, a number of approaches to

management, including "new forestry" (Franklin 1989)

and "ecosystem management" (Kaufmann and others

1994), are converging on the same sort of rationale.

Hunter (1993), for example, characterizes the approach

as follows: "timber harvesting regimes should be

designated to imitate the natural disturbance regimes,

such as fire and windfall, which shape the structure of

forest ecosystems".

In western Canada the forested landscape reflects a

conspicuous influence of wildfire (Rowe 1972, Stelfox

1995). Analogues to fire are already being applied in

management regimes for the mixedwood zone of

western Canada (Stelfox 1995). From the premise that

the boreal biota is well adapted to deal with the natural

disturbance of periodic wildfire, it is inferred that

imposing similar patterns of forest regeneration on the

landscape through harvesting should only minimally

disrupt natural processes and patterns of biodiversity.

Proponents reason further, and probably correctly, that

we cannot know enough to micro-manage our forestry

activities for the benefit of every forest species and,

even if we did, complete knowledge would doubtlessly

paralyze us. The pragmatic approach to resource

development in the light of uncertainty is to apply a

coarse, general filter in forest management. Thus, we

seek to develop and adopt policies for harvest that

ensure existence of an appropriate mosaic of habitats

on the landscape, as defined by natural disturbance

patterns, to maintain our biota as a whole. Fine-filter

management geared to the needs of particular biotic

elements may be established under this approach, but

only for charismatic or exploitable species that are

negatively influenced under the coarse filter approach.

We tend to ignore small, unengaging or poorly known

species, especially those that appear to be unimportant

in socio-economic reasoning.

The above arguments ignore practical constraints and

make significant leaps of faith. Although the approach

seems rational and represents a stage in the process of

improving forest management, it is critical that the

approach not be canonized as a 'magic bullet', and

thus, simply be defined as sustainable forestry. Magic

bullets are simple but powerful presciptions that seem

to give exactly the results we want (Van Den Bosch

1978). Unfortunately, magic bullets, used to solve

environmental problems, frequently fail when the

temporal or spatial frame is expanded and unpleasant

side effects are noticed. Scientists and research

managers may stall recognition of side effects and

development of alternative approaches. This comes

from insisting on the sort of local focus guaranteed to

deliver research dollars from magic bullet makers and

those who want fixed truths on which to maximize

efficiency of profit-oriented operations. A strictly firebased natural disturbance model seems like a magic

bullet to us. It has been most touted by those familiar

with creatures that range widely over a landscape. For

many birds and most larger mammals such as bears

and ungulates, about which we already have

considerable biological knowledge, such an approach

has real appeal. Although our comparative ignorance

of invertebrates may be blissful for a time, there is no

guarantee that little known invertebrates, including

those involved in the biotic engine that renews soil

fertility and other ecosystem functions, will not be

negatively affected by forest management directed at

and monitored using the more charismatic fauna. Such

negative effects could compromise sustainability of

forestry operations in the long term. For this reason, it

is imperative that we continue to study all aspects of

forest systems brought into harvest rotation.

The strong focus on large-scale spatial patterns,

ignores significant concerns from which probable sideeffects may spring. Let us consider four rational

counter arguments to a strict landscape approach. This

is not to dismiss landscape level work but rather to

amplify why it remains essential to study processes

operating at both smaller and larger scales if we are

going to manage them effectively. Afterall, it is the

interaction of these processes that produces natural

forests.

1) In current discussions about pre-eminence of

landscape processes, generation of stand-level variation

across successional scales meaningful for forestry is

frequently ignored. The within-stand diversity of such

harlequin environments is critical to ensure the

persistence of species that might be favoured as

82
background conditions change. Within-stand

homogenization resulting from forestry activities could

move environmental conditions out of the range of

today's levels of natural variation, causing extreme

reductions or extirpations of some species dependent

on rare microhabitats. If so, we would constrain the

system's ability to respond and there would be little

basis for assessing the long-term success of adopting a

natural disturbance model. For example, the boreal

biota is probably not in meaningful equilibrium (Boktin

1990), but is still recovering from the last glaciation

(Prentice and others 1991) and probably responding to

anthropogenic climate change (Lenihan and Neilson

1995). Biodiversity, including small, rare and/or cryptic

species, is a key element in ecosystem response to

change (Wilson 1992, Naeem and others 1994). We

ignore these changes and the ecosystem features

required to buffer them at our peril. Tomorrow need

not be like yesterday.

2) Innovative managers are encouraged to emulate

natural patterns of stand origin. Under a wildfire-driven

natural disturbance model, these are those seen on fire

maps. However, natural but erratic spatial-temporal

patterns are unlikely to be either socially or

economically acceptable for industrial forestry. Even if

some vague approach to pattern matching is

acceptable, what time-frame should be used to

establish the base pattern? Fire frequency and patterns

have clearly varied, even recently, both as an effect of

active fire suppression and as a result of natural

processes (Bergeron and Dansereau 1 993, Gauthier and

others 1995). Should we, for example, force

landscapes to develop in the form defined by processes

that operated in some limited past time window?

3) On occassion, practical implications are reduced to a

worrisome short-cut stated as follows: if we don't cut

more land annually than was scorched in former times,

estimated to be in the range of 2-3% for the boreal

mixed wood, the biota will be conserved. However,

there is little hard scientific evidence to show that any

wildfire-based model, applied strictly at the landscape

scale, will be any better for biodiversity than the old

"cut it and pray" model. Adopting an untested model

does not automatically confer the mantra of

sustainability on boreal forestry (Fanklin 1993). Models

are tested most stringently by experiment and there are

excellent but largely untapped opportunities to develop

appropriate experiments in the context of adaptive

managment (Hilborn and others 1995). It is unlikely

that we have inferred the optimal answer based on

incomplete information.

4) The scenario depicted in figure 2 is generally

accepted as a description of what happens to forested

uplands in the western boreal region (Rowe 1972,

Bonan and Shugart 1989; Stelfox 1995). Upland

stands are initiated by wildfire and, after a very poorly

Figure 2-Schematic diagram indicating the general course of

succession in the boreal mixedwood forest. The two major

harvest points presently envisioned are indicated by the two

dark vertical arrows.

understood early successional period, come to be

initially dominated by trembling aspen. These may

have a white spruce understory right away, or may be

slowly colonized by spruce, depending on the distance

to seed source (Lieffers and others 1996). As a stand

becomes more dominated by spruce, the probability of

stand replacing fire goes up and eventually the stand

burns again to reinitiate the cycle. Two main harvest

points will be imposed as we come to a stable rotation

in the boreal mixedwood forest. The earlier one,

representing a major part of today's harvest in Alberta,

is directed at an ecosystem that is rarely re-initialized

by wildfire. It is not clear that we will allow natural

processes to guide the intervening period to spruce

harvest, but even if we do, the age distribution and

variation of these spruce stands will surely be

truncated. Thus, with respect to the ecological

processes inherrent in natural stand development, it is

fair to ask whether a natural disturbance model is being

followed at all. Should the wildfire analogue be used

more deeply than as a weak justification for removal of

fiber?

The natural disturbance approach is perpetrated mainly

as an idea about risk management. As in a game of

bingo, it is argued that "chance" will determine the

winners and losers as in natural situations. If each

species has a fair or "natural" chance to win (i.e., to

stay in the game) we might legitimately dodge

responsiblity for extinctions. However, we need to

know if the rules of the game, under anthropogenic

forest managment, are really natural for all players.

Thus, we prefer an approach that manages for the

results we want to achieve over the long run, rather

than rushing blindly to emulate selected aspects of

natural process that fit into our socio-economic

constraints. We will be responsible for the outcome

either way because clearly we will not allow nature to

take her own course, even under a natural disturbance

model. This is not to claim that we will ever be able to

83
employ perfect knowledge, but we argue that testing

and applying the limits of what we know is preferable

to 'biodiversity bingo' when we can be sure that the

probabilities of extinction will be affected however we

manage the system.

A superficial, landscape-oriented approach to basing

management on natural disturbance may encourage

false security with respect to biodiversity. In the words

of song-writer Joni Mitchell, it is common for humans

to "pave paradise" as a result of focus on specific

objectives. If our forestry practices homogenize forest

stands, we might not even know what biodiversity

"we've got 'til it's gone". Organizing the size, age and

distribution of "parking lots" carefully according to

some natural rhythm of land disturbance won't help

much if the parking lots themselves are part of the

problem.

BEETLES AND SPIDERS OF BURNS AND

HARVESTS

Recently burned and recently harvested stands are

quite dissimilar to most observers. The question for

those interested in managing in the context of

sustainability criteria is whether such differences throw

biotic succession onto different courses that might

threaten some of the biota. As part of a team of forest

ecologists, we set out to determine how close

succession from burns and havests might be in the

aspen-dominated mixed wood. We asked whether

harvest was a reasonable subsitute for fire with respect

to development of the litter-dwelling invertebrate fauna.

If so, we reasoned that fire- and harvest-origin stands

would develop increasingly similar faunas as they aged.

If the fauna diverged significantly, however, the logic

of using harvest as an analogue for fire may be flawed

with respect to protection of biodiversity.

Finding suitable harvested and burned stands was a

challenging matter, undertaken by P. Lee and his

colleagues at the Alberta Research Council. Using

forest inventories, aerial photography and ground

truthing, they found aspen stands either harvested or

burned in 1968, 1982 and 1995. These sites are

spread out over a wide geographical area in

northeastern Alberta, and thus interpretations of

resulting data may be confounded by geographical

variation. However, this is the best that can be done

with a chronosequence study of aspen forest, given

that harvest of aspen was rare in Alberta before the

late 1980s.

We have focused our studies on two families of litterdwelling beetles (Carabidae, Staphylinidae), and spiders

(Araneae), using pitfall trapping (Spence and Niemela

1994, Digweed and others 1995). The specific

sampling designs will be detailed in future papers.

Here, it suffices to say that we ran trap lines of 6

traps, each separted by c. 50 m, in two stands of each

origin-age combination during 1996 and 1997. More

than 17,000 specimens were collected during the two

year sampling period and the definitive taxonomic

analysis is still in progress. However, our preliminary

findings can be used to illustrate the general level of

similarity between fire- and harvest-origin stands.

There were no great differences in overall abundance of

these litter-dwelling taxa between harvest and wildfireorigin stands (fig. 3). There is a marginally significant

trend for staphylinids to be less abundant in fire-origin

stands, but the other taxa show no significant

differences in overall abundance. Furthermore, the

differences in the first year after fire appear to be much

greater than in year 2. We used rarefaction (Simberloff

1978) as a measure of diversity, and the results show

that overall diversity of litter dwelling arthropods did

not differ dramatically between fire and harvest origin

stands at any age (fig. 4). Although there were

prominent differences in spider diversity between

wildfire and harvest in the first year of the study, these

had largely disappeared in the second year. However,

analyses of diversity and abundance, however, do not

provide much comfort about the fit of harvest effects

to natural disturbance, if the species differ between

harvested and pyrogenic stands.

Figure 3-Least square mean abundance (± standard errors)

by stand origin for beetles and spiders collected by pitfall traps

during 1996.

We have just begun to tackle the problem of

distribution of particular species on the

chronosequence. As a start, we've used cluster

analysis to ask questions about patterns of relative

abundance of species. Results using all litter-dwelling

taxa collected in year one, suggests that pyrogenic and

harvested stands tend to converge on similar species

associations. It appears that litter dwelling

invertebrates recover faster from disturbance by

harvesting than by wildfire as the stands harvested in

84
1982 are most similar to the 1968 fire-origin stands

(fig. 5).

Figure 5-Cluster analysis of Bray-Curtis measures of percent

similarity for carabids, staphylinids and spiders collected by

pitfall trapping during 1996. Data were standardized to 800

trap-days.

Figure 6-Detrended corespondence analysis (DCA) of sample

scores (pitfall traps) derived from 1 53 species of carabids,

staphylinids and spiders collected in 1996.

two chronosequences followed quite different

trajectories to reach the present positions in ordination

space, it may be risky to predict that community

development will be parallel in fire- and harvest-origin

stands in the future.

From a biodiversity perspective we still must get down

to the level of species. Some species of each taxon

are collected only in stands of one age or origin (table

1). Many of these are records of only one or a few

individuals and may thus be somewhat discounted.

However, some records refer to more abundant species

and the data about them are thus less comforting. As

we replace fire with harvest, for example, is it possible

that early successional species characteristic of

pyrogenic stands could drop out of the system? The

aspen-dominated mixedwood seems to have relatively

few litter-dwelling species that specialize in old stands,

but especially among saproxylic species, some may be

at risk (Spence and others 1996, 1997).

We've used detrended correspondence analysis (DCA)

(ter Braak 1987a,b) to further describe these

assemblages. An example of this approach is provided

by figure 6, in which individual trap data about three

taxa are plotted in ordination space defined by the first

two DCA axes. The first axis, which has a high

eigenvalue, separates the stands quite clearly and

consistently with respect to age. As would be

expected from the results of the cluster analysis, the

harvest-82 and harvest-68 stands are closer together in

the ordination than are the comparable fire-origin

stands. Although fire- and harvest-origin assemblages

start off from quite different points in this ordinationspace, they appear to converge in 30-year old stands.

Thus, both cluster analysis and ordination support the

claim that litter invertebrates converge with respect to

general community structure of litter invertebrates by

28 years post-disturbance. However, given that the

BEETLES OF FIRE SKIPS

We are studying the ecological significance of oldgrowth patches of residual forest nested within large

burns by looking in detail at the actions of wildfire, and

at what sorts of habitats it creates. Fire brings about

profound changes in community structure, but it also

creates patchiness and spatial heterogeneity on several

scales. Today's foresters attempt to re-create some of

this local structure by leaving patches of standing

green trees in patch-retention logging systems.

However there has been little research to characterize

invertebrate assemblages of 'fire skips' and to assess

whether they are significant for recolonization of the

surrounding burned sites. We have set out to describe

what happens in both unburned patches and the

burned matrix of large burns by comparing the fauna

85
Table 1-Number of species unique to one type of stand origin (fire or harvest) or age (origin in

1995, 1 982 or 1 968) for two orders of beetles and for ground dwelling spiders. Data from

pitfall traps run during 1996.

between two large sub-alpine burns, one 15 and

another 30 yrs-old, in conifer-dominated forests on the

east slope of the Rockies.

We studied the 15 yr-old burn intensively in 1997 and

will do the same for the older burn in 1 998. As above,

pitfall traps have been used to sample and describe the

carabid fauna of unburned patches of various sizes, of

the burned matrix and of the surrounding old-growth

forest, consisting mainly of Englemann spruce,

lodgepole pine and sub-alpine fir. Transects of traps

have been laid out to run across the boundaries of fireskips and burned areas and of burned forest and the

undisturbed forests at the edges of the burn. The

details of the sampling design will be presented

elsewhere.

Although definitive analyses must await completion of

the second year of field work, some points of interest

emerge from the first year's work. Several numerically

dominant old-growth carabid species, including

Calathus advena and Nebria crassicornis, were found

in 15-year old fire-skips and the surrounding unburned

forest, but were not taken in the burned areas except

near skip edges. One interesting parthenogentic

species, Pterostichus empetricola, which is a

predominantly northern species existing in relict

populations southward along the Rockies, was

collected exclusively from skips. We suggest that this

species requires a narrow range of cool, wet habitat

that is relatively rare in the forest at large, but which

may be disproportionately represented in skips because

of their resistance to burning.

The data suggest that beetles move more readily into

the burns from the fire skips (fig. 7), perhaps because

of population pressures in the smaller skip habitats.

Clearly beetle activity-abundances are higher in fireskips than in either burned or surrounding old-growth

forest. Also, it appears that skips of different sizes

may retain different mixes of beetles species, perhaps

as a result of interspecific interactions intensified in

Figure 7-Evidence for recolonization of burned areas from fire

skips by the carabid beetle, Calathus advena.

these small areas. For example, Nebria crassicornis

was found only in larger skips, where it sometimes

overlapped with small populations of C. advena.

However, in smaller skips, C. advena was abundant

(fig. 8).

Already this study provides several results with

implications for the patch retention approach to

harvesting. Fire skips do provide refuges for oldgrowth invertebrate assemblages and these species do

re-colonize burned patches from the skips. Patches of

different sizes vary in respect to carabid assemblage

retained. Thus, distribution of patch size will be an

important aspect of emulating natural process.

Furthermore, it appears that the location of these skips

is not random with respect to microtopographic

features, and that especially the largest skips, which

occupy wetter microsites, may have remained

unburned through at least two fires. This is not the

random pattern that proponents of the Natural

Disturbance Paradigm have suggested. Thus, the

spatial distribution of patches in a logged stand should

be adjusted to aspects of intrastand variation that

remain to be studied and defined.

86

Taxon

Carabidae

Staphylinidae

Araneae

Disturbance Type

Fire

4

4

19

Harvest

12

12

17

1995

5CJl

15

Year of Stand Origin

1982

1

3

10

1968

3

4

8
Figure 8-Segregation of two carabid species in fire-skips of

differring size.

THE EMEND PROJECT: STAND

STRUCTURE AND INVERTEBRATES

Unmanaged stands differ in structure from stands

managed for fiber production, by generally having more

coarse woody material (CWM) lying about in various

stages of decay. This probably has little to do with

differences in stand origin, but instead stand

developing exclusively under natural processes vary

conspicuously in size, age and spacing of trees and

CWM as result of ongoing GAP dynamics (Botkin 1972;

Shugart and West 1977; Solomon 1986). These

smaller scale, within stand disturbances, which

diversify stands as habitats for other organisms, kick

into high gear as canopy trees begin to senesce, and

thus, aspects of these processes are generally

eliminated or truncated by harvest rotations. Although

surely important components of insect habitats, the

effects of complex and variable stand structure are not

well studied.

We have recently launched a cooperative project that

seeks to determine the significance of stand structure

and to put it to work in management of the northern

mixedwood forest system. The project goes by the

acronym EMEND (Ecosystem Management by

Emulating Natural Disturbance) and occupies >1000

ha of mixed wood forest, c. 120 km northwest of

Peace River Alberta. The project was developed

through nearly 2 years of detailed consultation among

research scientists, practicing industrial foresters and

public land managers. EMEND has been designed both

to meet the needs of industry and to satisfy the

requirements for a rigorous scientific experiment. It

has been linked to the operating plans of Canadian

Forest Products LTD. (CANFOR) and DaishowaMarubeni International LTD. (DMI) and has been

envisioned from the start as a significant research

component for adaptive management. The work

involves a large, multi-disciplinary team, including

biologists, foresters and social scientists.

Our general objectives are as follows: 1) to investigate

how retention of forest structure on cut blocks affects

forest renewal, including regeneration of biodiversity;

2) to compare a number of alternative harvest options

to the effects of wildfire; and 3) to investigate the

cost-benefit ratios of alternative prescriptions. The two

main driving variables are the extent and type of

disturbance and the original forest cover type (aspen

dominated [>70% of canopy], aspen dominated with

strong spruce understory, mixed forest [white spruce

and aspen both <60% of canopy], and conifer

dominated [>70% of canopy]). Experimental burns are

planned at several intensities and the impacts of these

will be compared with those of several practical

harvesting regimes. Integrated study of a wide range

of response variables, including comprehensive work on

invertebrate biodiversity, is planned. A central

component of this project is an effort to connect

studies of nutrient cycling and long-term site

productivity to invertebrate assemblages characteristic

of living trees, dead and dying wood, leaf litter and soil

organic matter.

During 1997, we collected the ground-level baseline

data required to lay out an effective experiment. Our

main goal was to select blocks of relatively

homogeneous stands for the experiment from a pool of

candidate stands identified initially from the enhanced

Alberta Vegetative Inventory. Data about the stands

selected suggest that we are indeed dealing with rather

distinct points along the successional pathway (fig. 9).

Spruce volume increases over the sequence and that of

aspen falls. There are clear differences in age among

the oldest trees found on site. These are somewhat

less different than might be expected from the standard

successional model, adding support to recent ideas that

the process is more complex than frequently envsioned.

The various treatments have been laid out over three

replicate stands representing each of the four cover

types. During summer 1998, we will make pretreatment measurements on all these stands, with

burning and harvest scheduled to begin in the following

fall and winter. Further information about the EMEND

project, including experimental design, specific

biodiversity projects and contacts for questions, can be

found at

http://www.biology.ualberta.ca/emend/emend.html.

Proposals for additional work to take advantage of the

unusual opportunities provided by the EMEND site are

welcome.

CONCLUSIONS

It is appropriate that reason and scientific study

structures the way in which biodiversity is applied in

forest management. However, at present we have too

little sound information about the biota of northern

87
Figure 9-Characteristics of overstory vegetation in stands

selected for the EMEND experiment.

forests to corroborate the hypothesized value of new

management models based on natural disturbance.

Perhaps more than ever before in northern forestry,

adaptive management that weds research and

operations is in order. Society is clear in the message

that biodiversity is to be valued in the context of

management of public lands (Kuusipalo and Kangas,

1994; Abramovitz, 1998). Thus, invertebrate

biologists must clearly connect their studies of

biodiversity to studies of consequences of particular

land-use patterns, if concerns about invertebrates are

to influence those making the management decisions.

It is time for those involved in forest management to

leave the value judgments about biodiversity alone and

ask rigorous questions about the consequences of

adopting any particular concept as a model for forestry

development.

At this point we offer the following tentative answers

to the questions posed at the end of our first paragraph

about the possible influences of extensive forestry

activity in the western boreal forest.

• Will it matter? Leopold's dictum about intelligent

tinkering cautions us against throwing away things

that we don't understand. Even if many forest

species contribute little to the sustainability of the

forest resource in human terms, many are

concerned about the ethical aspects of humancaused extinction.

• How will we know? Forest health monitoring

should incorporate some assessment of poorly

understood taxa like invertebrates, otherwise we

may understand what we have lost only when it is

too late to reverse the process. Experiments that

compare effects of real management prescriptions

and follow these for significant periods of time will

also increase our understanding.

Should we do anything about it? In addition to

monitoring, the fire-skip study suggests that

refugia are important and that their locations could

be critical. Thus, we should characterize the

patterns of sub-landscape level structure that

contribute to the function of these refuges and

incorporate such information into cutting plans.

Can we do something about it? This depends both

on forestry policies, market pressures and land

owners. Balancing study of and attention to

biodiversity concerns against the need to generate

revenues from a fully allocated forest resource will

be tangible proof of commitment.

ACKNOWLEDGMENTS

We gratefully acknowledge research funding from both

the public (Biodiversity Challenge Grants Program,

Canadian Forest Service, Natural Sciences and

Engineering Research Council of Canada, Sustainable

Forest Management Network of Centres of Excellence

and private sectors (CANFOR, DMI and Weyerhaeuser

Canada). In addition, we thank personnel of AlbertaPacific Forest Industries, Alberta Research Council,

CANFOR, DMI and Weyerhaeuser for direct field

support and encouragement of our research.

Discussions with D. Gilmore, J. Lattin, V. Lieffers, S.

Luchkow, J. Niemela, B. Stelfox and T. Vinge helped

shape the ideas presented in this paper. J. Scott

polished and produced the final figures.

QUESTIONS FROM THE FLOOR

Q: I'm wondering if the "magic bullet" isn't stand

dynamics and the focus on these scale processes.

What about the relationship between biodiversity and

landscape dynamics, the "neighborhood" effect on

disturbance processes landscape structure?

A: Let me begin by saying that I have no faith in any

sort of magic bullets. What we've tried to illuminate is

how the present preoccupation with landscape-level

processes leaves out concerns important to

conservation of species with dynamics on smaller

spatial, but perhaps longer temporal scales. I do not

argue that work at the landscape scale is not important

or that is should not be done, only that it isn't the beall and end-all for sustainable forestry. We need to be

concerned that our forestry practices, as applied at the

stand-level, do not homogenize whole regions with

respect to microhabitats critical for invertebrate

species, like for example, in the case of the saproxylic

beetles studied by Sittonen and Martikainen (1994).

88
Assuming that by "landscape dynamics" you mean

changes in the the size and distribution of forest stands

of various ages over time, then I think this is possibly

important and we've commented on it to some degree

in Spence and others (1996, 1997). Indeed colonists

available to any isolated habitat unit come from the

surrounding matrix. However, the surrounding matrix in

forest land is composed of stands as well, and stands

are the focus of how harvest technology affects the

land, repeatedly across a landscape. The important

point that I want to re-emphasize is that we don't

know enough about how stand level processes like gap

dynamics create habitat complexity that may be crucial

for biodiversity. It is clear that short rotation forestry

limits the action of these processes. Will there be

undesirable consequences? The available evidence does

not rule them out.

I cannot address the last part of the question because

I just don't understand it.

LITERATURE CITED

Abramovitz, J. N. 1998. Taking a stand: cultivating a

new relationship with the world's forests.

Worldwatch Paper 140, Worldwatch Institute,

Washington, D.C. 84 p.

Angelstam, P. 1997. Landscape analysis as a tool for

the scientific management of biodiversity. -

Ecological Bulletins 46:140-1 70.

Begeron, Y.; Dansereau, R.P. 1993. Predicting the

composition of Canadian southern boreal forest in

different fire cycles. Journal of Vegetation Science

4:827-832.

Bonan, G.B.; Shugart, H.H. Jr. 1989. Environmental

factors and ecological processes in boreal forests.

Annual Review of Ecology and Systematics 20: 1 -

28.

Botkin, D.B. 1972. Some ecological consequences of

a computer model of forest growth. Journal of

Ecology 60:849-872.

Botkin, D.B. 1990. Discordant harmonies: a new

ecology for the 21st century. Oxford University

Press, New York. 241 p.

Digweed, S.C.; Currie, C.R.; Cáircamo, H.A; Spence,

J.R. 1995. Digging out the "digging-in effect" of

pitfall traps: influences of depletion and

disturbance on catches of ground beetles

(Coleoptera: Carabidae). Pedobiologia 39:561-576.

Erwin, T.L. 1991. An evolutionary basis for

conservation strategies. Science 253:750-752.

Franklin, J.F. 1989. Toward a new forestry.

American Forests 95:37-44.

Franklin, J.F. 1993. Preserving biodiversity: species,

ecosystems, or landscapes. Ecological

Applications 3:202-205.

Gauthier, S.; Gagnon, J.; Bergeron, Y. 1995. Forest

dynamics modelling under a natural fire cycle: A

tool to define natural mosaic diversity in forest

management. Environmental Monitoring and

Assessment 39:417-434.

Haila, Y,; Hanski, I.P.; Niemelä, J.; Punttila, P.; Raivio,

S.; Tukia, H. 1994. Forestry and the boreal

fauna: matching management with natural forest

dynamics. Anna/es Zoologici Fennici 31:187-202.

Hilborn, R.; Walters, C.J.; Ludwig, D. 1995.

Sustainable exploitation of renewable resources.

Annual Review of Ecology and Systematics 26:45-

67.

Hunter, M.L. 1993. Natural fire regimes as spatial

models for managing boreal forests. Biological

Conservation 65:115-120.

Kaufmann, M.R.; Graham, R.T.; Boyce, D.A., Jr.; Moir,

W.H; Lee, P.; Reynolds, R.T.; Bassett, R.L.;

Mehlhop, P.; Edminster, C.B.; Block, W.M.; Corn,

P.S. 1994. An ecological basis for ecosystem

management. USDA Forest Service, General

Technical Report RM-246. 22p.

Kuusipalo, J.; Kangas, J. 1994. Managing biodiversity

in a forestry environment. Conservation Biology 8:

450-460.

Lennihan, J.M.; Neilsen, R.P. 1995. Canadian

vegetation sensitivity to projected climatic change

at three organizational levels. Climatic Change 30:

27-56

Leopold, A. 1953. Round River. Oxford University

Press, New York. 173p.

Lieffers, V.J.; Stadt, K.J.; Navratil, S. 1996. Age

structure and growth of understory white spruce

under aspen. Canadian Journal of Forest Research.

26:1002-1007

Mikkola, K. 1991. The conservation of insects and

their habitats in northern and eastern Europe.

Pages 109-119 in Collins, N.M.; Thomas, J.A.

(eds.). The Conservation Insects and Their

Habitats. 15th Symposium of the Royal

89
Entomological Society of London, 14-15

September, 1989, London, England. Academic

Press, Toronto.

Naeem, S.; Thompson, L.J.; Lawler, S.P.; Lawton,

J.H.; Woodfin, R.M. 1994. Declining biodiversity

can alter the performance of ecosystems. Nature

368:734-736.

Niemela", J. 1997. Invertebrates and boreal forest

management. Conservation Biology 11:601 -610.

Niemela, J.; Langor, D.W.; Spence, J.R. 1993.

Effects of clear-cut harvesting on Boreal groundbeetle assemblages (Coleoptera: Carabidae) in

western Canada. Conservation Biology 7:551-562.

Nilsson, S.G.; Arup, U.; Baranowski, R.; Ekman, S.

1995. Tree-dependent lichens and beetles as

indicators in conservation forests. Conservation

Biology 9:1208-1215.

Pratt, L.; Urquart, I. 1994. The Last Great Forest.

NeWest Press, Edmonton, Canada. 222 p.

Prentice, I.C.; Bartlein, P.J.; Webb, T. III. 1991.

Vegetation and climate change in eastern North

America since the last glacial maximum. Ecology

72:2038-2056.

Probst, J.R.; Crow, T.R. 1991. Integrating biological

diversity and resource management. Journal of

Forestry 89:12-17.

Rowe, J.S. 1972. Forest regions of Canada.

Canadian Forest Service Publications, Ottawa.

172 pp.

Shugart, H.H.; West, D.C. 1977. Development of an

Appalachian deciduous forest succession model

and its application to assessment of the impact of

chestnut blight. Journal of Environmental

Management 5:161-179.

Siitonen, J.; Martikainen, P. 1994. Occurrence of rare

and threatened insects living on decaying Populus

tremula : a comparison between Finnish and

Russian Karelia. Scandinavian Journal of Forestry

Research 9:185-191.

Simberloff, D.S. 1978. Use of rarefaction and related

methods in ecology. Pages 150-165 in Dickson,

K.L.; Garins Jr., J.; Livingston, R.J. (eds.).

Biological data in water pollution assessment:

quantitative and statistical analysis. American

Society for Testing and Materials, STP 652.

Solomon, A.M. 1986. Transient response of forests

to CO2-induced climate change: simulation

modeling experiments in eastern North America.

Oecologia 68:567-579.

Spence, J.R.; Langor, D.W.; Hammond, H.E.J.; Pohl,

G.R. 1997. Beetle abundance and diversity in a

boreal mixedwood forest. In Watt, A.D.; Stork

N.E., eds. Forests and Insects: Proceedings of the

18th Royal Entomological Society Symposium; 1 3-

15 September 1995; London: Chapman and Hall:

285-299.

Spence, J.R.; Langor, D.W.; Niemela, J.; Carcamo, H.

A.; Currie, C.R. 1996. Northern forestry and

carabids: the case for concern about old-growth

species. Annales Zoologici Fennica 33:1 73-1 84.

Spence, J.R.; Niemelä, J.K. 1994. Sampling carabid

assemblages with pitfall traps: the method and the

madness. Canadian Entomologist 126:881-894.

Stelfox, J.B. (ed.). 1995. Relationships between

stand age, stand structure, and biodiversity in

aspen mixedwood forests in Alberta. Jointly

published by Alberta Environmental Centre

(AECV95-R1), Vegreville, Alberta, and Canadian

Forest Service (Project No. 0001 A), Edmonton,

Alberta.

ter Braak, C.J.F. 1987a. Ordination. Pages 91-173 in

Jongman, R.H.G.; ter Braak, C.J.F.; van Tongeren,

O.F.R. (eds.). Data Analysis and Landscape

Ecology. Center for Agricultural Publishing and

Documentaion (PUDOC), Wageningen, The

Netherlands.

ter Braak, C.J.F. 1987b. CANOCO- a fortran program

for canonical community ordination by partial

detrended correspondence analysis, principal

components analysis and redundancy analysis.

TNO Institute of Applied Computer Science,

Statistics Department, Wageningen, Netherlands.

Van Den Bosch, R. 1978. The pesticide conspiracy.

Doubleday & Co., Garden City, New York. 226 p.

Wilson, E.O. 1992. The Diversity of Life. W.W.

Norton and Company, New York, New York.

424 p.

90
MANAGEMENT EFFECTS ON SOIL ORGANISMS

AND PROCESSES
SOIL ORGANISMS AND RANGELAND SOIL HYDROLOGICAL FUNCTIONS

Jeffrey E. Herrick1

ABSTRACT

Soil organisms control water distribution in rangelands by creating macropores which rapidly conduct water into the

soil, by generating stable soil aggregates which prevent crusting and increase water holding capacity, and by controlling

litter decomposition. Ants, termites and earthworms have all been shown to increase infiltration capacity; anecdotal

evidence suggests that macropore formation by a variety of other macroinvertebrates may have equally dramatic

effects. Macroinvertebrates generate aggregates in the form of fecal pellets. Lichens, mycorrhizal fungi, cyanobacteria

and other microorganisms also contribute to aggregate formation and soil surface stabilization. The direct effects of

rapid litter decomposition on infiltration and soil water storage in arid and semi-arid rangelands is generally negative:

litter removal exposes the surface to raindrop impact, which leads to the formation of physical crusts, and increases

evaporation from the soil surface. These negative impacts are balanced by the creation of surface-connected

macropores and the incorporation of soil organic matter, resulting in the formation of stable aggregates.

Keywords: hydrology, infiltration, soil water, available water holding capacity, invertebrate, ant, termite, earthworm,

fungi, rangeland, soil structure, aggregate, soil organic matter

INTRODUCTION

Biomass production in most North American rangelands

is limited by soil moisture availability during at least

part of the year. Annual precipitation is low, variable

often arrives during periods which are unsuitable for

plant growth. As a result, monthly potential

evapotranspiration (the amount of water which could

be lost to the atmosphere from a free water surface)

may exceed precipitation by a factor of 100 or more

during periods when temperatures are high enough to

support plant production (Hargreaves and Samani

1991).

These limitations are compounded on some parts of the

landscape by high runoff rates. In closed-basin

systems, this water is conserved in lower parts of the

landscape, while in open systems it is effectively lost.

These losses are not necessarily negative from a

regional perspective, as the runoff supports riparian

systems and is frequently collected in reservoirs for

urban and agricultural uses. Losses from the rooting

zone may also contribute to groundwater recharge.

High rates of runoff, however, are rarely supported as a

management objective.

Water redistribution also occurs at much smaller scales.

Redistribution at the scale of tens of meters has been

cited to explain patterns of banded vegetation in many

arid ecosystems (Tongway and Ludwig 1994).

Schlesinger and others (1990) suggested that water

and nutrient redistribution at the level of the individual

plant are associated with the transition of grasslands to

shrublands.

In addition to redistribution, the capacity of the soil to

store water also varies across the landscape. The

volume of plant-available water which can be stored at

any particular point and depth in the profile depends on

complex plant-soil interactions. In general, however,

more highly structured soils have a higher capacity to

store water at tensions at which it is accessible to

plants, and provide conditions which favor root

exploration of a larger soil volume.

Soil organisms can affect water redistribution, soil

water holding capacity and the accessibility of water to

plants through a variety of mechanisms. They modify

soil structure, alter the form and spatial distribution of

above- and below-ground detritus and, ultimately,

change the relative growth rates of different plants and

therefore plant cover and community composition. The

objective of this paper is to review the role of soil

organisms in altering rangeland soil hydrology through

aggregate and macropore formation. Microbiotic crusts

are only briefly considered here. For a more extensive

discussion of their role please see recent reviews by

Warren (1995), Ladyman and Muldavin (1996), the

NRCS (1997), Belnap (this volume) as well as a

forthcoming book in the Advances in Ecology series.

Recent papers by Eldridge and Greene (1994) and

1

Jeffrey E. Herrick is a soil scientist with the USDA-ARS Jornada Experimental Range, P.O. Box 30003, MSC 3JER,

Las Cruces, NM 88003-8003 (jherrick@nmsu.edu).

91
Eldridge and others (1997) have explored the effects of

these crusts on soil surface hydrology in Australia.

SOIL WATER INFILTRATION

The volume of water which enters the soil profile at a

particular point on the landscape is a function of (1)

supply or the depth of water which is available for

infiltration, (2) residence time, or the amount of time

which water is resident on the soil surface, and (3)

infiltration characteristics. Hydrologists often describe

soil water infiltration in terms of sorptivity, which is

related to the initial rate of water movement into a dry

soil, and the saturated hydraulic conductivity, or the

rate of water movement into the soil after it has

become saturated. Soil organisms affect both

residence time and infiltration characteristics. They can

also affect the water supply at one location by

modifying the soil surface in adjacent patches, or at

other locations higher in the landscape.

Supply

Water available for infiltration is a function of both total

precipitation and run-on. The unique plant communities

associated with riparian zones and playa lakes owe

their existence to supplemental water from higher parts

of the landscape. Most studies focus on the effects of

soil organisms at a single point and down-slope effects

of changes in the populations and activity of soil

organisms are rarely considered.

Residence Time

Soil biota affect residence time by modifying surface

roughness and the distribution of litter on the soil

surface. Both soil surface roughness and the presence

of litter slow the movement of water across the

landscape. In some cases, small "ponds" may be

formed by litter dams (Ludwig and Tongway 1995) or

by micro-depressions formed by surface features

stabilized by microbiotic crusts (West 1990). Increases

in residence time by litter can be important in

rangelands which depend on brief, high-intensity

storms. Ponds, even if only 10-25 millimeters deep,

can be locally significant: 25 millimeters is equivalent

to over 10 percent of the average annual precipitation

in much of the western United States and northern

Mexico.

Soil organisms affect litter distribution and residence

time through at least four processes: stabilization,

shredding, burial, and consumption. Fungi can

temporarily stabilize litter and limit redistribution by

wind and water by binding it to the soil surface.

Shredders, including both vertebrates and

invertebrates, modify the susceptibility of litter to both

redistribution and further degradation by reducing

fragment size. Litter is removed from the surface

through both burial and consumption. Ants tend to

bury material in their nests, where it may later be

consumed directly or used to culture fungi. Anecic

species of earthworms, such as Lumbricus terrestris,

drag litter into burrows, while epigeic species consume

material at the surface (Edwards and Bohlen 1996). A

third group (the endogeic species) function solely in the

mineral soil and do not contribute to removal of litter

from the surface. Dung beetles exhibit an even wider

variety of behaviors. Some species bury dung in the

soil directly below dung patches (paracoprids), others

roll dung balls to another location where it is buried

(telecoprids), while a third group lays eggs directly

inside the dung patch, leaving the larvae to consume at

the soil surface (Bornemissza 1969). Most other

important species consume material at the surface and

may deposit fecal material both above- and belowground.

While high levels of soil biotic activity are normally

viewed as an indicator of a healthy system, the direct

effects on infiltration of litter removal from the soil

surface are generally negative. In addition to slowing

runoff, litter protects the soil surface from raindrop

impacts. This limits the formation of soil physical

crusts. This is particularly important in arid and semiarid rangelands where vegetative canopy cover is

generally much lower than 100 percent.

The net impact of direct and indirect effects of litter

decomposition on infiltration depend on complex

interactions with a number of other processes including

the formation of soil structure by the decomposers and

changes in soil microtopography. For example, Trojan

and Linden (1992) reported that infiltration was

deepest where earthworm burrows were located at the

bottom of depressions in the soil surface.

Infiltration Characteristics

Infiltration depends on the density, size distribution,

and continuity of soil pores. It also depends on the

degree of repulsion between soil and water

(hydrophobicity), particularly in the early stages of

precipitation events. Soil organisms form pores

through two basic processes: excavation and

aggregation. A third process, root decomposition, also

results in the development of new voids.

Pore formation by excavation-Excavation generally

results in the formation of transmission pores, or those

pores which transmit water during infiltration. These

pores are > 30μ in diameter (Marshal and others 1996)

and are commonly classified as mesopores (30 - 75μ)

and macropores (> 75μ) (Soil Science Society of

America 1997). Earthworms, ants and termites are

arguably the three groups of invertebrates responsible

92
for the majority of biopore formation in rangeland soils.

Direct observations during rainstorms indicate that a

number of other species including burrow-forming

vertebrates such as pocket gophers (Family

Geomyidae), prairie dogs (Cynomys spp.) and kangaroo

rats (Spectabilis spp.) are probably locally important.

Lee and Smettem (1995) state that, "[Earthworm]

burrows are the most numerous and most important

macropores of animal origin in nearly all soils."

Earthworms form burrows by forcing their bodies

through the soil, and by swallowing soil to be later

excreted as casts either elsewhere in the soil, or at the

soil surface. These burrows may range from one to

over ten millimeters in diameter. Anecic worms, or

those which move between the surface and the

subsoil, have been shown to be particularly effective at

improving infiltration capacity. Up to 2,000 burrow

entrances per square meter have been reported (Lee

and Foster 1991). Even much lower densities can

dramatically increase infiltration due to the fact that the

burrows frequently connect the surface to much deeper

layers of soil.

The relative importance of earthworms in rangelands

has not been evaluated, however. A recently revised

book by Edwards and Bohlen (1996) provides a

comprehensive review of the role of earthworms in

soils, but contains few references to non-arable

systems. The key anecic species are virtually absent

from many arid and semi-arid rangelands, including the

Chihuahuan (Whitford 1996) and Namib Deserts

(Crawford and others 1993). They do occur in moister

parts of the Chilean Arid Zone where coastal fogs

appear to maintain higher levels of soil moisture in spite

of annual rainfall of less than 200mm (Crawford and

others 1993). The lack of earthworms in many

rangeland ecosystems suggests that other species,

including ants and termites, dominate biological

macropore formation.

Rangeland termites can be divided into two groups

based on whether or not they build mounds. Only soildwelling termites (non-mound builders) are found in

North America, occurring throughout the southwestern

United States and Mexico. They form extensive nearsurface galleries and tunnels which extend deep into

the soil. Most are detritivores, feeding on dead

material below the soil surface. Those which have the

highest direct impact on infiltration feed on litter and

standing dead vegetation at the surface. Some, like

Gnathamitermes tubiformans form protective sheaths

around vegetation, while others, including Amitermes

wheeleri feed directly at the surface during periods

when evaporation rates are minimal. Soil-dwelling

termites can generate very high densities of surfaceconnected macropores, particularly under concentrated

sources of organic matter such as cattle dung.

Infiltration rates in a Costa Rican pasture were over

twice as high (71 mm/h) under dung patches which

had been decomposed primarily by termites than they

were in adjacent control plots (34 mm/h) (Herrick and

Lal, 1995).

The positive impact of termites on infiltration through

macropore development may be balanced by the

negative effects of high rates of litter decomposition.

This is illustrated by a comparison of two studies in

which termites were eliminated with the use of the

insecticide, Chlordane. Rainfall simulation was used in

both cases to measure infiltration capacity. Four years

after termite removal from a Chihuahuan Desert

creosote shrubland, infiltration rates had declined by

over 40 percent relative to control in areas with less

than 5 percent perennial cover. In plots centered on

creosote bushes (Larrea tridentata), there was no

change (table 1; Elkins and others 1986). In a

shortgrass prairie in Texas, cumulative infiltration

actually increased relative to control plots two to three

years after termites were removed (table 2; Spears and

others 1975). This increase was associated with a 50

percent increase in organic carbon in the top 1

centimeter of soil and an 80 percent increase in litter

mass.

Table 1 —Final measured infiltration rates (mean ± S.E.) based

on rainfall simulation on 1m2 plots on a Chihuahuan Desert

bajada four years after termite exclusion (data from Elkins and

others 1986).

No shrub; < 5 percent

Erioneuron pulchellum

Larrea tridentata canopy

Termites

excluded

Termites

present

mm/hour

51.3±6.8

106.4±9.7

88.4 ± 5.6

100.6±6.1

In Australia, subterranean termites (Drepanotermes

spp.) increase the density of surface-connected

macropores and reduce it in others, leading to

redistribution of water at the patch (1 meter) scale.

These termites form a cap of up to 2 meters in

diameter directly above their nests, and macropores

lead to the surface in the area surrounding the nest.

Eldridge (1994) reported that ponded infiltration rates

into the nest margins were over 16 times higher than in

the caps. Rates in control areas were similar to slightly

lower than those recorded in the nest margins. In one

of the few studies in which soil-dwelling termite

activity has been evaluated as a potential management

tool, Mando (1997) found that infiltration into a bare,

crusted soil was increased by termite activity, and that

there was a strong positive interactive effect of mulch

and termite treatments on both infiltration and soil

93
water storage. Subsequent microscopic and

computerized image analysis confirmed that 60 percent

of the macroporosity in the top 10 centimeters could

be attributed to termite tunnels and chambers (Mando

and Miedema 1997). Enhanced termite activity has

also been shown to increase macropore flow in two

degraded soils in the Chihuahuan Desert (unpublished

data).

Table 2 —Characteristics of plots two to three years after

initiation of termite exclusion from a shortgrass prairie (data

from Spears and others 1975).

nests of a Camponotus species apparently had no

effect in the woodland (fig. 1c). Two species of

Rhytidoponera also had no effect on infiltration into a

shrub-dominated sand (fig. 1d) (Lobry de Bruyn and

Conacher 1994).

The high variability in nest morphology of different

species and even within species suggests that effects

on infiltration should also vary. Some authors have

speculated that while infiltration is increased at nest

entrances, it may be reduced in the surrounding area

due to vegetation removal (Lobry de Bruyn and

Conacher 1990) or by the generation of impermeable

caps by species such as Trachymyrmex smithii and

Pogonomyrmex maricopa (Whitford 1993) or

Pogonomyrmex maricopa (W.G. Whitford pers.

commun.).

Where ant nests do increase infiltration, their

importance is magnified by the fact that they are able

to conduct water deep into the profile (Eldridge 1 993)

where it is protected from rapid evaporative loss. This

is a function of both the diameter and the continuity of

the macropores in the nests. Soil moisture content

under harvester ant nests (Pogonomyrmex wheeleri,

Cole) in a southeastern Idaho sagebrush-grassland was

less than or equal to that in control areas to a depth of

40 cm. From 60 to 100 cm, however, soil moisture

content was higher under nests (Laundré 1990).

Pore formation by aggregation-Pores are formed

each time three or more soil particles are linked

together as aggregates. These intra-aggregate pores

tend to be much smaller than those formed by

excavation, and are more likely to be involved in water

storage (Hindell and others 1994), while pores formed

between individual aggregates are more likely to

contribute to water transmission. Stable aggregates at

the soil surface also help maintain infiltration capacity

by limiting physical crusting and blockage of surfaceconnected macropores during storms.

A general hierarchical model of aggregate structure and

formation has been proposed by Tisdall and Oades

(1982) and refined by Oades and Waters (1991). This

model is based on the assumption that soil

macroaggregates (> 250μ in diameter) are formed from

more stable microaggregates (50 - 250μ) and that

different binding mechanisms are dominant at each

scale (table 3). Elliott (1986) found that the model

could be applied to a Nebraska soil which had been

under sod for the previous 14 years. More recent

studies have confirmed that the form of soil organic

matter is at least as important as total content in the

formation of stable aggregates (reviewed in Herrick and

Wander 1998).

Soil organisms contribute to soil aggregation directly by

forming fecal pellets, by re-arranging particles during

94

Cumulative infiltration

Organic carbon: top 0 - 1 cm

Litter

Termites

excluded

Termites

present

- mm/40 minutes --

18.5 15

- percent --

1.8

- gm2 -

63

1.2

35

Ants create extensive networks of voids and

macropores in many rangeland ecosystems. Whitford

and DiMarco (1995) calculated that they move one

centimeter of soil to the surface per 100 years in a

Chihuahuan Desert grassland. Assuming that most of

the soil was derived from the top meter, this represents

a porosity increase of 1 percent, or a 2 percent addition

to the typical pore volume for a sandy loam soil. At

nature reserve in New South Wales, Australia, funnel

ants (Aphaenogaster barbigula) alone cover 2.5 percent

of the soil surface per year with their mounds, which

they move approximately every 9 months (Eldridge and

Pickard 1994). The authors calculate that up to 92.5

percent of the soil surface would be affected after 100

years, resulting in a net transport to the surface of 2.8

centimeters of soil.

Infiltration around entrances has been measured for

only a few species. Several Australian studies suggest

that ant activity has a generally positive impact on

infiltration, at least when the entrance is included.

Ponded rates were four (sandy loam) and eight (loam)

times higher over funnel ant (A barbigula) nests

(Eldridge 1993, 1994). Infiltration was positively

correlated with nest entrance diameter on both soils. A

study by Lobry de Bruyn and Conacher (1994)

indicates that these effects probably depend on ant

species, soil type and land use. Water infiltration under

ponded or near-ponded (5 millimeter tension) conditions

was higher under Pheidole sp. nests located on a sandy

loam in a woodland and on farmland (fig. 1a, b), while
Figure 1 —Effects of Pheidole sp. nests on ponded steady state infiltration into a yellow sand (a) and grey sandy loam (b) in farmland,

of Camponotus sp. and Pheidole sp. into a grey sandy loam in a woodland (c), and of two Rhytidoponera species into a yellow sand

under heath (d) in Western Australia.

movement through the soil, by cross-linking particles

with their bodies (fungal hyphae, for example) and by

generating products such as mucilages which act as

binding agents. Fecal pellets are formed in a wide

variety of shapes and sizes. Microarthropods produce

pellets which are generally less than 1 millimeter in

diameter (Lee and Foster 1991), while earthworm casts

can be well over one hundred times that size (Edwards

and Bohlen 1996). As fecal pellets tend to be relatively

concentrated sources of energy and nutrients, they are

rapidly colonized or consumed by other organisms in

the detrital food web. In cattle dung patches colonized

by both dung beetles and termites, fecal pellets

produced by dung beetle larvae are rapidly attacked by

termites (Herrick and Lai 1995). Some pellets,

however, persist to form the nucleus for long-term

stable aggregates. Earthworm casts, which are

predominantly inorganic and therefore less susceptible

to attack, may account for over 50 percent of the soil

mass in some grassland soils (Lee and Foster 1991).

Table 3 —Primary binding agents for aggregates of different

diameters. Note that larger aggregates may be formed from

smaller aggregates and/or from primary particles such as sand

grains.

Aggregate Diameter (μ)

200 - 2000

20 - 200

2 - 20

<2

Primary binding agents

Roots

Hyphae

Fecal pellets

Hyphae

Bacteria

Decomposition products

Persistent organic materials

Primarily inorganic

95
The presence of fungal hyphae has been shown to

promote stable aggregate formation in many grassland

soils (Tisdall 1991; Degens and others 1996;

Chantigny and others 1997). Organic amendments

including manure lead to rapid increases in hyphal

length density (Roldan and others 1996; Degens and

others 1996). The impact of these amendments on

aggregate formation appears to depend on

environmental factors controlling extension of the

hyphae away from the organic substrate into the

surrounding soil (Degens and others 1996).

SOIL WATER HOLDING CAPACITY

Soil organisms affect soil water holding capacity

through effects on pore volume and size distribution

and on the content and form of soil organic matter.

Increases in soil organic matter generally increase soil

water holding capacity (Hudson 1994) both because of

its high absorptive potential and because of its role in

the formation and maintenance of soil structure. The

water holding capacity of organic matter varies widely

depending on its form. Carbon in the form of plant

lignin holds relatively little water, but microorganisms

can convert this to polysaccharides, which can retain

many times their weight in water.

The most important pores for water retention are those

located within soil aggregates. Water is held more

tightly in small pores than in large pores and water in

the smallest pores is unavailable to plants. Clay soils

hold much more water than loamy soils, but much of

this water is never extracted because it is tightly bound

in microscopic pores. Information on soil water

content is difficult to interpret without information on

the tension of the water. The smallest pores inhabited

by bacteria are on the order of 0.2 μ in diameter

(Hassink and others 1993). This is the minimum pore

diameter from which it is assumed that most plants can

extract water (Marshall and others 1996), although

there are reports of rangeland plants continuing to

transpire at much lower (more negative) tensions.

RELATIONSHIPS TO VEGETATION

Soil hydrology, soil biota and vegetation are intimately

linked in rangelands. While the discussion here

emphasizes direct impacts of soil biota on infiltration, it

is essential to consider the positive and negative

feedbacks between soil biota, changes in the soil water

regime and vegetative cover, composition and spatial

distribution (fig. 2). Increases in plant production

generally follow an improvement in the soil water

regime in water-limited environments. Higher primary

production means more material to support the soil

food web, leading to a positive feedback loop. Other

Figure 2 —Simplified illustration of relationships between ant

and termite activity, and changes in plant-available water.

papers in this volume describe interactions between soil

biota and vegetation. Plant canopy cover is frequently

cited as the single most important factor controlling

infiltration on rangelands (Spaeth and others 1996).

Much of the strength of this relationship is probably

due to the association of soil biotic activity with plants,

as described above. However, plant canopies also

directly enhance infiltration by protecting the soil

surface from destructive raindrop impacts. Plant bases

increase residence time by increasing tortuosity,

slowing the movement of water across the landscape.

Different types of soil organisms alter the hydrologic

regime in different ways, affecting both the amount

and vertical distribution of water in the profile. For

example, anecic earthworms and deep-burrowing

harvester ants may favor shrubs by rapidly conducting

water to deeper soil layers, while epigeic worms and

the smaller Dolichoderine ants would tend to

concentrate water in the upper layers of the profile.

Similarly, colonization of litter by different species of

fungi will change the relative hydrophobicity of the soil

surface, altering water redistribution patterns at the

surface. Few of these soil organism-vegetation

feedbacks have been explored and even fewer tested.

RESILIENCE

The contribution of soil organisms to the resilience of

hydrologic functions following disturbance has received

relatively little attention. Rangeland soils are constantly

subjected to structural degradation caused by raindrop

impact, overland flow, wind and the activities of both

native and domesticated animals. These processes

96
ultimately lead to physical crusting and compaction.

Physical crusts limit infiltration at the surface. This

type of crust is distinguished from biologically-stabilized

crusts which can enhance infiltration in many cases

(Belnap in this volume). Compaction reduces

movement of water to deeper layers and can limit

surface infiltration during prolonged storms.

The balance between structural degradation and

recovery is a function of both biological and physical

processes. The importance of biological processes

increases in coarse-textured soils. In fine-textured soils

charged clay particles bind to each other and to other

particles when they are brought into contact by wetdry and freeze-thaw cycles. In some cases,

information on the contribution of soil organisms may

be used to promote recovery of degraded systems

(DePuit and Redente 1988; Whisenant 1996; Herrick

and others 1997). There is a high potential to change

the quantity and distribution of plant water availability

in rangelands by manipulating soil organism populations

through changes in disturbance regimes and organic

matter inputs (Lee and Smettem 1995). However, with

the exception of studies on earthworms in croplands,

the inclusion of Rhizobium with seed, the addition of

mycorrhizal fungi to extremely degraded mineland

(Allen 1988), and the work on termites by Mando

(1997), there have been few attempts to develop

management tools which directly incorporate soil

organisms.

The studies cited in this review, along with preliminary

results from recently established studies in southern

New Mexico (unpublished data) indicate that recovery

of soil structure in a variety of rangeland ecosystems

following both small and large disturbances depends on

soil organism activity. Resilience is believed to increase

with the number of species per functional group

(Naeem 1998). Species redundancy appears to be

extremely low for hydrologic functions in at least some

rangeland ecosystems. For example only two species

of termites {Gnathamitermes tubiformans and

Amitermes wheeleri) contribute significantly to

macropore formation in the northern Chihuahuan

desert. The impacts of removing or limiting the activity

of just these two species can be quite dramatic, as

illustrated by the termite removal study described

above (Elkins and others 1986).

SUMMARY AND CONCLUSIONS

Soil organisms have significant impacts on hydrologic

processes and may play a key role in controlling water

redistribution on the landscape across a broad range of

rangeland ecosystems. These impacts vary across

functional groups, microsites, plant communities and

regions. Larger macroinvertebrates are responsible for

macropore formation and for the breakdown and

incorporation of litter, controlling the residence time on

the soil surface. Smaller invertebrates, bacteria and

fungi contribute to aggregate formation and thereby

soil water storage capacity. An understanding of soil

biology and the direct and indirect (through vegetation)

effects on hydrology can already be used in some areas

to direct or guide management practices involving soil

surface disturbance and the removal or incorporation of

organic matter.

ACKNOWLEDGMENTS

M. Walton provided an early review of this manuscript.

QUESTIONS FROM THE FLOOR

Q: What is the relative importance of hydrophobicity for

water infiltration in more xeric rangeland environs?

A: Hydrophobicity varies widely across both space and

time in rangeland environments. It is commonly

associated with the accumulation of decomposing litter

under shrubs, but can also occur in interspaces between

plants. In some rangeland environments it can be

extremely important during at least part of the year, while

in others it is scarcely evident.

Q: Do bacteria & fungi play a bigger role in infiltration

rates with black grama (Bouteloua eriopoda) grasslands

that are not degraded when compared to ants and

termites?

A: I am not aware of any studies which have directly

compared the relative contribution of these soil organisms

to soil structure in black grama grasslands. All appear to

be important. Shrub invasion of black grama grasslands

is generally associated with a decline in soil lichen cover.

The lichens are frequently replaced by cyanobacteria,

which are less apparent, but more resilient. Fungal

density is correlated with both live roots and with organic

matter inputs. As the system moves from grass- to shrubdomination, these distribution of these resources become

more patchy at the decimeter to meter scale. Ants and

termites are ubiquitous throughout both degraded and

non-degraded systems.

Q: You described binding agents for various size

aggregates. Are these representative binding agents, in all

ecosystems and if not, what agents occur in different

ecosystems or soil types?

A: The binding agents which I described for various size

classes of aggregates have been found to occur in all

97
ecosystems. Elliott (1986) found that the basic model

described here applies at least in general to a Nebraska

prairie soil. However, the relative importance of each

type of binding agent does vary across both ecosystems

and soil types. Very little information is currently

available, particularly for low-organic matter perennialdominated rangeland soils. A USDA-University of

Illinois study was initiated in 1998 to address this issue

for the Chihuahuan Desert at the Jornada Experimental

Range.

LITERATURE CITED

Allen, E.B. (ed.). 1988. The Reconstruction of

Disturbed Arid Lands. Boulder, Colorado: Westview

Press. 267 p.

Bornemissza, G.F. 1969. A new type of brood care

observed in the dung beetle Oniticellus cinctus

(Scarabaeidae). Pedobiologia 9:223-225.

Chantigny, M.H.; Angers, D.A.; Prévost, D.; Vézina, LP.; Chalifour, F.-P. 1997. Soil aggregation and

fungal and bacterial biomass under annual and

perennial cropping systems. Soil Science Society of

America Journal 61:262-267.

Crawford, C.S.; Mackay, W.P.; Cepeda-Pizarro, J.G.

1993. Detritivores of the Chilean Arid Zone (27-

32°S) and the Namib Desert: a preliminary

comparison. Revista Chilena de Historia Natural 66:

283-289.

Degens, B.P.; Sparling, G.P.; Abbott, L.K. 1996.

Increasing the length of hyphae in a sandy soil

increases the amount of water-stable aggregates.

Applied Soil Ecology 3:149-159.

DePuit, E.J.; Redente, E.F. 1988. Manipulation of

ecosystem dynamics on reconstructed semiarid

lands. In Allen, E.B. ed. The Reconstruction of

Disturbed Arid Lands: An Ecological Approach editor

Edith B. Allen, 162-204. Boulder, Colorado:

Westview Press: 162-204.

Edwards, C.A.; Bohlen, P.J. 1996. Biology and

Ecology of Earthworms. London: Chapman and Hall.

426 p.

Eldridge, D.J. 1993. Effect of ants on sandy soils in

semi-arid eastern Australia: local distribution of nest

entrances and their effect on infiltration of water.

Australian Journal of Soil Research 31:509-518.

Eldridge, D.J. 1994. Nests of ants and termites

influence infiltration in a semi-arid woodland.

Pedobiologia 38:481-492.

Eldridge, D.J.; Greene, R.S.B. 1994. Microbiotic soil

crusts: a review of their roles in soil and ecological

processes in the rangelands of Australia. Australian

Journal of Soil Research 32:389-415.

Eldridge, D.J.; Pickard, J. 1994. Effects of ants on

sandy soils in semi-arid eastern Australia: II.

Relocation of nest entrances and consequences for

bioturbation. Australian Journal of Soil Research 32:

323-333.

Eldridge, D.J.; Tozer, M.E.; Slangen, S. 1997. Soil

hydrology is independent of microphytic crust cover:

further evidence from a wooded semiarid Australian

woodland. Arid Soil Research and Rehabilitation 11:

113-126.

Elkins, N.Z.; Sabol, G.V.; Ward, T.J.; Whitford, W.G.

1986. The influence of subterranean termites on

the hydrological characteristics of a Chihuahuan

Desert ecosystem. Oecologia 68:521-528.

Elliott, E.T. 1986. Aggregate structure and carbon,

nitrogen, and phosphorus in native and cultivated

soils. Soil Science Society of America Journal 50:

627-633.

Hargreaves, G.H.; Samani, Z.A. 1991. WeatherPredictive Models. In Skujins, J. (ed.). Semiarid

Lands and Deserts: Soil Resource and Reclamation.

New York: Marcel Dekker, Inc.: 581-604.

Hassink, J.; Bouwman, L.A.B.; Zwart, K.B.; Brussaard,

L. 1993. Relationships between habitable pore

space, soil biota and mineralization rates in

grassland soils. Soil Biology and Biochemistry 25:

47-55.

Herrick, J.E.; Havstad, K.M.; Coffin, D.P. 1997.

Rethinking remediation technologies for desertified

landscapes. Journal of Soil and Water Conservation

52:220-225.

Herrick, J.E.; Wander, M.M. 1998. Relationships

between soil organic carbon and soil quality in

cropped and rangeland soils: the importance of

distribution, composition and soil biological activity.

In: Lal, R.; Kimble, J.; Follett, R.; Stewart, B.A.

(eds.). Advances in Soil Science-Soil Processes and

the Carbon Cycle, editors. Boca Raton, Florida: CRC

Press: 405-425.

Herrick, J.E.; Lal, R. 1995. Evolution of soil physical

properties during dung decomposition in a tropical

pasture. Soil Science Society of America Journal

59:908-912.

Hindell, R.P.; McKenzie, B.M.; Tisdall, J.M.; Silvapulle,

M.J. 1994. Relationships between casts of

98
geophagous earthworms (Lumbricidae, Oligochaeta)

and matric potential I. Cast production, water

content, and bulk density. Biology and Fertility of

Soils 18:119-126.

Hudson, B.D. 1994. Soil organic matter and water

holding capacity. Journal of Soil and Water

Conservation 49:189-194.

Ladyman, J.A.R.; Muldavin, E. 1996. Terrestrial

cryptogams of pinyon-juniper woodlands in the

Southwestern United States: a review. Gen. Tech.

Rep. RM-GTR-280. U.S. Department of Agriculture,

Forest Service, Rocky Mountain Forest and Range

Experiment Station. 33 p.

Laundr6, J.W. 1990. Soil moisture patterns below

mounds of harvester ants. Journal of Range

Management 43:10-12.

Lee, K.E.; Foster, R.C. 1991. Soil fauna and soil

structure. Australian Journal of Soil Research 29:

745-775.

Lee, K.E.; Smettem, K.R.J. 1995. Identification and

manipulation of soil biopores for the management of

subsoil problems. In Jayawardane, N.S.; Stewart,

B.A., eds. Subsoil Management Techniques. Boca

Raton, Florida: Lewis Publishers: 211-243.

Lobry de Bruyn, L.A.; Conacher, A.J. 1990. The role

of termites and ants in soil modification: a review.

Australian Journal of Soil Research 28:55-93.

Lobry de Bruyn, L.A.; Conacher, A.J. 1994. The

effect of ant biopores on water infiltration in soils in

undisturbed bushland and in farmland in a semi-arid

environment. Pedobiologia 38:193-207.

Ludwig, J.A.; Tongway, D.J. 1995. Spatial

organization of landscapes and its function in semiarid woodlands, Australia. Landscape Ecology 10:

51-63.

Mando, A. 1997. The impact of termites and mulch

on the water balance of crusted Sahelian soil. Soil

Technology 11:121-138.

Mando, A.; Miedema, R. 1997. Termite-induced

changes in soil structure after mulching degraded

(crusted) soil in the Sahel. Applied Soil Ecology 4:

241-249.

Marshall, T.J.; Holmes, J.W.; Rose, C.W. 1996. Soil

Physics, 3d ed. Camridge, England: Cambridge

University Press: 453 p.

Naeem, S. 1998. Species redundancy and ecosystem

reliability. Conservation Biology 12:39-45.

NRCS. 1997. Introduction to Microbiotic Crusts. U.S.

Department of Agriculture, Natural Resources

Conservation Service, Soil Quality Institute and

Grazinglands Technology Institute. 13 p.

Oades, J.M.; Waters, A.G. 1991. Aggregate hierarchy

in soils. Australian Journal of Soil Research 29:

815-828.

Roldán, A.; Albaladejo, J.; Thornes, J.B. 1996.

Aggregate stability changes in a semiarid soil after

treatment with different organic amendments. Arid

Soil Research and Rehabilitation 10:139-148.

Schlesinger, W.H.; Reynolds, J.F.; Cunningham, G.L.;

Huenneke, L.F.; Jarrell, W.M.; Virginia, R.A.;

Whitford, W.G. 1990. Biological feedbacks in

global desertification. Science 247:1043-1048.

Soil Science Society of America. 1997. Glossary of

Soil Science Terms. Madison, Wisconsin: Soil

Science Society of America.

Spaeth, K.E.; Pierson, F.B.; Weltz, M.A.; Awang, J.B.

1996. Gradient analysis of infiltration and

environmental variables as related to rangeland

vegetation. Transactions of the American Society of

Agricultural Engineers 39:67-77.

Spears, B.M.; Ueckert, D.N.; Whigham, T.L. 1975.

Desert termite control in a shortgrass prairie: effect

on soil physical properties. Environmental

Entomology 4:899-904.

Tisdall, J.M. 1991. Fungal hyphae and structural

stability of soil. Australian Journal of Soil Research

29:729-743.

Tisdall, J.M.; Oades, J.M. 1982. Organic matter and

water-stable aggregates in soils. Journal of Soil

Science 33:141-163.

Tongway, D.J.; Ludwig, J.A. 1994. Small-scale

resource heterogeneity in semi-arid landscapes.

Pacific Conservation Biology 1:201-208.

Trojan, M.D.; Linden, D.R. 1992. Microrelief and

rainfall effects on water annd solute movement in

earthworm burrows. Soil Science Society of

America Journal 56:727-733.

Warren, S.D. 1995. Ecological role of microphytic soil

crusts in arid ecosystems. In Allsop, D.; Colwell,

R.R.; Hawksworth, D.L. (eds.). Microbial Diversity

and Ecosystem Function. London: CAB

International: 199-209.

West, N.E. 1990. Structure and function of

microphytic soil crusts in wildland ecosystems of

99
arid to semi-arid regions. Advances in Ecological

Research 20:179-21?

Whisenant, Steven G. 1996. Initiating autogenic

restoration on degraded arid lands. In: Neil E. West,

ed. Proceedings of the Fifth International Rangeland

Congress. Salt Lake City, Utah, 23 July 1995-28

July 1995. Denver, Colorado: Society for Range

Management: 597-598.

Whitford, W.G.; DiMarco, R. 1995. Variability in soils

and vegetation associated with harvester ant.

100

Pogonomyrex rugosus, nests on a Chihuahuan

Desert watershed. Biology and Fertility of Soils 20:

169-173.

Whitford, W.G. 1993. Animal feedbacks in

desertification: a review. Revista Chilena de Historia

Natural 66:243-251.

Whitford, W.G. 1996. The importance of the

biodiversity of soil biota in arid ecosystems.

Biodiversity and Conservation 5:185-195.
EFFECTS OF DISTURBANCE ON MICROBIAL ACTIVITY AND N-CYCLING IN FOREST

AND SHRUBLAND ECOSYSTEMS

John M. Stark1 and Stephen C. Hart2

ABSTRACT

Disturbances involving removal or destruction of plant biomass or mixing of soil typically increase nitrogen (N)

availability. Increased N availability above plant requirements can have detrimental effects by increasing soil nitrate

concentrations, which can lead to increased N losses or increased abundance of undesireable weedy plant species. The

increase in nitrate concentrations following disturbance can be attributed in part to increases in populations and activity

of nitrifying bacteria, but an even more important cause appears to be a reduction in rates of assimilation of nitrate by

soil microorganisms. This decline in microbial assimilation rates appears to be direct result of smaller amounts of

belowground plant carbon-inputs following disturbance. Maintaining C-inputs to the soil following disturbance will

allow continued microbial assimilation of N, which will minimize the adverse affects associated with excess N

availability.

Keywords: disturbance, nitrification, microbial assimilation of nitrogen, nitrate, isotope dilution, nitrogen loss

INTRODUCTION

The term "disturbance" is used frequently in ecological

discussions, however, the term is quite ambiguous and

may describe events that have very different effects on

ecosystem processes. In addition, the degree of

disturbance that a particular event creates in an

ecosystem may be quite specific to that ecosystem.

For example, a period of two or three months of

drought may represent a disturbance to plants adapted

to mesic environments, while that same drought might

have little or no impact on plants adapted to semi-arid

and arid systems. Likewise, a moderate amount of

herbivory might have minimal effects on a plant species

that has evolved in the presence of grazing, but could

be devastating to a plant that has not. In this paper,

we will limit our discussion to two types of

disturbances that have direct effects on ecosystem

function, soil microbial communities, and nitrogen (N)

cycling. These disturbances are: a) destruction or

removal of plant biomass (during processes such as

fire, grazing, harvesting, etc); and b) physical mixing of

soil (during processes such as cultivation, and forest

harvesting and site preparation).

These two types of disturbances may have a number of

different effects on N cycling. Removal of plant

biomass reduces water extraction, uptake of N, and

most importantly, reduces the input of carbon (C) to

the soil system. All of these factors will tend to

promote increased N availability: higher soil moisture

contents will lead to greater microbial activity, but with

reduced C-inputs soil microorganisms will produce less

biomass, and thus, immobilize less N. In addition,

since the plants themselves are taking up less N, there

will be more left in the available pool.

Soil mixing can increase substrate availability to soil

microorganisms by exposing organic materials that

were previously trapped in inaccessible locations. Soil

mixing may also increase microbial activity by changing

soil aeration and water relations. In addition, soil

mixing may disrupt fungal hyphae that link soil

microsites. Disruption of these fungal linkages reduces

the ability of the microbial community to exploit

microsites of high N availability, which is likely to

increase N availability to plants. Therefore, one of the

important outcomes of many disturbances is that there

is a temporary increase in availability of N, as well as a

variety of other nutrients.

CHANGES IN NITROGEN AVAILABILITY

Increased N availability is in many respects a double

edged sword. It may be something to strive for if you

are interested in growing highly productive agricultural

crops, but where sustained production of more slowgrowing native perennial vegetation is the goal, N

availability in excess of plant needs may have adverse

1

John M. Stark is an associate professor of microbial ecology in the Department of Biology, Utah State University, Logan, UT

84322-5305.

2Stephen C. Hart is an associate professor of forest ecosystem ecology at the School of Forestry, College of Ecosystem Science

and Management, Northern Arizona University, Flagstaff, AZ 8601 1-5018.

101
effects. Increased availability of N generally results in

greater accumulation of nitrate, the primary anion form

of soil inorganic N. This may cause problems for two

reasons: nitrate is the form of N that is most readily

lost from soils, and high nitrate concentrations may

stimulate growth of annual weeds.

Increases in the availability of either form of inorganic

N, ammonium (NH4

 + ) or nitrate (NO3-

) may cause

increased N loss from soil, but the most important

pathways for N-loss from most soils are associated

with the NO3

-

 pool (fig.1). Denitrification occurs under

anaerobic conditions, as bacteria rely on NO3

-

 instead

of O2 to carry out respiration processes. During

denitrification, NO3

-

 is converted to gaseous forms of N

which are lost to the atmosphere. High soil NO3

-

concentrations can also lead to increased losses from

leaching. Because N03

-

 is an anion, it is not adsorbed

to soil colloids as readily as NH4

 +

 , and thus it is easily

washed out as water percolates through the soil.

ground water

Figure 1 —Pathways for nitrogen loss from soils.

High N availability may also promote growth of fast

growing weedy plant species at the expense of the

native perennials. Annual weeds frequently show a

greater growth response to high N-availability than

many slow-growing native perennials (Kay and Evans

1965, Trent and others 1994, McLendon and Redente

1994). In addition, certain annual weeds show a

preference for NO3

-

 as an N-source, whereas many

native perennials show a preference for NH4

+

 (Young

and others 1995). Therefore the shift away from

NH4

 + -dominated soils and toward NO3-

-dominated soils

that usually accompanies soil disturbance, may favor

establishment of annuals rather than the more desirable

perennial species. It has been suggested, for example,

that changes in N-availability and in the abundance of

NO3

-

 relative to NH4

 +

 , play important roles in allowing

cheatgrass stands to persist in much of the western

US.

Because many types of disturbance result in

accumulations of NO3

-

 in soil, much of the research

related to disturbance effects have focused on the

process of NO3

-

 formation (Robertson and Vitousek

1981). Nitrate is formed from NH4

+

 by a process

called nitrification (fig. 2). This process is carried out

by two groups of soil bacteria: the first group oxidizes

NH4

+

 to nitrite (NO2

-

), and consists primarily of

members of the genera Nitrosomonas, Nitrosospira, and

Nitrosolobus. These bacteria get 100% of their energy

from the oxidation of NH4

+

, and to create biomass they

fix CO2. Thus, they are quite responsive to changes in

the availability of NH4

 +

 . The second step of

nitrification, the oxidation of NO2

-

 to NO3

-

, occurs quite

rapidly and is carried out by soil bacteria belonging to

the genus Nitrobacter.

Figure 2 —The two-step process of nitrification and the

genera of bacteria that carry out the processes.

N-CYCLING-PAST AND CURRENT

VIEWS

Previous research on the process of nitrification has led

to the view that in soils of undisturbed late

successional ecosystems, very little NO3

-

 is produced,

and what is produced is taken up by plants (fig. 3a).

The vast majority of the N-cycling has been believed to

occur between the organic and NH4

+

 pools. With low

rates of nitrification, and plant uptake of the small

amount of NO3

-

 that is produced, N-losses are kept to a

minimum.

Another conventional view of the N-cycle has been that

soil microorganisms are not important sinks for N03

-

(Jones and Richards 1977). It has been shown in

laboratory studies that the presence of NH4

+

 inhibits

microbial uptake of NO3

-

 (Broadbent and Tyler 1962).

Since NH4

+

 is almost always present in soils, it has

been assumed that this will prevent significant amounts

of uptake of NO3

-

 by soil microorganisms, and thus,

plants will be the primary sink for NO3

-

. Upon

disturbance, such as removal of plant biomass, plant

uptake of ammonium stops (or is greatly reduced), and

NH4

+

 pools increase (fig. 3b). Increased NH4

 +

availability allows nitrification to occur, and populations

of nitrifying bacteria expand, causing even more NO3

-

to be produced. Because plants are either gone or

present in only small amounts, there is no sink for NO3

-

,

and NO3

-

 accumulates. Under these conditions

increased denitrification and leaching cause substantial

amounts of N to be lost from the soil. As mentioned

102
Figure 3-Revised view of the N-cycle in a) late successional

ecosystems; and b) early successional ecosystems following

disturbance. Even in mature undisturbed ecosystems, a

substantial amount of N appears to cycle through the nitrate

pool, but microbial assimilation prevents nitrate from

accumulating and leading to N-loss.

earlier, higher soil water contents also may follow plant

biomass removal due to reduced transpiration, which

will increase the likelihood of anaerobic conditions and

water percolation, and thus further increase N loss by

denitrification and leaching.

Recent studies have provided evidence that this

conventional view of N-cycling is not quite correct.

The phenomenon of increased N-availability and NO3

-

leaching following disturbance has been well

documented, but the mechanisms for the phenomenon

have not. Almost all of the work showing that

nitrification rates are slow or nonexistent in mature

undisturbed forests have utilized measures of net

nitrification rates, such as the buried bag technique. In

this technique, a soil sample is collected and incubated

in the field. The net change in the soil NO3

-

 pool during

the incubation period is used to estimate nitrification

rates (fig. 4). If all of the consumption processes are

eliminated, then this net change in the NO3

-

 pool will

equal the gross rate of nitrification. However, if there

is consumption occurring, then the net rate will

underestimate the actual gross rate of nitrification.

Recent studies using

 15N have shown that microbial

assimilation of NO3

-

 may be much more prevalent than

previously thought, and thus net nitrification rates may

Figure 4—Sources and sinks of nitrate in soils. Net

nitrification techniques measure the difference between

nitrate production and consumption. If consumption

processes are significant, then net nitrification rates will

underestimate gross nitrification rates. In most undisturbed

soils, microbial assimilation of nitrate is a substantial

consumptive fate of nitrate.

frequently underestimate gross nitrification rates

(Schimel and others 1989, Davidson and others 1990,

Davidson and others 1992).

We decided to use a relatively new isotope dilution

technique (Hart and others 1994) to evaluate whether

or not nitrification rates are low in mature undisturbed

forests and shrublands. We used this technique to

measure nitrification and NO3

-

 consumption in soils of a

wide variety of undisturbed forests (Stark and Hart

1997). These sites ranged from the highly productive

western hemlock-sitka spruce forests on the central

coast of Oregon to a juniper woodland in central

Oregon. We also measured rates at a big sagebrush

steppe in northern Utah and at sites along an

elevational gradient in northern New Mexico, ranging

from a subalpine spruce-fir site at 3,400 m to a pinyonjuniper woodland at 2,400 m.

Consistent with previous observations, net nitrification

rates were low at virtually all of the sites (Stark and

Hart 1997). In fact, in 13 of 19 cases, rates were

actually less than zero. In contrast, gross nitrification

rates, measured by isotope dilution, were quite high.

These rates ranged from about one to two orders of

magnitude faster than estimates of plant uptake,

indicating that cycling of NO3

-

 through the soil microbial

population is extremely rapid, even in soils of mature

undisturbed forests. Because of the high nitrification

rates and relatively small NO3

-

 pool sizes, soil NO3

-

pools turned over about once each day.

Another interesting observation was that when

measurements were made in intact soil cores, NO3

-

consumption was very tightly coupled to nitrification.

103
The correlation between NO3

-

 consumption and

nitrification followed a nearly 1:1 relationship and had

an r2 of 0.89 (Stark and Hart 1997). In other words,

virtually all of the NO3

-

 that was produced in

undisturbed soil samples was rapidly consumed by soil

microorganisms. This result contradicts the previously

held view and shows that nitrification rates can be

quite high in soils of undisturbed forests. In addition, it

shows that it is not plants, but soil microorganisms

that are primarily responsible for promoting NO3

-

retention in undisturbed ecosystems.

The tight coupling between NO3

-

 consumption and

production that was seen in intact cores was destroyed

when the soil samples were physically disturbed.

Following disturbance, the r2 for the correlation

between NO3

-

 consumption and nitrification dropped

from 0.89 to 0.38. One explanation for this change

may be that in undisturbed soils, microsites with high

rates of microbial immobilization are linked to

microsites with high rates of mineralization and

nitrification by fungal hyphae. Through these fungal

linkages, N released in high-mineralization microsites is

transported back to the high immobilization microsites

where it is consumed. Physical disturbance of the soil

would break these hyphal connections, and uncouple

NO3

-

 consumption from production. Therefore,

immobilization would stop due to N-limitations, but Nmineralization and nitrification would continue.

This work suggests that we need to revise our view of

the N-cycle in undisturbed systems, and our view of

how disturbance affects changes in N-availability (fig.

5a,b). Nitrification appears to be a significant part of

the N cycle even in undisturbed forest soils; however,

nitrification does not necessarily result in high rates of

N loss because microbial assimilation of NO3

-

 prevents

NO3

-

 accumulation. Following disturbances that result

in removal of plant biomass, belowground inputs of

plant C will slow or stop all together, which causes

microbial assimilation of N to stop. Without microbial

assimilation or plant uptake, NO3

-

 accumulates, and

increased denitrification and leaching result in

substantial N loss.

MANAGEMENT IMPLICATIONS

The management implications of this work are relatively

simple: because rates of microbial uptake of N are

regulated by carbon availability, maintaining the carbon

supply belowground will minimize the adverse effects

of disturbance on the N-cycle. Therefore,

management practices that maximize the amount of

organic debris left on-site and minimize the period of

time between vegetation removal and reestablishment

will minimize N-losses and limit invasion by weedy

plant species.

Figure 5 —The conventional view of the N-cycle in a) late

successional ecosystems; and b) early successional

ecosystems following disturbance.

LITERATURE CITED

Broadbent, F.E.; Tyler, K.B. 1962. Laboratory and

greenhouse investigations of nitrogen immobilization. Proceedings Soil Science Society of America

26:459-462.

Davidson, E.A.; Stark, J.M.; Firestone, M.K. 1990.

Microbial production and consumption of nitrate in

an annual grassland. Ecology 71:1968-1975.

Davidson, E.A.; Hart, S.C.; Firestone, M.K. 1992.

Internal cycling of nitrate in soils of a mature

coniferous forest. Ecology 73:1148-1156.

Hart, S.C.; Stark, J.M.; Davidson, E.A.; Firestone, M.K.

1994. Mineralization, immobilization, and

nitrification. Pp. 985-1018 in Weaver, R.; and

others (eds.). Methods of Soil Analysis. Part 2,

Microbiological and Biochemical Properties. Soil Sci.

Soc. Am., Madison, Wl.

Jones, J.M.; Richards, B.N. 1977. Effect of

reforestation on turnover of

 15N-labelled nitrate and

ammonium in relation to changes in soil microflora.

Soil Biology & Biochemistry 9:383-392.

104
Kay, B.; Evans, R. 1965. Effects of fertilization on a

mixed-stand of cheatgrass and intermediate

wheatgrass. Journal of Range Management 18:7-

11.

McLendon, T.; Redente, E.F. 1994. Role of nitrogen

availability in the transition from annual-dominated

to perennial-dominated serai communities. Pp. 352-

362 in Monsen, S.B.; Kitchen, S.G. (eds.).

Proceedings: Ecology and management of annual

rangelands. Intermountain Research Station,

Ogden, UT.

Robertson, G.P.; Vitousek, P.M. 1981. Nitrification

potentials in primary and secondary succession.

Ecology 62:376-386.

Schimel, J.P.; Jackson, L.E.; Firestone, M.K. 1989.

Spatial and temporal effects on Plant-microbial

competition for inorganic nitrogen in a California

grassland. Soil Biology & Biochemistry 21:1059-

1066.

Stark, J.M.; Hart, S.C. 1997. High rates of

nitrification and nitrate turnover in undisturbed

coniferous forests. Nature 385:61-64.

Trent, J.D.; Young, J.A.; Blank, R.R. 1994. Potential

role of soil microorganisms in medusahead invasion.

Pp. 140-142 in Monsen, S.B.; Kitchen, S.G. (eds.).

Proceedings: Ecology and management of annual

rangelands. Intermountain Research Station,

Ogden, UT.

Young, J.A.; Blank, R.R.; Longland, W.R. 1995.

Nitrogen enrichment-immobilization to control

succession in arid land plant communities. Journal

of Arid Land Studies. 5S:57-60.

105
NITROGEN TRANSFORMATIONS AND MANAGEMENT INTERACTIONS IN FORESTS

David D. Myrold1

ABSTRACT

Nitrogen is often the most limiting nutrient, and often the most limiting of all environmental constraints, to the

productivity of forests. The N cycle is also one of the most complicated of all elemental cycles in terrestrial

ecosystems. Consequently, it is important to understand its regulation in order to understand the effect that forest

management practices have on N cycling and sustainable productivity of forests. Harvesting and site preparation result

in an immediate removal of N from the site and can result in elevated losses of N for a period of time as a result of

increased N turnover. The long-term impact of these management practices depends upon the size of the remaining soil

N pool, how quickly it takes for N consumptive processes to come back into balance with the enhanced production of

NH4

+

 and NO3

-

, and an ameliorative efforts to add N to the site through fertilization or use of N2-fixing plant symbioses.

This recovery of the N cycle is site dependent, of course, but can likely be achieved on many sites by carefully

balancing N inputs with N removals and using the principles that regulate N cycling to design management strategies

that minimize N losses and time to recovery.

Keywords: nitrogen cycle, nitrogen fixation, mineralization, immobilization, nitrification, denitrification

INTRODUCTION

Of all plant nutrients, N is needed in the largest amount

and is the nutrient most often limiting productivity of

terrestrial ecosystems. Consequently, management

activities that alter the amount or availability of N can

greatly alter ecosystem productivity, over both the

short- and long-term. Most often, forest management

practices have the potential for decreasing the N

supply, which can result in lower growth and

potentially alter forest composition; however, some

activities can lead to an excess of N, which may have

detrimental environmental consequences.

In this paper I will review the basic principles of N

cycling in forest soils and demonstrate how these

principles can help us understand the impact that forest

management practices are likely to have. In doing so, I

will draw primarily on examples of research done in the

Pacific Northwest.

OVERVIEW OF N CYCLE PROCESSES

To help understand the cycling of N, it is useful to

think in terms of pools and processes. A forest

ecosystem has several distinguishable pools of N, for

example, N stored in living plants, soil organic N, NH4

and NO3

-

 in soil solution. It is interesting to note that

the pools of N that can be readily used by plants are

the smallest and those that are least available are the

largest (fig. 1). Because the pools of available N are

small, the balance between productive and

consumptive processes that affect these pools-the

inputs and outputs-are extremely important. To aid in

conceptualizing the N cycle, it is useful to separate N

processes into inputs and outputs to the system from

those transformations that are internal to the plant-soil

system (fig. 2).

Figure 1 —Sizes of N pools typical for a 30-m Douglas-fir

stand in the Pacific Northwest. The N2 pool is based on the

N2 contained in the top 1 m of soil plus that in the

atmosphere within the canopy of the 30-m trees. Soil

organic N and plant N are based on Biggar and Cole (1983)

and Cole and Johnson (1981). Soil inorganic N pools are

typical for conifer forests of the Pacific Northwest (for

example, Binkley and others 1992, Vermes and Myrold

1992).

1David D. Myrold is a professor of forest soils and soil microbiology, Oregon State University, Corvallis, OR 97331-7306.

106
Figure 2—Major pools and processes involved in N cycling in

forest ecosystems.

Inputs

Inputs of N into forested ecosystems can be natural or

anthropogenic in origin. Natural inputs include

precipitation and biological N2 fixation. Both of these

can, of course, be altered or manipulated by human

activity. Fertilization is the primary management

practice that directly manipulates the input of N to

forests.

Precipitation-The most prevalent natural input is

inorganic N from precipitation and dry deposition. In

pristine forests far from sources of atmospheric

pollution, atmospheric inputs of N are low; no more

than a few kg N ha-1 annually in the Pacific Northwest

(Cole and Johnson 1981). Forests located downwind

from heavily industrialized regions, high traffic areas, or

animal stockyards can have inputs in excess of 10 kg

ha-1 yr-1 (Myrold and Nason 1992).

Nitrogen fixation-The conversion of atmospheric N2

into plant available forms by bacteria is the second

major natural input of N. In most instances, the

amount of N accrued by N2 fixation in forest

ecosystems is small, a few kg ha-1 yr-1 (Dawson 1983).

More substantial inputs can occur when plants with

root-nodulating, N2-fixing symbioses are present. In the

Pacific Northwest, N2 fixation rates can exceed 10 kg

ha-1 yr-1 on sites with Ceanothus and 50 kg ha-1 yr-1 on

sites with alder (Hibbs and Cromack 1990). It is

possible to manage these N2-fixing plant symbioses to

manipulate N inputs at some sites.

Internal Cycling

Within a forest stand, the conversion of N from one

form to another is highly dynamic and the major pools

of N are interconnected by several biological processes

(fig. 2). The relative rates of these concurrent

processes determine the availability of N, are generally

tightly regulated, and can be influenced by

management.

Ammonification-Heterotrophic soil organisms can

produce NH4

+

 as they decompose plant material and

mineralize soil organic matter. This conversion of

organic N to NH4

+

 is known as ammonification.

Whether ammonification results in the net accumulation

of NH4

+

 depends upon the N requirements of the

heterotrophic organism relative to the concentration of

N in the organic substrate. The general rule of thumb

is that NH4

+

 will be produced from an organic substrate

with a C:N ratio of about 25 or less; higher C:N ratios

will result in inorganic N being immobilized by

heterotrophic organisms.

Immobilization-When heterotrophic organisms are

limited by available N, they assimilate, or immobilize,

inorganic forms of N. Assimilation of NH4

+

 is generally

considered to be the dominate route of immobilization,

however, recent work in forest soils suggest that

sizeable amounts of NO3

-

 can also be immobilized (Hart

and others 1994). This finding may alter the

commonly held conception that NO3

-

 production is

insignificant in many forest soils.

Nitrification-The production of NO3

-

 is key factor in

determining N losses from forest ecosystems. Nitrate

can be produced by the oxidation of NH4

+

 by

specialized genera of nitrifying bacteria, a process

known as autotrophic nitrification, or from organic N by

a range of heterotrophic microorganisms, a process

known as heterotrophic nitrification. Although the

autotrophic process is generally considered to be the

most important in soils, recent work has shown that

heterotrophic nitrification can be dominant in at least

some forest soils (Hart and others 1997).

Plant uptake-Plants take up both NH4

+

 and NO3

-

 to

meet their N requirements. In a forest stand, both

ground vegetation and trees take up inorganic N,

however, after crown closure, trees are responsible for

most plant N uptake. The amount of N taken up by

forest trees varies with site productivity (the growth

rate of trees, and hence N uptake rates are often

dependent upon soil N availability), but typical values

for Douglas-fir in the Pacific Northwest are about 20-

40 kg ha-1 yr-1 (Cole and Gessel 1992).

Outputs

Nitrogen leaves forest ecosystems in various forms and

by several natural and anthropogenic mechanisms.

Management activities can directly remove N through

harvest and the use of prescribed fire as well as more

indirectly by altering other, more continual outputs,

such as erosion, leaching, and denitrification. Although

107
wildfire is a natural component of many forest

ecosystems and can result in catastrophic losses of N,

most natural outputs of N are continual and generally

small.

Leaching-Because it is an anion, NO3

-

 is often the

dominant form of N lost by leaching, although losses of

dissolved organic N can be of a similar magnitude. In

general, undisturbed forest ecosystems retain N very

tightly and little is lost below the rooting zone.

Coniferous forests of the Pacific Northwest loose only

a few kg ha-1 yr-1 through leaching (Sollins and others

1980), although rates can be higher in high productivity

coastal forests (Binkley and others 1992). Leaching

losses >20 kg ha-1 yr-1 are not uncommon in stands

dominated by red alder (Binkley and others 1992).

Denitrification-Undisturbed forest ecosystems lose

small amounts (<1 kg ha-1 yr-1) of N by denitrification

(Davidson and others 1990). Denitrification losses are

somewhat higher in the presence of N2-fixing plants or

following disturbance (Binkley and others 1992,

Vermes and Myrold 1992).

Regulation of Nitrogen Cycling

To better understand N cycling and predict the

influence of management practices on N cycling it is

important to consider the factors that control N cycling

in soils. This task is simplified if we divide the internal

N cycle into three subcycles: the organic N, NH4

 +

 , and

NO3

-

 subcycles (fig. 3).

Organic N subcycle-The major input into the soil

organic N pool is the physical incorporation and

biological decomposition of plant detritus (fig. 3a).

Except when N2-fixing plant symbioses are present, the

N in plant debris represents N that is recycling through

the plant-soil system. The bulk of N in plant debris is

present in organic forms, largely proteinaceous

compounds, which are metabolized by heterotrophic

soil organisms. Some of this N may be incorporated

into microbial biomass, a relatively small pool of organic

N that turns over several times per year; some may be

transformed into more recalcitrant compounds or

interact with soil minerals and thereby be stored in a

more stable soil organic N pool. Small, water soluble

organic N compounds, such as amino acids, that are

by-products of mineralization may leach out of the soil.

The mineralization process can also produce NH4

+

 by

ammonification when the C:N ratio of the decomposing

plant debris is low enough to provide the decomposers

with all the N they require for growth.

Ammonium subcycle-Ammonium produced by

biological ammonification, or entering the system as an

outside input (for example, precipitation, fertilization),

has a variety of fates (fig. 3b). At high pH (above

8), NH4

+

 can be lost by volatilization to NH3 gas, but

this is highly localized to microsites immediately

surrounding animal excreta or urea fertilizer granuales.

Because it is positively charged, NH4

+

 is held on cation

exchange sites and is not readily leachable. There can

also be some chemical fixation of NH4

+

 by illitic clays

or soil organic matter, but this is generally a minor fate

in most forest soils. Assimilation of NH4

+

 by

108

Figure 3c

Figure 3 —The three subcycles of the N cycle: a. soil organic

matter subcycle, b. NH4

+

 subcycle, c. NO3

-

 subcycle.
heterotrophic microorganisms and plants is a more

important fate. These organisms compete with

nitrifying bacteria for NH4

+

. In many forest soils, C

availability is high, consequently heterotrophic

microorganisms out-compete plants and nitrifiers,

making NH4

+

 immobilization the dominant fate and

resulting in plant growth being N limited and relatively

low rates of NO3

-

 production.

Nitrate subcycle-Nitrate is often the smallest pool of

N in the internal N cycle. It is produced biologically by

nitrification and can enter the system in precipitation or

through fertilization (fig. 3c). Because it is negatively

charged, NO3

-

 does not interact strongly with soil

organic matter or soil minerals and is subject to

leaching. In vegetated soils much of the NO3

-

 is taken

up by plants, however. Heterotrophic microorganisms

can also assimilate NO3

-

, and recent research has

shown that this can be a major fate of NO3

-

 in C-rich

forest soils. Some reduction of NO3

-

 to N2 by

denitrification can occur in anaerobic microsites of

forest soils, but this is usually a minor fate of NO3

-

.

Guiding Principles

Although the exact impacts of forest management

activities will always be site specific, there are a few

principles that allow us to generalize about the

outcome of these activities. First, it is important to

remember that N cycling is dominantly biological. Thus

the effect of management activities will be determined

by how these activities alter the major regulators of

biological activity: substrate supply (both quantity and

quality), water availability, temperature, etc. Second,

with the exception of autotrophic processes (plant

growth and nitrification), the cycling of N is intimately

linked with C cycling (most of the N in an ecosystem is

in organic compounds). This close linkage between C

and N means that it is the heterotrophic need for C that

often drives the cycling of N and that storage of N is

mainly linked to C storage. Third, plant, and even

microbial, pools of available N are often in short supply,

thus competition between plants, heterotrophic

microorganisms, and nitrifying bacteria is the norm.

Another way of stating this principle is that competition

for available N keeps the sizes of the available N pools

small. Four, the turnover rates of N pool are inversely

related to their size; small pools, such as NH4

 + and NO3-

, turn over very rapidly. Thus, pool size is not

necessarily an indicator of biological activity.

MANAGEMENT IMPACTS

Forest management can affect inputs and removals of

N as well as influencing the processes of the internal N

cycle. In each of these cases, the effect of

management can be direct, an immediate response, or

indirect, a delayed response. Furthermore, the degree

of the effect can vary widely from imperceptible to

catastrophic. More subtle effects come about by

alterations in soil physical and chemical properties or

changes in the microenvironment.

Nitrogen Removals

Nitrogen in forest ecosystems is removed by, or lost as

a result of, various management practices. Direct

removal of N occurs during timber harvest and the use

of prescribed fire. Forest operations can also have less

immediate but longer lasting effects through altering

rates of erosion, leaching, or denitrification.

Harvesting-The immediate impact of logging is to

remove boles, and potentially other portions, of trees.

Although significant, the removal of boles represents

only about 5-20% of the N stored on forested sites

(fig. 4). About an equal amount of N is stored in

branches and foliage, therefore complete removal of

aboveground tree biomass is a loss of 10-35% of the

total system N (Biggar and Cole 1983). The fraction of

system N removed in boles is fairly constant for a given

species regardless of site productivity, however,

relatively more N is removed from low compared to

high productivity sites when branches and foliage are

harvested. By itself, the amount of N removed by

harvesting is not usually large enough to adversely

affect the long-term productivity of a site, except

possibly on sites with very small reserves of N in the

soil.

Figure 4 —Distribution of N in a Douglas-fir stand (Biggar and

Cole, 1983).

As mentioned previously, the major indirect effect of

removing trees is to reduce the competition for

available N, at least until vegetation regrows. Thus, it

is common to find greater rates of net N mineralization

and enhanced NO3

-

 production during the first few

years after harvesting (table 1). This is partly from

reduced competition for N but may also be fostered by

an improved microclimate (moister and warmer soils) as

a result of removing the forest canopy.

Harvesting operations not only remove trees but usually

result in exposure of mineral soil, soil displacement, and

109
Table 1-Indirect effects of harvesting on N lost by denitrification and on internal cycling of N by mineralization and

nitrification. The Douglas-fir site is on the east side of the central Coast Range in Oregon; the western hemlock

site is on the west side of the Coast Range in southwestern Washington.

Site

Douglas-fir

Western hemlock

Treatment

Old-growth

Harvested only

Harvested/compacted

Old-growth

Harvested only

Harvested/compacted

Net N mineralization

- kg N ha-1 y-1 --

<10

70

40

<10

120

200

Net nitrification

-- % -

0

60

70

0

65

50

Denitrification

-- kg N ha-1 y-1 --

<0.1

2

4

0

0.3

0.4

compaction. These changes can also alter N cycling in

less direct, but in some instances, longer lasting ways.

Decreased vegetative cover and soil exposure can

accelerate losses of N through erosion, leaching, and

denitrification, at least until there is sufficient regrowth

to restore pre-harvest N uptake rates.

A review of studies in several western Oregon

watersheds showed that the effect of logging on

suspended sediment load in streams is highly variable

(Fredriksen and Harr 1981): about half of these studies

showed no significant change; the remainder showed

2- to 25-fold increases in sediment load following

harvest operations and continued elevated levels of

sediment for five or more years. Greater rates of

sedimentation were associated with steeper slopes.

Disturbance caused by harvesting operations often

causes an increase in NO3

-

 leaching (Vitousek and

others 1979), although the intensity and duration of

this response is highly variable. For example, a study

on the H. J. Andrews Experimental Forest in the

Oregon Cascades showed that the effect of

clearcutting and burning lasted at least six years,

although streamwater NO3

-

 concentrations were only

greatly elevated in the year following slash burning and

annual losses at this site were small (Fredriksen and

others 1975).

The soil disturbance and compaction that often

accompanies harvesting operations can also influence

the internal N cycle. Soil mixing may cause a

temporary increase in net N mineralization, analogous

to the effect of plowing in agriculture. Compaction, on

the other hand, by decreasing macropore space and

thereby inhibiting aeration, can sometimes slow aerobic

microbial activity (table 1). This change in aeration

may also enhance denitrification losses on some sites

(table 1) but are relatively small in forests of the Pacific

Northwest (Vermes and Myrold 1992).

Fire-Prescribed fire has often been used in the past

following harvesting for hazard reduction or to prepare

the soil for regeneration. The amount of site N lost

from burning depends upon the intensity of the fire.

Light burns may remove only the small fuel portion of

the slash and only the litter (L or Oi) layer of the forest

floor; more severe burns may remove all of the slash

and forest floor and even volatilize some N from the

mineral soil. As illustrated in fig. 4, even a hot burn

removes only about the same amount of N as is

removed in bole wood.

Because fire is often used in conjunction with timber

harvesting practices, it is often difficult to separate its

influence from those of tree removal and physical

disturbance on N losses by erosion, leaching, and

denitrification. Elevated concentrations of NH4

+

 and

NO3

-

 are often found during the first few years after

burning (Weetman and others 1990), which

corresponds to the period of enhanced leaching losses.

Nitrogen Inputs

Forest management can also affect the input of N to

the system. This is done most directly through

fertilization. Nitrogen capital of forest sites can also be

managed by including native or introduced N2-fixing

symbiotic plants in mixtures or as part of a rotation.

Fertilization-Forest productivity, particularly on poorer

sites, can often be enhanced by N fertilization. This

practice has been quite extensive in the Pacific

Northwest (Cole and Gessel 1992). In this region, a

typical fertilization regime is the addition of about 200

kg N ha-1, which has a growth response of up to about

10 years. The immediate, and most beneficial, result

of N fertilization is enhanced tree growth because of

plant uptake. In general, the greatest relative response

is found on the poorest (most N limited) sites, although

greater absolute increases in growth may occur on

110
better sites. Greater response is often seen when

fertilization is done in conjunction with thinning.

A review of forest N fertilization studies shows that

about 20-30% of the added N is taken up by trees and

ground vegetation (Nason and Myrold 1992). Most of

the fertilizer N (50-60%) is immobilized into soil organic

matter. The remaining 10-30% is lost from the

system. When urea is used, much of the N can be lost

through NH3 volatilization. The pulse of available N

coming from fertilization usually results in a short-term

enhancement of nitrification (Heilman, 1974). This

NO3

-

 can potentially be lost by leaching and

denitrification. The fertilizer N that is incorporated into

soil organic matter represents only a small increase

(<5%) in the total soil N pool and therefore probably

has relatively little long-term impact on N cycling,

although there is some evidence that litter

decomposition and C mineralization may be depressed

for a period of time.

Nitrogen-fixing symbioses-The Frankia that form

root nodules on actinorhizal plants are the major N2-

fixing symbiosis in temperate forests. In the Pacific

Northwest, these actinorhizal plants include Alnus spp.,

Ceanothus spp., Shepherdia spp., Cercocarpus spp.,

and Purshia tridentata. Rates of N2 fixation range from

20-150 kg N ha-1 yr-1 for alders, 5-100 kg N ha-1 yr-1 for

Ceanothus, and <20 kg N ha-1 yr-1 for Purshia

tridentata (Hibbs and Cromack 1990). Whether part of

natural succession, grown in rotation, or in mixed

stands, these levels of N input can have a significant

impact on the growth of associated trees (for example,

Hibbs and others 1994) and building of soil organic N

reserves (table 2). The continual input of N through

the leaf litter and fine root turnover of actinorhizal

plants increases the availability of N, which can alter

the internal N cycling processes in the soil. Hart and

others (1997) showed that the presence of red alder

increased rates of N cycling, enhancing rates of NH4

 +

and NO3

-

 production relatively more than those of

consumptive processes. This often leads to an

accumulation of NO3

-

 in soils of red alder stands, which

can lead to greater leaching losses and sometimes

greater denitrification (table 2).

IMPLICATIONS FOR SUSTAINABILITY

There are several points we can draw from the

information on N cycling in forests that has been

presented:

1. Human activities can decrease or increase N pool

sizes depending upon the type of activity.

2. Human activities generally speed up N turnover; N

cycles faster as the result of management activities.

3. The combined affect of these two generalizations is

that human activities usually result in greater losses

of N from forest systems compared to unmanaged

forests.

It is always important, however, to keep these

generalizations within the context of the degree, or

magnitude, of management activities and the inherent

characteristics of the site. Some sites are better

buffered to changes in N inputs and outputs (resistant)

and some recover more rapidly (resilience). Resistance

and resilience to management activities, and hence

long-term sustainability, will require us to:

1. Balance inputs of N with removals of N.

2. Understand the many factors that control N storage

and turnover in forest soils.

ACKNOWLEDGMENTS

This work was made possible by the support of the

Oregon State University Agricultural Experiment Station

and Forestry Research Laboratory, and by past grant

support by the National Science Foundation and

U.S.D.A. National Research Initiative programs.

QUESTIONS FROM THE FLOOR

Q: How important is lichen-nitrogen input to forest

ecosystems?

A: It is known that lichens that form symbioses with

cyanobacteria, such as the Lobaria oregana, which

grows on the branches of old-growth Douglas-fir trees,

can fix N2. The annual input from these sources is a

few kg N ha-1 y-1 which is similar to inputs from

precipitation, and probably balances N outputs. The

importance of these lichens in younger forests is

probably less and not sufficient to compensate for

harvest removals.

Q: How effective are riparian reserves in mitigating

loss of NO3

-

 through leaching? What characteristics of

riparian reserves make them more effective?

A: Vegetated riparian zones have the capacity to take

up N and store it in plant biomass and, ultimately, as

soil organic matter. Nutrients, such as NO3

-

, in the soil

water that moves through the rooting zone of a riparian

buffer strip, can be taken up by plants, immobilized in

the soil, and, in the case of NO3

-

, volatilized by

denitrification. This has been most clearly shown using

riparian buffer strips adjacent to agricultural lands.

Their exact capacity to mitigate NO3

-

 movement to

111
Table 2-lnfluence of red alder on N cycling in Pacific Northwest forests (Binkley and others 1992).

Site

Wind River, WA

Cascade Head, OR

Treatment

Conifer

Conifer-alder mix

Conifer

Conifer-alder mix

Soil N

- kg N ha-1 -

2100

4700

10,000

14,000

N Leached

- kg N ha-1 y-1 -

5

25

20

50

Denitrification

- kg N ha-1 y-1 -

0.1

0.2

0.1

0.1

streams will certainly vary depending upon the size of

the buffer strip, soil type, plant community, net primary

productivity, and other factors. Obviously bigger

buffers have greater capacity and riparian vegetation

with greater growth rates can take up more NO3

-

.

Q: Explain what happens to added N fertilizer after 10

years. Does the ecosystem revert back to the Nitrogen

cycle balance before added fertilizer? Does the tree

growth slow down to what it was before?

A: Empirical studies in the Pacific Northwest have

shown that the growth response to the standard 200

kg N ha-1 urea fertilization lasts up to ten years,

eventually returning to control (non-fertilized) growth

rates. There is very little process-level work examining

the long-term (>10 y) effect of N fertilization on N

transformations in soils. The few

 15N studies have

shown that 50-60% of the N is immobilized into soil

organic matter and 20-30% is taken up by trees and

understory vegetation; the remaining N is presumably

lost by NH3 volatilization, leaching, and denitrification.

Urea fertilization often results in enhanced nitrification

for the first few years following application but this

usually returns to pre-fertilization rates after 10 years.

Some studies have shown N fertilization to depress

decomposition and C mineralization (presumably by

inhibiting lignin degrading fungi) during the first few

years following fertilization, but longer lasting effects

have not be researched.

Q: Yesterday, it was mentioned that if there is less of

something within soils-it becomes more critical if some

is removed from the system, versus if there is an

abundance of something and some is removed there is

less of an impact. Question: Much talk about elements

C, N, and P. Do forest activities affect micronutrients?

Do we know to what degree? What role do these

micronutrients play for soil organisms and soil

health/resiliency?

A: There are certainly examples when micronutrients,

such as Cu and B, are limiting to tree growth, although

they are not as widespread as macronutrient growth

limitations and have not been studied very intensively.

Because micronutrients are required in very small

amounts, management practices that selectively

remove them (e.g., harvesting, burning of litter layer,

etc.) would be expected to be most detrimental. There

is not much hard data, however. Soil organisms, like

trees, also require micronutrients but there is even less

information that they may actually be limited by them

in forest soils, except for a study that suggested that

N2-fixing bacteria may be Mo-limited on some sites.

LITERATURE CITED

Biggar, CM. ; Cole, D.W. 1983. Effects of harvesting

intensity on nutrient losses and future productivity

in high and low productivity red alder and Douglasfir stands. In: Ballard, R.; Gessel, S.P. (eds.).

I.U.F.R.O. symposium on forest site and continuous

productivity. Gen. Teh. Rep. PNW-163, Portland,

OR: U.S. Department of Agriculture, Forest Service,

Pacific Northwest Forest and Range Experiment

Station: 167-178.

Binkley, D.; Sollins, P.; Bell, R.; Sachs, D.; Myrold, D.

1992. Biogeochemistry of adjacent conifer and

alder-conifer stands. Ecology. 73(6)2022-2033.

Cole, D.W.; Gessel, S.P. 1992. Fundamentals of tree

nutrition. In: Chappell, H.N.; Weetman, G.F.; Miller,

R.E. (eds.). Forest fertilization: sustaining and

improving nutrition and growth of western forests.

Seattle, WA: Institute of Forest Resources

Contribution 73, College of Forest Resources,

University of Washington: 7-16.

Cole, D.W.; Johnson, D. 1981. The cycling of

elements within forests. In: Heilman, P.E.;

Anderson, H.W.; Baumgartner, D.M. (eds.). Forest

soils of the Douglas-fir region. Pullman, WA:

Cooperative Extension Service, Washington State

University: 185-198.

Davidson, E.A.; Myrold, D.D.; Groffman, P.M. 1990.

Denitrification in temperate forest ecosystems. In:

Gessel, S.P.; Lacate, D.S.; Weetman, G.F.; Powers,

R.F. (eds.). Sustained productivity of forest soils:

112
Proceedings of the 7th

 North American Forest Soils

Conference; 1988 July 24-28; Vancouver, British

Columbia. Vancouver, British Columbia: University

of British Columbia, Faculty of Forestry Publication:

196-220.

Dawson, J. 1983. Dinitrogen fixation in forest

ecosystems. Canadian Journal of Microbiology.

29:979-992.

Fredriksen, R.L.; Harr, R.D. 1981. Soil, vegetation and

watershed management. In: Heilman, P.E.;

Anderson, H.W.; Baumgartner, D.M. (eds.). Forest

soils of the Douglas-fir region. Pullman, WA:

Cooperative Extension Service, Washington State

University: 231-260.

Fredriksen, R.L.; Moore, D.G.; Norris, L.A. 1975. The

impact of timber harvest, fertilization, and herbicide

treatment on streamwater quality in western Oregon

and Washington. In: Bernier, B; Winget, C.H., eds.

Forest soils and forest land management:

Proceedings of the 4th North American Forest Soils

Conference; 1973 August; Laval, Quebec. Laval,

Quebec: Laval University, Les Presses de

L'Universite Laval: 283-313.

Hart, S.C.; Binkley, D.; Perry, D.A. 1997. Influence of

red alder on soil nitrogen transformations in two

conifer forests of contrasting productivity. Soil

Biology & Biochemistry. 29(7):1 111-1123.

Hart, S.C.; Nason, G.E.; Myrold, D.D.; Perry, D.A.

1 994. Dynamics of gross nitrogen transformations

in an old-growth forest: the carbon connection.

Ecology. 75(4):880-891.

Heilman, P. 1974. Effect of urea fertilization on

nitrification in forest soils of the Pacific Northwest.

Soil Science Society of America Proceedings.

38:664-667.

Hibbs, D.E.; Cromack, K. Jr. 1990. Actinorhizal plants

in Pacific Northwest forests. In: Schwintzer, C.R.;

Tjepkema, J.D. (eds.). The biology of Frankia and

actinorhizal plants. San Diego, CA: Academic

Press, Inc.: 343-363.

Hibbs, D.E.; DeBell, D.S.; Tarrant, R.F. 1994. The

biology and management of red alder. Corvallis,

OR. Oregon State University Press. 256 p.

Myrold, D.D.; Nason, G.E. 1992. Effect of acid rain

on soil microbial processes. In: Mitchell, R. (ed.).

Environmental microbiology. New York, NY: WileyLiss, Inc.: 59-82.

Nason, G.E.; Myrold, D.D. 1992. Nitrogen fertilizers:

fates and environmental effects in forests. In:

Chappell, H.N.; Weetman, G.F.; Miller, R.E. (eds.).

Forest fertilization: sustaining and improving

nutrition and growth of western forests. Seattle,

WA: Institute of Forest Resources Contribution 73,

College of Forest Resources, University of

Washington: 67-81.

Sollins, P.; Grier, C.C.; McCorison, F.M.; Cromack, K.

Jr.; Fogel, R.; Fredriksen, R.L. 1980. The internal

element cycles of an old-growth Douglas-fir

ecosystem in western Oregon. Ecological

Mongraphs. 50(3):261-285.

Vermes, J.-F.; Myrold, D.D. 1992. Denitrification in

forest soils of Oregon. Canadian Journal of Forest

Research. 22:504-512.

Vitousek, P.M.; Gosz, J.R.; Grier, C.C.; Melillo, J.M.;

Reiners, W.A.; Todd, R.L. 1979. Nitrogen losses

from disturbed ecosystems: interregional

comparative studies show mechanisms underlying

forest ecosystem response to disturbance. Science.

204:469-474.

Weetman, G.F.; Fournier, R.; Schnorbus Panozzo, E.;

Barker, J. 1990. Post-burn nitrogen and

phosphorus availability of deep humus soils in

coastal British Columbia cedar/hemlock forests and

the use of fertilization and salal eradication to

restore productivity. In: Gessel, S.P.; Lacate, D.S.;

Weetman, G.F.; Powers, R.F. (eds.). Sustained

productivity of forest soils: Proceedings of the 7th

North American Forest Soils Conference; 1988 July

24-28; Vancouver, British Columbia. Vancouver,

British Columbia: University of British Columbia,

Faculty of Forestry Publication: 451-499.

113
WILDFIRE VS. CLEARCUTTING: IMPACTS ON ECTOMYCORRHIZAL AND DECOMPOSER

FUNGI

S. Visser1 and D. Parkinson2

ABSTRACT

Intense wildfire is more destructive to the forest floor than timber harvesting with potentially more impact on fungal

communities as loss of forest floor structure, microhabitat and resource diversity is more extreme after wildfire. After

intense wildfire, decomposer and EM fungi associated with the forest floor are virtually eliminated since the organic

matter may be almost completely combusted exposing the underlying mineral soil. Clearcutting and site scarification

may compress and redistribute forest floor organic matter, but losses are low relative to those from intense wildfire.

Studies in lodgepole and jack pine forests in Alberta indicate that wildfire destabilizes the EM community to a greater

extent than clearcutting and recovery to pre-disturbance conditions is more rapid following clearcutting. Decomposer

fungi in the forest floor are significantly reduced by intense wildfire while the fungal community in the underlying

mineral soil remains relatively unaltered. Composition of the decomposer fungal community in the forest floor recovers

rapidly after burning while community structure requires a longer timeframe to stabilize. However, changes in the

decomposer fungal community following intense wildfire has negligible repercussions on litter decomposition.

Clearcutting, either with or without site preparation appears to have minimal effects on the composition and structure

of decomposer fungal communities. Management of harvested sites for the protection of EM and decomposer fungi

should concentrate on minimizing soil disturbance during site preparation, eliminating prescribed burning, and ensuring

rotation times are sufficiently long to allow slow-growing fungi to complete their lifecycles.

Keywords: wildfire, clearcutting, site preparation, ectomycorrhizal fungi, decomposer fungi, recovery, community

composition, community structure, decomposition

INTRODUCTION

As parasites, symbionts and decomposers, fungi

contribute significantly to the regeneration, growth and

maintenance of forests. Forest fertilility, resilience and

sustainability are determined by the ability of the fungi

to acquire, translocate, immobilize and recycle nutrients

essential to ecosystem function. Diversity of

substrates and microhabitats in forest soils as a result

of annual inputs of a wide range of dead plant residues

(wood, leaves, roots) creates ideal conditions for the

development of species-rich fungal communities with

wide enzymatic potentials. Resiliency of forest soils to

disturbance has been attributed to the presence of

highly diverse microbial communities with inherent

functional redundancy although the degree of resiliency

as related to the extent and intensity of disturbance are

not well understood.

biodiversity and ecosystem productivity (Bergeron and

Harvey 1997). However, in order to develop

silvicultural regimes patterned after natural (wildfire)

disturbance it is essential that the abiotic and biotic

processes associated with forest recovery be well

understood. Plant community dynamics following

wildfire or clearcutting have been well-researched

(Keenan and Kimmins 1993; Johnson 1981; Shafi and

Yarranton 1973) while the dynamics of fungal

communities and associated nutrient cycling processes

have received much less attention. This aspect of

forest recovery is considered in the present paper

which reviews the relative impacts of wildfire versus

clearcutting disturbance on ectomycorrhizal and

decomposer fungal communities in lodgepole (Pinus

contorta), jack pine (Pinus banks/ana) or mixedwood

(Populus tremuloides; Picea glauca) forests in the boreal

zone of Alberta and Ontario.

Two of the main types of disturbance in northern

temperate forests are wildfire and clearcutting.

Recently, there has been much interest in basing

silviculture on natural ecosystem dynamics such as

those following wildfire, with the goal of preserving

1Suzanne Visser is a research associate and adjunct assistant professor, Dept. of Biological Sciences, The University of Calgary,

Calgary, AB T2N 1N4.

2Dennis Parkinson is a faculty professor, Dept. of Biological Sciences, The University of Calgary, Calgary, AB T2N 1N4.

114
COMPARISON OF WILDFIRE AND

CLEARCUTTING ON ECOSYSTEM

ATTRIBUTES

Wildfires in the boreal forest tend to be large-scale with

fires over 10,000 ha in area accounting for 90% of the

area burned (wildfires 100,000 ha in area are not

uncommon) (Johnson 1992). Fire return intervals

(recovery time) average 100 years (Johnson 1992).

Disturbance intensity is often high in that there is high

mortality of the canopy, understory and groundcover

and extensive combustion of the forest floor organic

matter.

In contrast, clearcutting is conducted on a smaller scale

with allowable cuts in Alberta ranging from 60 to 100

ha (Alberta Environmental Protection, Lands and Forest,

personal communication). Rotation times are unknown

but have been estimated to be 80 to 100 years, a

timescale similar to the average fire return interval.

Harvesting eliminates the canopy while understory and

groundcover species may increase in abundance and

productivity (Keenan and Kimmins 1993). Forest

regeneration practices generally involve some form of

site preparation such as scarification where the forest

floor is disrupted by mechanical means to expose the

mineral soil so regenerating conifers will have a

competitive advantage over more rapidly growing

species. Scarification results in redistribution and

mixing of soil organic matter.

Abiotic and biotic parameters important in the

composition and structure of both mycorrhizal and

decomposer fungal communities and which may be

altered by either wildfire or clearcutting disturbance

include soil temperature, pH, organic matter content,

total and available nutrients and plant community

composition and structure. Removal of the canopy

following wildfire or timber harvesting causes an

increase in soil temperature (Keenan and Kimmins

1993; Kozlowski and Ahlgren 1974). The production

of ash during a wildfire increases the pH of the surface

soil more so than disturbance of the surface soil

horizons caused by harvesting and site preparation

where the pH may increase slightly or remain unaltered

(Kozlowski and Ahlgren 1974; Pietikainen and Fritze

1995). Combustion of the forest floor during an

intense wildfire significantly reduces soil organic matter

(fig. 1) thereby reducing pools of total nutrients.

However, available phosphorus and nitrogen increase

following wildfire disturbance as nitrogen mineralization

is enhanced and ashing renders phosphorus more

available (Raison 1979; Raison and McGarity 1980).

Indeed in a study by Visser (unpublished), there was

little difference in extractable nitrogen in the upper 20

cm of soil in 6, 41 and 65 year old jack pine stands.

Recovery of the forest floor in jack pine stands

following wildfire in northern Alberta is relatively rapid

Figure 1 —Forest floor mass or depth in a chronosequence of

pine or mixedwood stands following wildfire or clearcutting.

as no differences in organic matter accumulation were

detectable amongst 41 , 65 and 122 year old stands

which had originated from wildfire (fig. 1).

Harvesting alone appears to have little impact on mass

of organic matter in aspen/spruce mixedwood stands

over the short term (fig. 1). Although harvesting

followed by scarification may cause compaction and

redistribution of soil organic matter in the surface

horizons, thereby reducing the depth of the forest floor

(fig. 1), it may not have a significant impact on mass of

organic matter in the surface soil horizon (fig. 1).

These data indicate that intense wildfire is more

destructive to the forest floor than timber harvesting

with potentially more impact on fungal communities as

loss of forest floor structure, microhabitat and resource

diversity is more extreme in the case of intense

wildfire. As observed for wildfire, available nitrogen

and phosphorus often increase following clearcutting

and scarification (Keenan and Kimmins 1993).

Alterations in soil organic matter quality and quantity in

conjunction with changes in plant community

composition and structure following either wildfire or

harvesting/scarification disturbance would be expected

to have significant impacts on the soil fungi since these

factors drive symbiont and decomposer fungal

communities. Removal of the canopy and increased

dominance of forbs and grasses after both types of

disturbance would lead not only to major shifts in the

composition, availability, distribution and physiology of

regenerating mycorrhizal hosts, but would also shift the

115
quality and quantity of plant residues feeding the

decomposer fungi. How these fungi respond to these

changes and their patterns of recovery in regenerating

forests and their forest floors is unknown.

WILDFIRE VS CLEARCUTTING/SITE

PREPARATION EFFECTS ON

ECTOMYCORRHIZAL FUNGAL

COMMUNITIES

Fruitbody-producing Ectomycorrhizal (EM) Fungi -

Occurrence and Abundance

It is recognized that fruitbody presence does not

provide a complete picture of the symbionts colonizing

the roots (Jansen and De Nie 1988) since some

ectomycorrhizal fungi do not form conspicuous

fruitbodies and fruiting occurs at various stages of the

life-cycle depending on species, host and environmental

conditions. Nevertheless, when combined with

observations of ectomycorrhizal fungi colonizing the

roots, fruitbodies confirm the presence of a particular

species and can provide valuable information on the

dynamics and life-history strategies of mycorrhizal fungi

involved in forest recovery.

Based on fruitbody observations, recovery patterns of

ectomycorrhizal fungi in pine stands disturbed by either

wildfire or clearcutting/scarification appear to be very

similar (Bradbury and others 1998; Visser 1995) (tables

1, 2). The initial impact of both clearcutting and

wildfire disturbance is a reduction in the number of

fruiting mycorrhizal fungi, but with time the number of

fruiting species increases and the complement of fungi

becomes more similar to that observed in mature 90 to

122 year old stands. However, the greater similarity

between fungi in young and old stands following

clearcutting than following wildfire suggests that the

ectomycorrhizal community may recover more rapidly

or may not be as impacted by harvesting as it is by

wildfire. Species which are typical of younger stands,

for example Suillus spp., persist throughout the lifetime

of the stand although their abundance may change

with forest maturation, thus there is no evidence of

complete replacement of species (succession) with time

(Bradbury and others 1998; Dahlberg and Stenstrom

1989; Visser 1995). The increase in number of fruiting

species with stand age may reflect the growth rates of

the various species and the time it takes each species

to amass enough energy to fruit. For example, species

such as Hygrophorus purpurascens and Hydnellum

peckii fruit only in older stands presumably because

they are very slow growing and require decades to

produce a mycelium capable of fruiting. As a result,

these are species which may be most susceptible to

clearcutting disturbance if rotation times are shorter

than the time required for them to recover and

complete their lifecycles.

Root-colonizing Ectomycorrhizal (EM) Fungi —

Occurrence and Abundance

A more accurate evaluation of the EM fungi associated

with a particular host is obtained by conducting a

microscopic examination of the fungi colonizing the

feeder roots and identifying them using criteria outlined

in identification manuals published by Agerer (1987-

1994) and Ingleby and others (1990). This approach

has been applied to assess the recovery of the EM

fungi in lodgepole and jack pine stands in Alberta

(Bradbury and others 1998; Visser 1995).

In these pine stands, both wildfire and

clearcutting/scarification disturbance reduced the

number of EM species colonizing pine roots, but the

reduction was greater following wildfire (56%) than

following clearcutting/scarification (32%) (Bradbury and

others 1998; Visser 1995 ) (tables 3, 4). Although EM

community similarity to that in the mature stand

increased with stand age, pre-disturbance conditions

were attained much more rapidly following clearcutting

than following intense wildfire. Six years after

clearcutting and scarification, EM community

composition and structure were very similar to that in

mature, uncut stands; this was not observed in stands

regenerating after wildfire where community

composition and structure in the young stand were

distinctly different from that in the older stands (tables

3, 4; fig 2). Thus wildfire appears to destabilize the

EM community to a greater extent than clearcutting

and recovery to a large, multispecies community is

more rapid following clearcutting than after wildfire.

Fungi colonizing the roots of regenerating trees in both

clearcut and wildfire-disturbed stands were also found

on roots in mature, undisturbed stands suggesting no

distinct EM succession with stand maturation which

supports the fruitbody observations (Bradbury and

others 1998; Visser 1995). However, the abundances

of the fungi comprising the communities appeared to

change with time as was evident also in the fruitbody

observations. The functional significance of changes in

abundance of various EM species with time is unclear

but an increase in the size and density of perennial

mycelia of particular species over time may result in

more efficient acquisition of water and nutrients and

closer linkages with the decomposer microflora. Also,

it has been proposed that EM fungi which dominate

roots of mature stands have greater enzymatic

capabilities providing access to organic carbon,

nitrogen and phosphorus in forests which become more

nutrient-limited with age (Dighton and Mason 1985).

116
Table 1 —Most frequent ectomycorrhizal fungi observed fruiting over three growing seasons in a chronosequence of lodgepole pine

stands following clear-cutting disturbance. (+ = 1-9, + + = 10-99, + + + = 100-999, + + + + > 1000 fruitbodies). Data

modified from Bradbury and others (1998).

Species

Cortinarius spp.

Laccaria laccata

La ctarius rufus

Leccinum ponderosum

Suillus tomentosus

Suillus umbonatus

Hygrophorus hypothejus

Hygrophorus eburneus

Hydnellum peckii

TOTAL (all species)

SIMILARITY (with time)a

6

+ +

+ +

+

+

+

+ +

19

0.6

Stand age

10

+ + +

+ +

+

+ +

+ +

+ + +

+

25

0.7

(years)

19

+ + +

+ +

+ + +

+ + +

+ +

+ + + +

+

+

30

0.8

90

+ + +

+ + +

+ + + +

+ + +

+ + +

+ + +

+ + +

+ + +

+ + +

35

1

a

 Based on Sorenson's qualitative index of similarity.

Table 2 —Most frequent ectomycorrhizal fungi observed fruiting over two growing seasons in a chronosequence of jack pine stands

following wildfire disturbance. (+ = < 10, ++ = 10-20, + + + = > 20 fruitbodies per growing season). Data based on Visser

(1995).

Species

Suillus brevipes

Inocybe spp.

Lactarius deliciosus

Tricholoma zelleri

Suillus tomentosus

Cortina rius spp.

Bankera fuligineo-alba

Russula spp.

Hydnellum peckii

TOTAL (all species)

SIMILARITY (with time)a

6

+ + +

+ + +

+

14

0.3

Stand

41

+

+ + +

+ + +

+ + +

+ + +

+ + +

+ + +

+ +

33

0.6

age (years)

65

+ +

+ +

+ +

+

+ + +

+ + + +

+ +

+ + +

+ + +

40

0.7

122

+ +

+

+ +

+ + +

+ + +

+

+ + +

+ +

33

1

a

 Based on Sorenson's qualitative index of similarity.

117
Table 3 —Most frequent ectomycorrhizal fungi colonizing

lodgepole pine roots in a chronosequence after clear-cutting.

Data modified from Bradbury and others (1998) and expressed

as % relative abundance.

Based on Sorenson's qualitative index of similarity.

Figure 2—Recovery of EM fungal community structure

following clearcutting/scarification or intense wildfire based on

rank abundance curves of fungi colonizing pine roots in an age

sequence of stands.

Table 4 —Most frequent ectomycorrhizal fungi colonizing jack

pine roots in a chronosequence after wildfire. Data from

Visser (1995) and expressed as % relative abundance.

Based on Renkonen similarity coefficients.

The few studies conducted to date indicate that in

inland boreal forests intense wildfire alters EM

communities more extensively than clearcutting and

scarification do, possibly because the extent and

degree of host/soil disturbance is greater after wildfire

than clearcutting. Harvesting accompanied by burning

of logging slash reduces EM fungal inoculum (Harvey

and others 1980; Parke and others 1984) more than

clearcutting and scarification suggesting EM fungi are

more sensitive to burning than other types of

silvicultural disturbances. Although patterns of EM

recovery are similar after either wildfire or clearcutting,

rates of recovery appear to be more rapid following

clearcutting than following wildfire.

Both jack pine and lodgepole pine forests are adapted

to frequent wildfires and can recover rapidly following

this type of disturbance. Adaptation of life histories to

fire cycles by both the hosts and their EM symbionts

would explain the resilience exhibited by these

communities. This resilience allows EM communites to

recover rapidly from other types of disturbance as well,

such as clearcutting and scarification. Management of

harvested sites for the protection of EM fungi should

concentrate on minimizing soil disturbance during site

preparation, avoiding burning of slash and ensuring

rotation times are long enough to allow slow-growing

EM fungi to complete their lifecycles.

118

Species

Lactarius deliciosus

Suillus brevipes/tomentosus

Mycelium radicis atrovirens

Piloderma fallax/byssinum

Cenococcum geophilum

Tomentella

Russula

Cortinarius/Dermocybe

Hygrophorus

TOTAL (all species)

SIMILARITY (with time)a

Stand age

6

21

28

27

4

7

8

2

2

0

13

0.8

10

1

35

21

24

13

1

4

2

0

15

0.

9

(years)

19

0

17

21

20

14

9

9

9

0

15

0.

9

90

2

2

22

30

6

1

9

7

2

19

1

Species

Cenococcum geophilum

Mycelium radicis

atrovirens

Suillus brevipes

Russula spp.

Tricholoma spp.

Suillus tomentosus

Lactarius deliciosus

Cortinarius spp.

Hydnellum peckii

Piloderma byssinum

Hygrophorus spp.

TOTAL (all species)

SIMILARITY (with time)a

6

7

12

77

1

1

0

0

0

0

0

0

12

0.2

Stand age

41

2

1

23

19

13

19

1

2

6

1

0

20

0.3

(years)

65

2

1

14

23

12

1

9

13

6

10

0

25

0.5

122

3

14

5

19

10

5

0

5

3

10

6

27

1
WILDFIRE VS. CLEARCUTTING/SITE

PREPARATION EFFECTS ON

DECOMPOSER MICROFUNGI

Wildfire Effects —Fungal Occurrence and

Abundance

Through their extensive enzymatic capabilities, the

decomposer fungi are essential in the decomposition

and mineralization of nutrients from dead plant

residues, nutrients which are subsequently taken up by

the EM fungi and translocated to the host. Therefore,

recovery of this component of the soil flora following

clearcutting versus wildfire disturbance should also be

considered in the development of forest management

plans. Few studies address decomposer fungal

community dynamics following disturbance; the

following discussion is based on research conducted in

a chronosequence of jack pine stands originating from

wildfire and boreal mixedwood stands in Alberta (Visser

unpublished) and Ontario (Houston and others in press)

after harvesting with and without site preparation.

In jack pine stands in northern Alberta, intense wildfire

almost totally eliminated the forest floor (fig. 1)

including the decomposer fungal community associated

with it (Visser 1995). However, as the forest floor

redeveloped in the regenerating stands, the fungal

community recovered so that 41 years after wildfire

disturbance (possibly sooner, stands between 6 and 41

years were not studied), species richness and the

composition of the decomposer fungal community were

very similar to that in a mature stand (table 5).

Although the composition of the community appeared

to recover within 41 years of wildfire disturbance,

species abundance curves revealed that community

structure required 65 years to stabilize. At 41 years,

the decomposer fungal community tended to be

dominated by three species which is typical of an early

successional community (fig. 3). By 65 years, an

increase in complexity of the plant community and

forest floor microhabitats probably resulted in a more

stable fungal community composed of more species

whose abundances were more equally distributed

(fig. 3). The structural difference between the fungal

communities in the 41 and 65 year old stands is

interesting and may reflect differences in the state of

forest floor decomposition between the two stands.

With age, humification of forest floor organic matter

and microhabitat and resource complexity would be

expected to increase leading to a more diverse and

stable decomposer community.

Regardless of stand age the fungal communities in jack

pine forest floors were dominated by Oidiodendron

griseum, Mycelium radicis atrovirens complex,

Trichoderma viride, Mortierella ramanniana, Penicillium

spp. and Mucor spp. while the communities in the

Table 5 —Dominant decomposer fungi in the forest floor

and mineral soil of a chronosequence of jack pine stands

originating from wildfire. Data expressed as % relative

abundance.

Species

FOREST FLOOR:

Oidiodendron griseum

Mycelium radicis

atrovirens

Trichoderma viride

Mortierella ramanniana

Penicillium spp.

Mucor sp.

TOTAL SPECIES.

MINERAL SOIL (0-10 cm):

Fusarium spp.

Trichoderma viride

Oidiodendron griseum

Penicillium spp.

Thysanophora

penicillioides

Mucor circinelloides

Mycelium radicis

atrovirens

TOTAL SPECIES

6

NA

25

13

41

64

25

13

6

38

Stand age (years)

41

2

1

87

21

10

11

34

0

26

17

35

10

6

4

30

65

21

11

39

23

34

5

42

0

6

48

31

19

31

12

38

122

56

15

17

28

38

14

40

0

1

55

41

33

24

19

35

NA = not applicable; forest floor absent.

underlying 0-10 cm deep mineral soil were dominated

by Fusarium spp. (6 year old stand only), Trichoderma

viride, Oidiodendron griseum, Penicillium spp.,

Thysanophora penicillioides, Mucor circinelloides and

Mycelium radicis atrovirens. Although the fungi in the

forest floor were significantly affected by wildfire, the

fungal community in the mineral soil appeared to be

unaltered relative to that in a 122 year old stand.

Similar observations have been made in burned and

unburned subalpine, coniferous forests in the Rocky

Mountains six years after a moderately severe wildfire

(Bissett and Parkinson 1980). The occurrence of

Fusarium spp. in the 6 year-old regenerating stand and

not in the older stands may be explained by the

increase in grasses following wildfire. Fusarium is

commonly associated with grassland soils. The

119
Figure 3 —Recovery of decomposer fungal community

structure following clearcutting or intense wildfire based on

rank abundance curves of fungi in the forest floor of variously

aged jack pine (wildfire) or mixedwood (clearcut) stands.

increase in abundance of Oidiodendron griseum, a

mycorrhizal symbiont of the Ericaceae, in the forest

floor with stand age (table 5), may be indicative of

increased colonization of the forest floor by roots of

ericaceous plants such as Vaccinium.

Most of the fungi found in the mineral soil also

occurred in the forest floor suggesting that colonization

of forest floor organic matter as it accumulates in

regenerating jack pine stands is to some extent

facilitated by fungi in the mineral soil. During wildfire,

killing temperatures are restricted to the upper 5 cm of

mineral soil (Sims 1976) and it is assumed that fungi

surviving below this depth would rapidly recolonize any

microhabitats vacated as a result of high temperatures.

Aerial and insect/animal dispersal of fungal propagules

would be another mechanism for reintroducing

decomposer fungi eliminated by fire.

Clearcutting/Site Preparation Effects —Fungal

Occurrence and Abundance

In contrast to wildfire disturbance, timber harvesting in

the absence of site preparation had no immediate

impact on forest floor organic matter (fig. 1) and no

detectable effects on either the composition or the

structure of the decomposer fungal community (table

6, fig. 3) two years after harvesting a mixedwood

(Populus tremuloides, Picea glauca) site near Edson,

Alberta. The decomposer fungi in both two year old

harvested stands and 80 year old unharvested stands

were very similar and dominated by the same species -

species ubiquitous to well-developed, mature

mixedwood forest floors. These results suggest that if

the forest floor is not extensively disturbed during

timber harvesting, short-term impacts on the

decomposer microfungi are negligible.

In a similar study conducted in two mixedwood sites in

Ontario, there were few differences in fungal

community composition and structure between

unharvested sites and adjacent sites seven to nine

years after clearcutting and site preparation (Houston

and others in press). Although a few species such as

Penicillium canescens and Oidiodendron tenuissimum

occurred in either the harvested or unharvested sites,

frequencies of the 14 most abundant fungi were similar

between disturbed and undisturbed sites (table 7;

Houston and others in press). Rank abundance curves

(not presented) revealed no differences in community

structure in the forest floors of harvested and

unharvested sites (Houston and others in press). These

observations suggest that harvesting and disturbance

of the forest floor by scarification either has no

immediate impact on fungi in the forest floor or that

recovery of the decomposer microfungi to predisturbance levels occurs relatively rapidly, or a

combination of both factors. Minimal effects on fungal

community structure following clearcutting have been

reported also by Bääth (1981) and Baath and others

(1995) for pine forests in Sweden.

Based on studies conducted to date, wildfire alters

microfungal community composition and structure to a

greater extent than clearcutting and site preparation,

primarily because high intensity wildfire heats and

combusts forest floor organic matter, thereby

eliminating microhabitats, reducing fungal biomass and

changing the quality and quantity of organic matter

available for microbial consumption. Indeed, timber

harvesting appears to have few deleterious effects on

the microfungal community, particularly if disturbance

to the forest floor is minimized. Recovery of the

decomposer fungi in the forest floor to pre-disturbance

conditions may take many years following intense

wildfire whereas recovery following clearcutting/site

preparation appears to be rapid. The extent to which

wildfires affect fungal communities is probably related

to the intensity of the fire and recovery rates probably

vary depending on fire intensity and the amount of heat

generated. Research is required to confirm this.

Implications of Fungal Community Disruption on

Ecosystem Function

Although fungal community composition and structure

can be significantly disrupted by severe disturbances

such as high intensity wildfire, it is unclear how this

disruption translates into effects on ecosystem

function. A key ecosystem function is organic matter

120
Table 6 —Decomposer fungi in the forest floors of two unharvested and adjacent harvested mixedwood sites. The forest floor was

not scarified following harvesting. Data expressed as % frequency of occurrence.

Species

Trichoderma polysporum

Mycelium radicis atrovirens

Penicillium canescens

Trichoderma oblongisporum

Cylindrocarpon didymum

Mucor hiemalis

TOTAL (all species)

2 (Harvested)

61

28

15

13

10

15

37

Stand Age (years)

80 (Unharvested)

40

23

10

22

28

19

38

Table 7 —Decomposer fungi in the forest floors of two unharvested and adjacent harvested mixedwood sites in Ontario. The forest

floor was scarified following harvesting. Data based on Houston and others (in press) and is expressed as % frequency of

occurrence.

decomposition and measures of this function include

microbial respiration and rates of litter decomposition.

Six years after wildfire disturbance, microbial

respiration and biomass carbon of unburned grey

needles and FH material were significantly lower than

those measured in the forest floors of 41 , 65 and 1 22

year old jack pine stands, possibly because fire alters

the quality of microbially-available carbon in unburned

organic matter left after a fire (Pietikäinen and Fritze

1995). However, in a laboratory-controlled

decomposition study, mass losses of pine needles

placed on soil cores from 6, 41 , 65 and 122 year old

jack pine stands for three months were almost identical

even though the forest floor in the 6 year old stand

was almost completely absent (fig. 4). Therefore,

losses in the microfungal community as a result of

forest floor combustion does not translate into a

reduction in decomposition potential over the shortterm. This may be explained by colonization of preabscission needles by litter decay fungi introduced in

the air spora and continued decomposition of the

needles by these fungi after deposition. Also,

numerous forest floor microfungi which survive in the

mineral soil during wildfire have the capability of

degrading needles deposited by the regenerating stand

thereby recreating the forest floor. These dispersal and

survival mechanisms contribute to the maintenance of

resiliency within microfungal communities allowing

rapid recovery of the decomposition function.

Effects of timber harvesting on soil respiration appear

to be negligible either in the presence or absence of

scarification (Houston and others in press; Visser

unpublished) and this combined with no significant

impact on microfungal community composition and

structure strongly indicates that the litter

decomposition function would remain unaltered also.

This was confirmed in a 2 year litter decomposition

study in the field where mass loss of fireweed in two121

Species

Cladosporium cladosporioides

Mortierella vinacea

Penicillium janthinellum

Trichoderma viride

Penicillium canescens

Oidiodendron tenuissimum

TOTAL SPECIES

7 - 9 (Harvested/scarified)

8

23

7

5

0

6

42

Stand Age (years)

75-10 1 (Unharvested)

3

17

3

9

8

0

45
Figure 4—Laboratory decomposition of jack pine needles

placed on soil cores from a chronosequence of jack pine

stands originating from wildfire and field decomposition of

fireweed leaves placed in a two year old mixedwood clearcut

and adjacent uncut site.

year old harvested sites and adjacent unharvested sites

were not significantly different (fig. 4).

SUMMARY

When comparing the impacts of wildfire versus

clearcutting/site preparation on EM and decomposer

fungi, intense wildfire has more significant effects on

community composition and structure mainly due to

high mortality of EM hosts and combustion and

chemical alteration of forest floor organic matter. After

intense wildfire, decomposer and EM fungi associated

with the forest floor are virtually eliminated since the

organic matter may be completely combusted exposing

the underlying mineral soil. Clearcutting and site

scarification may compact and redistribute forest floor

organic matter, but losses are low relative to those

from intense wildfire. Data from lodgepole and jack

pine forests in Alberta indicate that wildfire destabilizes

the EM community to a greater extent than clearcutting

and recovery to predisturbance conditions is more rapid

following clearcutting. Decomposer fungi in the forest

floor are significantly reduced or eliminated by intense

wildfire while the fungal community in the underlying

mineral soil remains relatively unaltered and serves as a

source of inoculum in the regenerating forest floor.

Composition of the decomposer fungal community in

the forest floor recovers rapidly with the accumulation

of litter and development of the forest floor, but

community structure requires a longer timeframe to

stabilize. However, alteration of the decomposer

fungal community following intense wildfire does not

appear to have negative repercussions on ecosystem

function as expressed in litter decay rates.

Clearcutting, either with or without site preparation

appears to have minimal effects on the composition

and structure of decomposer fungal communities. Both

EM and decomposer fungi are sensitive to prescribed

burning when conducted in conjunction with

harversting resulting in significant reductions in EM

inoculum and decomposer fungal biomass.

Management of harvested sites for the protection of

EM and decomposer fungi should concentrate on

minimizing soil disturbance during site preparation,

eliminating burning of slash and ensuring rotation times

are of sufficient length to allow slow-growing fungi to

complete their lifecycles. This paper has emphasized

the impacts of intense wildfire; response of EM and

decomposer fungi to varying degrees of wildfire

intensity and variable combustion of the forest floor

requires further examination.

LITERATURE CITED

Agerer, R. (ed.). 1987-1994. Colour atlas of

ectomycorrhizae. Schwäbisch Gmund, Germany:

Einhorn-Verlag Eduard Dietenberger.

Baath, E. 1981. Microfungi in a clear-cut pine forest

soil in central Sweden. Canadian Journal of Botany

59:1331-1337.

Baath, E.; FrostegSrd, A.; Pennanen, T.; Fritze, H.

1995. Microbial community structure and pH

response in relation to soil organic matter quality in

wood-ash fertilized, clear-cut or burned coniferous

forest soils. Soil Biology and Biochemistry 27:229-

240.

Bergeron, Y.; Harvey, B. 1997. Basing silviculture on

natural ecosystem dynamics: an approach applied

to the southern boreal mixedwood forest of Quebec.

Forest Ecology and Management 92:235-242.

Bissett, J.; Parkinson, D. 1980. Long-term effects of

fire on the composition and activity of the soil

microflora of a subalpine, coniferous forest.

Canadian Journal of Botany 58:1704-1721.

Bradbury, S.M.; Danielson, R.M.; Visser, S. 1998.

Ectomycorrhizas of regenerating stands of lodgepole

pine {Pinus contorta). Canadian Journal of Botany

76:218-227.

122
Dahlberg, A; Stenstrom, E. 1989. The effect of clearcutting on the ectomycorrhizal flora on pine.

Agriculture, Ecosystems and Environment 28:91-94.

Dighton, J.; Mason, P.A. 1985. Mycorrhizal dynamics

during forest tree development. In: Moore, D.;

Casselton, L.A.; Wood, D.A.; Frankland, J.C. (eds.).

Developmental biology of higher fungi. Cambridge:

Cambridge University Press 117-139.

Harvey, A.E.; Larsen; M.J.; Jurgensen, M.F. 1980.

Partial cut harvesting and ectomycorrhizae: early

effects in Douglas-fir larch forests of western

Montana. Canadian Journal of Forest Research.

10:436-440.

Houston, A.P.C.; Visser, S.; Lautenschlager, R.A.

1998. Microbial processes and fungal community

structure in soils from clear-cut and unharvested

areas of two mixedwood forests. Canadian Journal

of Botany 76:630-640.

Ingleby, K.; Mason, P.A.; Last, F.T.; Fleming, L.V.

1990. Identification of ectomycorrhizas. ITE

research publication number 5. London:HMS0.

Jansen, A.E.; De Nie, H.W. 1988. Relations between

mycorrhizas and fruitbodies of mycorrhizal fungi in

Douglas fir plantations in The Netherlands. Acta

Botanica Neerlandia 37:243-249.

Johnson, E.A. 1992. Fire and vegetation dynamics:

studies from the North American boreal forest.

Cambridge: Cambridge University Press.

Johnson, E.A. 1981. Vegetation organization and

dynamics of lichen woodland communities in the

Northwest Territories, Canada. Ecology 62:200-

215.

Keenan, R.J.; Kimmins, J.P. 1993. The ecological

effects of clear-cutting. Environmental Reviews

1:121-144.

Kozlowski, T.; Ahlgren, C.E. 1974. Fire and

ecosystems. New York: Academic Press. 542 p.

Parke, J.L.; Linderman, R.G.; Trappe, J.M. 1984.

Inoculum potential of ectomycorrhizal fungi in forest

soil from southwest Oregon and northern California.

Forest Science 30:300-304.

Pietikainen, J.; Fritze, H. 1995. Clear-cutting and

prescribed burning in coniferous forest: comparison

of effects on soil fungal and total microbial biomass,

respiration activity and nitrification. Soil Biology

and Biochemistry 27:101-109.

Raison, R.J.; McGarity, J.W. 1980. Some effects of

plant ash on the chemical properties of soils and

aqueous suspensions. Plant and Soil. 55:339-352.

Raison, R.J. 1979. Modification of the soil

environment by vegetation fires, with particular

reference to nitrogen transformations: a review.

Plant and Soil. 51:73-108.

Shafi, M.I.; Yarranton, G.A. 1973. Vegetational

heterogeneity during a secondary (postfire)

succession. Canadian Journal of Botany 51:73-90.

Sims, H.P. 1976. The effect of presribed burning on

some physical soil properties of jack pine sites in

southeastern Manitoba. Canadian Journal of Forest

Research 6:58-68.

Visser, S. 1995. Ectomycorrhizal fungal succession in

jack pine stands following wildfire. New Phytologist

129:3889-401.

123
DECOMPOSITION AND NITROGEN MINERALIZATION IN FORESTS OF BRITISH

COLUMBIA: EFFECTS OF FOREST MANAGEMENT PRACTICES

Cindy E. Prescott1 and Lisa M. Zabek2

ABSTRACT

The productivity of most forests in the Pacific Northwest is limited by the availability of nitrogen. Nitrogen availability is

largely controlled by the rate at which N is recycled from organic matter, through the processes of decomposition and

mineralization. These processes are controlled by the factors that limit the activities of the soil organisms involved,

particularly temperature, moisture, and the physical and chemical nature of the organic matter. Forest management

activities may influence each of these factors and thereby affect rates of decomposition, N mineralization and N

availability.

The influences of several common forest management activities on decomposition and N mineralization were examined

in a suite of experiments across British Columbia. Rates of decomposition were compared in forests and adjacent

clearcuts at 22 sites, and rates were either the same or slower in the clearcuts. Several sites had additional silvicultural

treatments that provide a range of either opening size or removal intensity. Decomposition rates were not influenced by

opening size, but rates of net N mineralization and nitrification were increased in openings greater than 0.1 ha. Tree

species influences N mineralization in the forest floor, with highest rates occuring in Douglas-fir and broadleaf species

and lowest rates in cedar and pines. The decomposition rates of foliar litters was best predicted by its lignin

concentration. The influence of managing to maintain a component of broadleaf species was examined in

decomposition experiments with pure and mixed litter of aspen and spruce, Douglas-fir and alder, and lodgepole pine,

Douglas-fir and paper birch. There was no effect of mixing litters on their rates of decomposition. Likewise, N

fertilization had no effect on decomposition rates in trials in coastal Douglas-fir and aspen.

Keywords: decomposition, mineralization, nitrogen, fertilization, clear cut, alternative silviculture systems, species,

species mixtures

INTRODUCTION

Through the processes of decomposition and

mineralization, the nitrogen bound in organic matter is

released in inorganic forms (NH4-N and NO3-N) which

can be taken up by plants. Decomposition is the actual

breaking down of material and is measured as rates of

weight loss over time. Mineralization is the release of

nutrients from organic material as it is broken down.

Some portion of the mineralized nutrients are taken up

and temporarily immobilized by microorganisms, then

remineralized upon death. Rates of mineralization are

therefore net rates, i.e. the difference between rates of

mineralization and immobilization. Nitrification refers to

the production of NO3-N; this can be rapidly

reassimilated by microorganisms, so the net rate of this

process is measured. Rates of decomposition and

mineralization are affected by factors which influence

microbial activity, the most important of these are the

climate and the nature of the organic material. It is

critical to keep in mind that the effects of these

environmental factors are mediated through their

influence on the soil microbial and faunal communities,

whose activities directly affect rates of these processes

(fig. 1).

Figure 1 —Factors controlling rates of litter decomposition.

'Cindy E. Prescott is assistant professor in the Faculty of Forestry, University of British Columbia, Vancouver, B.C. V6T 1Z4.

2

Lisa M. Zabek is a graduate student in the Faculty of Forestry, University of British Columbia, Vancouver, B.C. V6T 1Z4.

124
The overriding influence of temperature and moisture

on rates of decomposition has been demonstrated in

laboratory experiments (Bunnell and others 1976), and

relationships between decomposition rate and climatic

variables such as actual evapotranspiration (AET)

across a wide range of climatic conditions have been

developed (Meentemeyer 1978). British Columbia has a

diversity of climates, due to its size (49°- 60°N),

maritime influence, and mountainous topography. The

result is a diversity of forest types, from coastal

rainforests to ponderosa pine savannas to boreal

mixedwoods, which have been used to divide the

province into 14 biogoeclimatic zones (Klinka and

others 1991). Rates of decompositon of standard litter

substrates have been measured at 26 installations in 9

forested biogeoclimatic zones within British Columbia.

At each installation, rates of mass loss of pine needle

litter, forest floor material and aspen leaf litter are being

measured for 5, 4 and 3 years, respectively. Rates of

mass loss will be compared with climate measurements

at each site or from the nearest climate station, to

determine the best relationships for predicting rates of

decomposition. Site fertility parameters such as soil

texture and C:N ratios of forest floor and soil will also

be measured to determine if they have an additional

influence on decomposition rates. Average values for

the weight of pine needles remaining after

decomposing for 3 years at each site were generally

fastest in forests on the coast (CWH, CDF and MH

zones) and in the interior wet-belt (ICH), and slowest in

dry interior forests (PP) (table 1). This suggests that

moisture is more influential than temperature on rates

of litter decomposition in British Columbia.

The physical and chemical nature (or "quality") of the

litter also determines its rate of decomposition. Of

particular importance are the relative concentrations of

labile and recalcitrant materials, i.e. those which are

readily metabolized by the microbes and those which

are not. Ratios such as C:N, lignin:N, lignin:cellulose:N

have proven to be useful predictors of decomposition

rates of different substrates (Berg and others 1987,

Taylor and others 1989a). Determination of C fractions

in litter by nuclear magnetic resonance spectroscopy

(

13C NMR) has also proven useful in characterizing

litters with respect to their potential to decompose and

release nutrients. Recent studies (Prescott and Preston

1994, Baldock and Preston 1995) have suggested that

the content of alkyl C (waxes and cutin) in particular,

Table 1 —Weight of lodgepole pine needle litter remaining after decomposing for 3 years at sites in each biogeoclimatic zone. Initial litter weight was 2.0 g. Mean (and standard deviation) of 7 samples per site.

Zonea

BWBS

CDF

CWH

ESSF

ICH

IDF

MH

PP

SBS

Mean annual

temperature

-1.4

9.5

9.2

1.1

6.9

4.2

5.0

8.6

3.3

Mean annual

precipitation

452

873

2140

1177

1063

4 1 4

2954

332

628

Site

Inga Lake

Bear Mt

Fairbanks

Shawnigan S

Shawnigan N

Blaney Lake

Pt McNeill

Otter Creek

Spanish Lake

Adams Lake

Malakwa

Hidden Lake

Valentine Lake

Boston Bar

Strachan

Garibaldi

Skihist

Trout Creek

Beedy Creek

Topley

Weight remaining (g)

1.07

0.83

1.41

0.88

0.77

0.86

0.69

0.77

1.05

1.01

0.63

0.87

1.07

1.06

0.89

0.88

1.55

1.24

0.66

0.88

(.06)

(.17)

(.04)

(.06)

(.11)

(.21)

(.12)

(.11)

(.20)

(.16)

(.10)

(.10)

(.11)

(.17)

(.15)

(.15)

(.08)

(.19)

(.11)

(.14)

125

;

 BWBS

CDF =

C W H =

ESSF =

ICH =

= Boreal Black and White Spruce

Coastal Douglas Fir

Coastal Western Hemlock

Engelmann Spruce Subalpine Fir

Interior Cedar Hemlock

IDF =

MH =

PP =

SBS

Interior Douglas Fir

Mountain Hemlock

Ponderosa Pine

= Sub-boreal Spruce
may be a useful indicator of litter decomposability.

High polyphenol concentrations in litter are generally

associated with reduced decomposition, mineralization

and nutrient availability (Baldwin and others 1983).

The reductions are thought to result from the inhibition

of microbial activity and the combination of

polyphenols with leaf proteins or soluble forms of

organic nitrogen to produce humic polymer precipitates

that are resistant to microbial attack (Handley 1954;

Haynes 1986; Palm and Sanchez 1991). The influence

of litter quality on rates of decomposition of 1 2

substrates, including needles, leaves, moss, roots,

cones and wood (Taylor and others 1991) are

illustrated in figure 2. Decomposition rate was the most

closely related to lignin content of the litter.

Figure 2—Decomposition of materials of different litter

quality in a Rocky Mountain Pine forest (Taylor and others

1991).

Any forest management activity that affects the

prevailing moisture and temperature conditions or the

nature of the litter produced has the potential to

substantially alter rates of decomposition and N

mineralization in forest floors. This will influence the

amount of organic matter present, the availability of N

and other nutrients, and losses of N from the site

through leaching. Predicting the effects of a given

activity is still difficult due to our imperfect

understanding of the controls on these processes and

the effects of each management activity on site

conditions. Following are results from a series of field

trials in British Columbia designed to test the influence

of several forest management activities on rates of

decomposition and N mineralization.

METHODS

Organic Matter Decomposition

Rates of weight loss of litter were measured using the

litterbag technique, in which known-weight samples are

enclosed in fiberglass mesh bags and incubated on site.

In addition to litter collected at the study site, most

experiments employed standard litters: lodegpole pine

needles and trembling aspen leaves from the

Kananaskis Valley in Alberta (Prescott and others

1989), and forest floor from a coastal forest of western

red cedar, western hemlock and Douglas-fir near

Vancouver (Prescott and Preston 1995). Bags

containing foliar litter were pinned to the surface of the

forest floor; bags containing forest floor material were

buried in the forest floor. Bags were constructed of

fiberglass screening and were 10x10 cm. They were

transported to the site in envelopes and spillage into

the envelope was weighed and subtracted from the

original weight. At annual intervals for 5 years, seven

bags of each type were collected from each plot. The

contents of each bag was dried at 65°C, and the

weight of litter remaining was measured.

Nitrogen Mineralization

Nitrogen mineralization in forest floor material and

mineral soil from plots of each treatment were

measured during incubation in the field. About ten

samples of the forest floor (FH layers) and 10 samples

of the upper 20 cm of mineral soil were collected from

one plot of each treatment. Roots and live vegetation

were removed and the sample was mixed well. A

subsample (about 100g) of each sample was put into a

polyethylene bag, sealed, and inserted into the forest

floor or soil near the place from which it had been

removed. After about 6 weeks, the bags were

retrieved and final concentrations of NH4-N and N03-N

were measured using an Alpkem RFA autoanalyzer

following extraction in 2M KCI (Page and others 1982).

Because initial concentrations of inorganic N were

ususally negligible, the final N concentrations were

used to estimate net N mineralization during the

incubation. For studies of tree species influences, rates

of potential N mineralization and nitrification were

measured by incubating samples of the forest floors of

each species under standard temperature and moisture

conditions in the laboratory.

CLEARCUTTING

Increased availability of N is commonly reported after

forests are clearcut (Bormann and Likens 1979,

Vitousek and others 1979, Smethurst and Nambiar

1990) and has been attributed to faster decomposition

and mineralization of the residual organic matter

126
(Covington 1981 , Kimmins 1987). This in turn has

been attributed to greater microbial activity resulting

from the warmer, moister conditions in clearcuts

(Edmonds and McColl 1989, Frazer and others 1990).

However, studies have shown that decomposition rates

may actually be faster, slower, or the same in clearcuts

compared with forests, depending on the regional

climate (Yin and others 1989). Further, clearcutting

may have different influences on decomposition rates

at different depths in the forest floor (Binkley 1984, Yin

and others 1989).

Experiments to determine the effects of forest removal

on rates of decomposition of litter and forest floors

have been established at 22 sites in several forest

types in British Columbia. Litterbags containing the

same standard substrates were installed in forests and

in adjacent clearcuts. After 3 years, rates of

decomposition of pine needle litter was slower in

clearcuts than in adjacent forests at almost all sites,

and significantly so at 7 of the 16 sites (table 2). This

may be an effect of drier conditions in clearcuts during

the summer, when temperatures are adequate for

decomposition.

Table 2 —Weight of lodgepole pine needle litter remaining after

3 years decomposition in adjacent forest and clearcut sites.

Original weight was 2.0 g.

Zonea

BWBS

CDF

CWH

ESSF

ICH

IDF

SBS

Site

Bear Mt

Shawnigan S

Shawnigan N

Pt McNeill

Blaney Lake

Otter Creek

Spanish Lake

Lucille Mtb

Adams Lake

Malakwa

Hidden Lake

Date Creekb

Valentine Lake

Boston Bar

Topley

Aleza Lakeb

Forest

0.83 (

0.88 (

0.77 (

0.69 (

0.86 (

0.77 (

1.05 (

1.04 (

1.01 (

0.63 (

0.87 (

1.06 (

1.07 (

1.06 (

0.88 (

1.01 (

. 17)

.06)

.11)

. 12)

.21)

.11)

.20)

.23)

.16)

.10)

.10)

. 11)

.11 )

.17)

. 14)

.13)

Clearcut

0.93

1.00

0.85

0.98

1.22

0.79

1.06

1.00

1.16

1.12

1.23

1.11

1.16

1.09

1.26

1.26

(.19)

(.12)c

(.08)

(.14) c

(.03)c

(.09)

(.19)

(.11)

(.14)

(.14)c

(.16)c

(.02)

(.16)

(.07)

(.06)c

(.07)c

abbreviations as in table 1.

2-year data.

clearcut significantly different from forest based on oneway

ANOVA.

ALTERNATIVE SILVICULTURAL

SYSTEMS

Alternative silvicultural systems such as partial cutting

or patch cutting have been proposed as a means of

reducing the negative consequences of clearcutting.

Little is known about the effects of these alternative

systems on organic matter and nutrient dynamics. In

an oak forest, Yin and others (1989) found rates of

decomposition in a shelterwood (15 years after

harvest) to be more similar to those in an uncut forest

than in a clearcut (5 years after harvest). Thinning

forests has been shown to increase (Piene and Van

Cleve 1978), decrease (Weetman 1965) or have no

effect (Will and others 1983) on decomposition rates.

Greater rates of litter decomposition and net N

mineralization were reported in a 0.25 ha patch cut

compared to an adjacent lodgepole pine forest (Prescott

and others 1992b). Parsons and others (1994)

reported increased rates of N mineralization only after

removal of at least 1 5 trees in another lodgepole pine

forest. In a beech forest, Bauhus and Barthel (1995)

found rates of N mineralization higher in 30-m gaps

than in the forest during the first year after cutting, but

lower rates during the second year. Rates of net N

mineralization and decomposition have been measured

in several silvicultural systems trials in British Columbia,

in openings of different sizes, or in patch or partial

cuts. Following are results from two of the trials.

Montane Alternative Silvicultural Systems Trial

In this trial in a montane forest of western hemlock and

amabilis fir on Vancouver Island, treatments were oldgrowth, shelterwood (200 stems ha-1), greentree

retention (25 stems ha-1), 1.5 ha patch cut and a 65 ha

clearcut (Prescott 1997). Rates of N mineralization and

nitrification in the forest floor were greatest in the

clearcut, intermediate in the other silvicultural systems,

and least in the old-growth forest (fig. 3). This

suggests that the alternative silvicultural treatments

have a smaller effect on N mineralization than

clearcutting. The lack of a clear gradient in

concentrations of mineralized N between clearcut,

greentree, shelterwood and old-growth suggests that

rates of N mineralization cannot be predicted from

basal area removed. The clearcut system appears best

for supplying N to regenerating conifers, since it

consistently had higher concentrations of extractable N

in both forest floor and mineral soil. However, the

higher nitrate concentrations in the clearcut indicate

that it is more prone to loss of N through leaching, and

so the other silvicultural systems may be preferable for

maintaining the N capital of the site.

Decomposition rates of foliar litter of hemlock, pine,

aspen and forest floors were measured in plots of the

five silvicultural treatments (old-growth, shelterwood,

patch cut, greetree retention, and clearcut) in a

montane coastal forest (Prescott 1997). After

decomposing for two years, the weight of hemlock and

pine needle litter remaining in the bags was

significantly less in the old growth forest than in the

other treatments (fig. 4). Forest floor material

127
Figure 3 —Net N mineralization from forest floors of a coastal

montane forest during a 63-day field incubation. OG = old

growth, PC = patch cut, SW = shelterwood, GT =

greentree, CC = clearcut. Means for treatments with

different letters are significantly different (p < 0.05)

(Prescott 1997).

Figure 4 —Decomposition of needle litter of western hemlock

(a) and lodgepole pine (b) in a coastal montane forest.

Treatments as described in figure 3. Means for treatments

with different letters are significantly different (p < 0.05).

Sicamous Creek Silvicultural Systems Trial

In a high elevation forest of Engelmann spruce and

subalpine fir in interior British Columbia, rates of net N

mineralization were measured in three plots of each

opening size (uncut control, single-tree, 0.1 ha, 1.0 ha

and 10 ha) during the first and second growing seasons

after harvesting. At each time, 8-12 samples of the

forest floor and of the mineral soil were collected from

and incubated in the center of each opening for about

6 weeks. Rates of N mineralization in the single-tree

removal plots usually did not differ from in the control

plots; those in the 0.1, 1.0 and 10 ha openings were

consistently greater than in the controls (fig. 5). There

was no further increase in mineralization rates in

128

decomposed slowly in all treatments. The overall

tendency was for decomposition to be fastest in the

old-growth followed by the clearcut, and slowest in the

alternative silvicultural treatments.

The faster decomposition in the old-growth forest in

this study suggests that temperature is not the major

factor controlling rates of litter decomposition in this

ecosystem. Average surface temperatures and growing

degree days (>4°C) were substantially lower in the old

growth forest than in any of the silvicultural systems

(Dunsworth and Arnott 1995). Moisture levels in the

forest floor in mid-summer of 1995 were higher in the

old growth forest than in the silvicultural systems,

suggesting that summer moisture may be more critical

for decomposition in this ecosystem. The similarity in

decomposition rates in all silvicultural treatments

indicates that the alternative treatments are similar to

clearcutting in terms of their effect of rates of litter

decomposition. This might suggest that surface

climatic conditions are more similar among the

silvicultural systems than in the old growth; some data

supporting this were presented by Dunsworth and

Arnott (1995). Alternatively, the higher rates of

decomposition in the old growth forest may reflect

more intact communities of soil microorganisms and

fauna in the undisturbed forests. Reductions in soil

microflora and fauna have been reported after

clearcutting in other forests (Blair and Crossley 1988,

Seastedt and Crossley 1981, Sundman and others

1978), including one on Vancouver Island (Vlug and

Borden 1973). Disruption of the forest floor and

reduced root activity in the silvicultural treatments may

have as great of an effect on the activities of soil

microbes and fauna as the microclimatic changes.
Figure 5—Net N mineralization in forest floor (a) and mineral

soil (b) from an Engelmann spruce-subalpine fir forest during a

6-week field incubation. Silvicultural treatments: clearcut (C);

single tree (ST); 0.1,1.0, and 10 ha openings. Means for

treatments with different letters are significantly different (p

< 0.05).

openings greater than 0.1 ha. During the first year,

nitrate formed a very small proportion (<7%) of the

extractable N pool in forest floors, but up to 25% in

mineral soil. Nitrate concentrations were much higher

during the second year, particularly in openings of at

least 0.1 ha, where nitrate accounted for 25-50% of

the available N. Decomposition of pine needles, aspen

leaves and spruce-fir needles were similar in plots of all

opening sizes during the first year (fig. 6). Similar

studies at 5 other silvicultural systems trials in British

Columbia also showed no opening effect.

Greater availability of inorganic N in clearcuts is often

attributed to increased rates of decomposition of

organic material in openings. In these studies, rates of

net N mineralization and nitrification were indeed

greater in the clearcuts, but rates of decomposition

were not. An alternative explanation is the reduced

input of fresh carbon in litter after clearcutting. This

would reduce the microbial biomass and its capacity for

Figure 6 —Decomposition of three litter types in an

Engelmann spruce-subalpine fir forest. Treatments as

described in figure 5. There were no significant differences

among treatments.

immobilizing N, thereby increasing the amount of

extractable NH4. Vitousek and Matson (1985) found

that immobilization of N in microbial biomass was the

major process retaining N in the ecosystem after

clearcutting loblolly pine plantations, and that retaining

residues as substrates for microorganisms increased N

retention. The higher concentrations of NH4 would also

stimulate nitrification, resulting in the higher

concentrations of NO3 after clearcutting. Assimilation

of NO3 would be reduced due to the diminished

microbial biomass (Davidson and others 1992), further

increasing NO3 availability in the clearcut. Hart and

others (1994) reported increased net nitrification

coincident with a decline in microbial biomass during

long-term incubations of a coniferous forest soil. The

sudden increase in NO3 concentrations was attributed

to a decrease in NO3 immobilization by the microbial

biomass as the quality of C substrates declined during

the incubation. A similar situation may exist in forest

floors in clearcuts that are deprived of C inputs in fresh

litter following the removal of trees. Thus, high rates

of N mineralization and leaching in the clearcut may be

related more to changes in C supply (litter inputs) than

alterations in microclimate.

SPECIES

The influence of tree species on soils has been of

interest for decades, but has been difficult to establish

because of the difficulty in separating the influences of

site and species in natural forests. Recent studies in

"common garden experiments" (Binkley 1995), i.e. pure

plantations of different species on one site, have

demonstrated considerable differences in N

mineralization rates in forest floors under different tree

species (Harmer and Alexander 1 986, Binkley and

129
Valentine 1991, Harris and Riha 1991, Gower and Son

1992). Each of these studies has been conducted on a

single site, which makes it difficult to generalize about

individual tree species. However, indications from

these and other studies have suggested that N

availability is relatively high under broadleaves (Stump

and Binkley 1993, Harris and Riha 1991), cedar

(Harmer and Alexander 1986, Turner and Franz 1985)

and larch (Gower and Son 1992, Harmer and Alexander

1986), and low under pine (Miller and others 1979,

Prescott and others 1992a,b) and spruce (Mardulyn

and others 1993, Pastor and others 1987). Gower and

Son (1992) found that differences in N availablity in

the forest floors under 5 tree species were closely

correlated with the lignin:N content of the litter, which

in turn is closely related to its rate of decomposition.

This suggests that the nutritional environment of a tree

is largely determined by, and therefore can be predicted

from, the resource quality of its litter and the rate at

which it decomposes.

The influence of tree species on N availability has been

examined by measuring rates of net N mineralization in

forest floors from plots of pure species during 30-day

laboratory incubations. In the first set of experiments

(Prescott and others 1995), three species trials were

examined: 35-year-old plantations of western red

cedar, western hemlock and Douglas-fir at the UBC

Research Forest near Vancouver, 25-year-old plots of

lodgepole pine, Douglas-fir and paper birch near

Skimikin in interior British Columbia, and 38-year-old

plantations of 13 species at Clonsast Bog in Ireland.

There were some consistent patterns in relative rates of

net N mineralization of tree species in the three trials

(table 3). The relative rates of the three species in the

University of British Columbia trial (Douglas-fir >

western hemlock > western red cedar) were the same

in the Clonsast Bog trial. Douglas-fir also had the

highest rate of N mineralization of the three species in

the Skimikin trial. We might expect, therefore, to find

relatively high rates of N mineralization in forest floors

under Douglas-fir, and relatively low rates under

western red cedar. High rates of net N mineralization

and nitrification in Douglas-fir forest floors have been

observed in several other trials in British Columbia

(Prescott 1996). Other studies have also suggested

that N mineralization under cedar is relatively low

(Prescott and Preston 1994). The slow mineralization

in pine plots is in keeping with low availability of N

reported in other pine forests (Fahey and others 1985,

Miller and others 1979, Prescott and others 1992a).

The observation of high rates of N mineralization in

forest floors under Douglas-fir and low under western

red cedar (Prescott and Preston 1994) was tested by

measuring rates of N mineralization in forest floors

under cedar, hemlock, spruce and Douglas-fir in

replicated experiments at four sites in coastal British

Columbia. As shown in figure 7, there was very little N

mineralized in forest floors from two sites, but large

amounts from the other two sites. At UK Main and

San Juan, rates were highest in cedar plots, followed

by Douglas-fir plots. The large between-site

differences in N mineralization were not related to

characteristics of the mineral soils. The sites with low

rates of N mineralization were located at higher

elevations on mid-slopes and the ground vegetation

was dominated by salal; the sites with high rates of N

mineralization were located at the base of slopes and

had vegetation more characteristic of moist sites, such

as salmonberry. Due to slow growth of cedar relative to

other species on all sites, cedar plots had more open

canopies and greater densities of ground vegetation.

Rates of N mineralization in the forest floors appear

therefore to be influenced by site moisture conditions

and related composition of the understory vegetation;

these site influences make it difficult to generalize

about the influence of tree species.

Foliar litter represents the largest single source of N to

be recycled within a forest, and as such its "quality"

and resulting rate of decomposition will influence N

availability in the forest. Rates of decomposition of

foliar litter are often related to their initial C:N, lignin or

lignin:N contents. Rates of decompositon of foliar litter

of 10 species of trees in British Columbia are being

measured at two locations, in the boreal forest and in

the coastal forest, and related to initial "quality". Rates

of weight loss during the first 3 years are shown in

figure 8. Western red cedar, firs and western hemlock

decomposed more slowly than other species. The

strongest relationship was with % lignin (r2 = 0.31),

which ranged from 18% in Engelmann spruce to 38%

in cedar (fig. 9).

SPECIES MIXTURES

Recent interest in management of mixedwood forests

in British Columbia has highlighted our limited

understanding of the interactions among species in

mixed forests, particularly as they influence

belowground processes. Studies in other forests have

demonstrated that having seral stages or interplanting

with other species, particularly hardwoods, can

significantly improve growth of regenerating conifers,

partly through improved nutrition of trees in mixed

forests. In the U.K., Sitka spruce in mixtures with

larch or pine grow significantly faster than those in

single species plantations, and this has been linked to

greater availability of N in the forest floor of mixed

plantations (Carlyle and Malcolm 1986). Williams and

Alexander (1991) suggested that the improved N

availability in mixed plantations was at least partly due

to more complex fungal-faunal interactions in mixed

plantations, which increased rates of decomposition

and N release from litter. Development of hardwood

130
Table 3 —Rates of N and C mineralization during laboratory incubations of forest floors from adjacent plantations of different trees species in three trials.

Species and Trial

UBC Forest

Douglas-fir

Western hemlock

Western red cedar

Skimikin

Douglas-fir

Paper birch

Lodgepole pine

Clonsast Bog

Sissile oak

Sitka spruce

Grand fir

Douglas-fir

Norway spruce

Lodgepole pine

Japanese larch

S. spruce + J. larch

Western hemlock

Scots pine

Monterey pine

Calluna

Western red cedar

N Mineralization

μg N g-1 d-1

5.5 a

3.2 ab

1.6 b

45.9 a

20.5 ab

15.7 b

226.3 a

151.3 b

151.1 bc

138.6 bc

127.0 bc

125.7 bcd

123.7 cde

98.0 cde

89.0 cdef

73.7 def

52.7 efg

41.0 fg

12.3 g

C mineralization

μ gcg -

191.4

202.1

144.5

415.2

406.5

820.3

355.2

313.8

234.9

282.6

239.0

195.3

279.0

216.1

204.0

385.4

138.1

446.5

409.4

'd-1

a

a

b

b

b

a

a bc

bcd

def

cde

def

ef

cde

def

ef

abc

f

a

ab

PH

5.2

4.2

5.1

5.8

6.4

4.9

4.7

4.8

5.0

4.4

4.9

3.9

5.7

4.3

3.9

4.7

4.1

4.6

6.1

a

b

a

b

a

c

bc

bc

b

cde

b

e

a

cde

e

bc

de

bcd

a

Note: Each value is the mean of 14 (UBC Forest), 24 (Skimikin), or 10 (Clonsast Bog) samples. Within each

trial, values for different species followed by the same letter are not significantly different (p>0.05), based

on oneway ANOVA and Sheffe's multiple range tests.

Figure 8 —Decomposition of foliar litter of ten conifer species

in a boreal spruce forest.

of aspen and poplar and subsequent faster decomposi

tion (Pare and Van Cleve 1993). Conifer litter often

decomposes more rapidly when mixed with hardwood

131

Figure 7 —Net N mineralization in forest floors of four tree

species at each of four sites in coastal BC during a 4-week

laboratory incubation.

clumps after clear-cutting white spruce in the boreal

forest contributes to acceleration of nutrient cycling,

due to higher concentrations of N and P in foliar litter
Figure 9 —Relationship between decomposition rate (k value)

and % lignin (acid insoluble fraction) of foliar litter of ten

coniferous species.

leaf litter (Taylor and others 1989b), suggesting that

decomposition of conifer litter may be faster in mixed

plots. The influence of mixing litters of either

hardwoods or conifers have been less consistent (Blair

and others 1990).

Experiments have been established to determine the

effects of mixing litter in the three major mixedwood

forest types in British Columbia: boreal (white

spruce/trembling aspen), interior wet-belt (Douglasfir/paper birch/lodgepole pine) and coast (Douglas-fir/

red alder). Rates of decay of pure and mixed litter of

each species is compared during incubation in pure and

mixed forests. After decomposing for four years at

four sites in the boreal forest, there were no significant

differences between pure and mixed litter of spruce

and aspen (fig. 10). After decomposing for 2 years in a

coastal forest, there was no difference in rates of

decomposition of pure and mixed alder and Douglas-fir.

Mixtures of lodgepole pine, birch and Douglas-fir

decomposed more slowly than pure litters.

FERTILIZATION

It is generally assumed that increasing N availability

through fertilization will increase the rates of litter

decomposition. However, direct studies of the effects

of N fertilization on decomposition rates have had

variable results, and indicate that external N supply has

little effect on decay rate (Hunt and others 1988,

Prescott 1994). The effect of higher N concentrations

in litter of a single species on decay rates is also

unclear. Berg and others (1987) found that greater N

availability in pine needles stimulated decay in the early

stages but inhibited decay during the later lignin decay

phase, but Prescott (1994) found no influence.

Prescott and others (1992b) suggested that decay of

litter with low lignin and high labile C contents may be

Figure 10—Decomposition of pure and mixed foliar litter of

trembling aspen and white spruce in a boreal spruce forest.

Means for treatments with different letters are significantly

different (p < 0.05).

stimulated by fertilization, whereas litter with high

lignin and low labile C contents will not.

Two experiments have been established, one in a

trembling aspen stand in northern British Columbia

fertilized once with N or NPK, and one in a Douglas-fir

stand in coastal Washington fertilized repeatedly with

N. At each site, litter from control and fertilized plots

is decomposing. In both trials, litter from fertilized

plots had higher initial concentrations of N than litter

from control plots. After two years, Douglas-fir litter

was decomposing more slowly in fertilized plots than in

control plots, but by year four there was no difference

in rate of decomposition (fig. 11). There was no

apparent influence of needle source (i.e. endogenous N

concentration). There was no effect of either higher N

concentrations in the litter or in the surrounding forest

floor on rates of decomposition of aspen litter (fig. 12).

CONCLUSIONS

Clearcutting

• increased net N mineralization and nitrification

• decomposition of pine litter same or slower than in

uncut forests

Alternative silvicultural systems

• increased net N mineralization and nitrification in

openings greater than 0.1 ha

• no effect on decomposition rates

132
ACKNOWLEDGMENTS

The experiments were funded by Forest Renewal British

Columbia, the British Columbia Ministry of Forests and

the Natural Sciences and Engineering Research Council

of Canada. The authors thank C. Staley for field and

laboratory work and L. Blevins and B. Taylor for

assistance with data analysis.

LITERATURE CITED

Baldwin I.T.; Olson R.K.; Reiners W.A. 1983. Protein

binding phenolics and inhibition of nitrification in

subalpine balsam fir soils. Soil Biology and

Biochemistry 15:419-423.

Baldock, J.A.; Preston, CM . In: Proceedings of 8th N.

American Forest Soils Conference. August, 1993.

Gainesville, USA.

Bauhus, J.; Barthel, R. 1995. Mechanisms for carbon

and nutrient release and retention within beech

forest gaps. 2. The role of soil microbial biomass.

Plant and Soil 168-169:585-592.

Berg, B.; Staff, H.; Wessen, B. 1987. Decomposition

and nutrient release in needle litter from nitrogenfertilized Scots pine (Pinus sylvestris) stands.

Scandinavian Journal of Forest Research 2:399-

415.

Binkley, D. 1984. Does forest removal increase rates of

decomposition and nitrogen release? Forest Ecology

and Management 8:229-233.

Binkley, D. 1995. Species influence on forest soils -

processes and patterns. In Proceedings of the Trees

133

Figure 11 —Decomposition of fertilized and control needle

litter of Douglas-fir incubated in fertilized and control plots.

Means for treatments with different letters are significantly

different (p < 0.05).

Figure 12—Decomposition of trembling aspen leaf litter from

fertilized and control plots (a), or incubated in control and

fertilized plots (b) in a boreal aspen forest. There were no

significant differences between treatments at any sampling

time.

Species

• net N mineralization and nitrification high under

Douglas-fir, low under western red cedar

• decomposition fastest in broadleaves, slowest in

cedar

• decomposition related to lignin content of litter

• influence of tree species confounded by site factors,

understory vegetation

Species mixtures

• no effect of mixing litter on decomposition rates

Fertilization

• no effect on decomposition rates
and Soil Workshop, Lincoln University, Christchurch,

New Zealand. Feb. 28 - Mar. 2, 1994.

Binkley, D.; Valentine, D. 1991. Fifty-year

biogeochemical effects of green ash, white pine,

and Norway spruce in a replicated experiment.

Forest Ecology and Management 40:13-25.

Blair, J.M.; Crossley, D.A. 1988. Litter

decomposition, nitrogen dynamics and litter

microarthropods in a southern Appalachian

hardwood forest 8 years following clearcutting.

Journal of Applied Ecology. 25:683-698.

Blair, J.M.; Parmalee, W.; Baere, M.H. 1990. Decay

rates, nitrogen fluxes, and decomposer communities

of single- and mixed-species foliar litter. Ecology 71:

1976-1985.

Bormann, F.H.; Likens, G.E. 1979. Pattern and

process in a forested ecosystem. Springer-Verlag,

New York.

Bunnell, F.L.; Tait, D.E.N.; Flanagan, P.W.; Van Cleve,

K. 1976. Microbial respiration and substrate

weight loss. 1. A general model of the influences of

abiotic variables. Soil Biology and Biochemistry

9:33-40.

Carlyle, J.C.; Malcolm, D.C. 1986. Nitrogen

availability beneath pure spruce and mixed larch +

spruce stands growing on deep peat. 1. Net N

mineralization measured by field and laboratory

incubations. Plant and Soil 93:95-1 1 3.

Covington, W.W. 1981. Changes in forest floor

organic matter and nutrient content following clear

cutting in northern hardwoods. Ecology 62:41-48.

Davidson, E.A.; Hart, S.C.; Firestone, M.K. 1992.

Internal cycling of nitrate in soils of a mature

coniferous forest. Ecology 73:1148-1156.

Dunsworth, B.G.; Arnott, J.T. 1995. Growth

limitations of regenerating montane conifers in field

environments, pp. 41-47 In: Arnott, J.T.; Beese,

W.J.; Mitchell, A.K.; Peterson, J. Montane

Alternative Silviculture Systems (MASS).

Proceedings of the MASS workshop , June 7-8,

1995, Courtenay, B.C. FRDA Report 238. Canadian

Forest Service, Victoria, B.C. 122 p.

Edmonds, R.L.; McColl, J.G. 1989. Effects of forest

management on soil nitrogen in Pinus radiata stands

in the Australian Capital Territory. Forest Ecology

and Management 29:199-212.

Fahey, T.J.; Yavitt, J.B.; Pearson, J.A.; Knight, D.H.

1985. The nitrogen cycle in lodgeole pine forests,

southeastern Wyoming. Biogeochemistry 1:257-

275.

Frazer, D.W.; McColl, J.G.; Powers, R.F. 1990. Soil

nitrogen mineralization in a clearcutting

chronosequence in a northern California conifer

forest. Soil Science Society of America Journal

54:1145-1152.

Gower, S.T.; Son, Y. 1992. Differences in soil and

leaf litterfall nitrogen dynamics for five forest

plantations. Soil Science Society of America Journal

56:1959-1966.

Handley W.R.C. 1954. Mull and mor formation.

Forestry Commission Bulletin No. 22. H.M.S.O.

Harmer, R.; Alexander, I. 1986. The effect of starch

amendment on nitrogen mineralization from the

forest floor beneath a range of conifers. Forestry

59:39-46.

Harris, M.M.; Riha, S.J. 1991. Carbon and nitrogen

dynamics in forest floor during short-term laboratory

incubations. Soil Biology and Biochemistry 23:1035-

1041.

Hart, S.C.; Nason, G.E.; Myrold, D.D.; Perry, D.A.

1994. Dynamics of gross nitrogen transformations

in an old-growth forest: the carbon connection.

Ecology 75:880-891.

Haynes R. J. 1986. The decomposition process:

mineralisation, immobilisation, humus formation and

degradation. In Mineral nitrogen in the plant-soil

system. Eds. R. J. Haynes Academic Press,

Orlando, Florida.

Hunt, H.W.; Ingham, E.R.; Coleman, D.C; Elliott, E.T.;

and Reid, C.P.P. 1988. Nitrogen limitation of

production and decomposition in prairie, mountain

meadow and pine forest. Ecology 69:1009-1016.

Kimmins, J.P. 1987. Forest Ecology. MacMillan, New

York.

Klinka, K.; Pojar, J.; Meidinger, D.V. 1991. Revision

of biogeoclimatic units of coastal British Columbia.

Northwest Science 65:32-47.

Krajina, V.J.; Madoc-Jones, S.; Mellor, G. 1973.

Ammonium and nitrate in the nitrogen economy of

some conifers growing in Douglas-fir communities

of the Pacific Northwest of America. Soil Biology

and Biochemistry 5:143-147.

Mardulyn, P.; Godden, B.; Amiano Echezarrets, P.;

Penninckx, M.; Gruber, W.; Herbauts, J. 1993.

Changes in humus microbiological activity induced

134
by the substitution of the natural beech forest by

Norway spruce in the Belgian Ardennes. Forest

Ecology and Management 59:15-27.

Meentemeyer, V. 1978. Macroclimate and lignin

control of litter decomposition rates. Ecology

59:465-472.

Miller, H.G.; Cooper, J.M.; Miller, J.D.; Pauline, O.J.L.

1979. Nutrient cycles in pine and their adaptation

to poor soils. Canadian Journal of Forest Research

9:19-26.

Palm C.A.; Sanchez, P.A. 1991. Nitrogen release from

the leaves of some tropical legumes as affected by

their lignin and polyphenolic contents. Soil Biology

and Biochemistry 23(1):83-88.

Page, A.L.; Miller, R.H.; Keeney, D.R. (eds.). 1982.

Methods of soil analysis. Part 2. Chemical and

microbiological properties. American Society of

Agronomy, Madison, Wisconsin.

Pare, D.; Van Cleve, K. 1993. Soil nutrient availability

and relationships with aboveground biomass

production on postharvested upland white spruce

sites in interior Alaska. Canadian Journal of Forest

Research 23:1223-1232.

Parsons, W.F.J.; Knight, D.H.; Miller, S.L. 1994. Root

gap dynamics in lodgepole pine forest: nitrogen

transformations in gaps of different size. Ecological

Applications 4:354-362.

Pastor, J.; Gardner, R.H.; Dale, V.H.; Post, W.M.

1987. Successional changes in nitrogen availability

as a potential factor contributing to spruce declines

in boreal Noth America. Canadian Journal of Forest

Research 17:1394-1400.

Piene, H.; Van Cleve, K. 1978. Weight loss of litter

and cellulose bags in a thinned white spruce forest

in interior Alaska. Canadian Journal of Forest

Research 8:42-46.

Prescott, C.E. 1994. Does nitrogen availability control

rates of litter decomposition in forests? Plant and

Soil 168:83-88.

Prescott, C.E. 1996. Influence of forest floor type on

rates of litter decomposition in microcosms. Soil

Biology and Biochemistry 28:1319-1325.

Prescott, C.E. 1997. Effects of clearcutting and

alternative silvicultural systems on rates of

decomposition and nitrogen mineralization in a

coastal montane coniferous forest. Forest Ecology

and Management 95:253-260.

Prescott, C.E.; Corbin, J.P.; Parkinson, D. 1989.

Biomass, productivity, and nutrient-use efficiency of

aboveground vegetation in four Rocky Mountain

coniferous forests. Canadian Journal of Forest

Research 19:309-317.

Prescott, C.E.; Corbin, J.P.; Parkinson, D. 1992a.

Availability of nitrogen and phosphorus in the forest

floors of Rocky Mountain coniferous forests.

Canadian Journal of Forest Research 22:593-600.

Prescott, C.E.; Corbin, J.P.; Parkinson, D. 1992b.

Immobilization and availability of N and P in the

forest floors of fertilized Rocky Mountain coniferous

forests. Plant and Soil 143:1-10.

Prescott, C.E.; Preston, CM . 1994. Nitrogen

mineralization and decomposition in forest floors in

adjacent plantations of western red cedar, western

hemlock, and Douglas-fir. Canadian Journal of

Forest Research 24:2424-2431.

Prescott, C.E.; Thomas, K.D.; Weetman, G.F. 1995.

Nitrogen mineralization in forest floor under adjacent

plantation of different tree species: lessons from

thre retrospective studies. Proc. Trees and Soil

Workshop, Lincoln University, Canterbury, New

Zealand, February 28 - March 2, 1994.

Read, D.J. 1983. The biology of mycorrhizae in the

Ericales. Canadian Journal of Botany 61:985-1004.

Rygiewicz, P.T.; Bledsoe, C.S; Zazoski, R. 1984.

Effects of ectomycorrhizae and solution pH on 1 5N

ammonium uptake by coniferous seedlings.

Canadian Journal of Forest Research 14:885-892.

Seastedt, T.R.; Crossley, D.A. Jr. 1981.

Microarthropod response following cable logging

and clear cutting in the southern Appalachians.

Ecology 62:126-135.

Smethurst, P.J.; Nambiar, E.K.S. 1990. Distribution

of carbon and nutrients and fluxes of mineral

nitrogen after clearfelling in a Pinus radiata

plantation. Canadian Journal of Forest Research

20:1490-1497.

Smirnoff, N.; Todd, P.; Stewart, G.R. 1984. The

occurrence of nitrate reductase in the leaves of

woody plants. Annals of Botany 54:363-374.

Stump, L.M.; Binkley, D. 1993. Relationships

between litter quality and nitrogen availability in

Rocky Mountain forests. Canadian Journal of Forest

Research 23:492-502.

135
Sundman, V.; Huhta, V.; Niemela, S. 1978. Biological

changes in northern Spruce forest soil after clearcutting. Soil Biology and Biochemistry 10:393-397.

Taylor B.R.; Parkinson, D.; Parsons, W.F.J. 1989a.

Nitrogen and lignin content as predictors of litter

decay rates: a microcosm test. Ecology 70:97-104.

Taylor B.R.; Parsons, W.F.J.; Parkinson, D. 1989b.

Decomposition of Populus tremuloides leaf litter

accelerated by addition of Alnus crispa litter.

Canadian Journal of Forest Research 19:674-679.

Taylor, B.R.; Prescott, C.E.; Parsons, W.F.J.;

Parkinson, D. 1991. Substrate control of litter

decomposition in four Rocky Mountain coniferous

forests. Canadian Journal of Botany 69:2242-2250.

Turner, D.P.; Franz, E.H. 1985. The influence of

western hemlock and western redcedar on microbial

numbers, nitrogen mineralization, and nitrification.

Plant and Soil 88:259-267.

Turner, D.P.; Sollins, P.; Leuking, M.; Rudd, N. 1993.

Availability and uptake of inorganic nitrogen in a

mixed old-growth coniferous forest. Plant and Soil

148:163-174.

Vitousek, P.M.; Gosz, J.R.; Grier, C.C.; Melillo, J.M.;

Reiners, W.A.; Todd, R.L. 1979. Nitrate losses

from disturbed ecosystems. Science 204:469-474.

Vitousek, P.M.; Matson, P.A. 1985. Disturbance,

nitrogen availability, and nitrogen losses in an

intensively managed loblolly pine plantation. Ecology

66:1360-1376.

Vlug, H.; Borden, J.H. 1973. Soil Acari and

Collembola populations affected by logging and

slash burning in a coastal British Columbia

coniferous forest. Environmental Entomology

2:1016-1023.

Weetman, G.F. 1965. The decomposition of confined

black spruce needles on the forest floor. Pulp &

Paper Research Institute of Canada. Woodlands

Research No. 165.

Will, G.M.; Hodgkiss, P.D.; Madgwick, H.A.I. 1983.

Nutrient losses from litterbags containing Pinus

radiata litter: influences of thinning, clearfelling and

urea fertiliser. N. Zeal. J. For. Sci. 13:291-304.

Williams, J.; Alexander, C.E. 1991. Interactions on

Mixing Litters from beneath Sitka spruce and Scots

pine and the effects on microbial activity and Nmineralization. Soil Biology and Biochemistry

21:183-188.

Yin, X.; Perry, J.A.; Dixon, R.K. 1989. Influence of

canopy removal on oak forest floor decomposition.

Canadian Journal of Forest Research 19:204-214.

136
TREE SPECIES COMPOSITION CHANGE - SOIL ORGANISM INTERACTION:

POTENTIAL EFFECTS ON NUTRIENT CYCLING AND CONSERVATION IN INTERIOR

FORESTS

Alan E. Harvey1, Russell T. Graham2 and Geral I. McDonald3

ABSTRACT

Current conditions in western Interior forests show large-scale changes in species compositions and accompanying

above- and below-ground structures and nutrient distribution. Typical forests occupy constantly changing, often

resource limited environments. Biological decomposition does not keep pace with production and natural wildfire is

critical to recycling. Appropriate protection and recycling of nutrients is key to both short- and long-term productivity

and sustainability. Comparisons of physical, physiological and genetic natures of the species dominating historic

(pines/larch), current and likely future forests (firs/cedar/hemlock) indicate microbe/fire codependant processes

mediating adaptation of forest vegetation, nutrient use rates, soil nutrient storage and recycling processes will all be

significantly altered. This is likely to produce forests that are drought, pest, competition and change intolerant with

high mortality rates, and high fuel accumulation/loss potentials, when compared to historic norms. Prudent future

management would emphasize reintroduction of appropriate pine stocks and western larch at every opportunity!

Control of density in dry forests can also make an important contribution.

Keywords: forest soils, forest health, forest fire, decomposition, ectomycorrhizae, nitrogen fixation, soil organic matter,

soil wood, long-term productivity, sustainability

INTRODUCTION

Widespread interest in and much reporting on the

"forest health crises," or lack thereof, in the Interior

West has occurred since the late 1980s. Opinions vary

as to the relative threat of changes in inland forests

this century, but there is widespread agreement that

significant changes have indeed occurred (Baker 1988,

Brown 1983).

For the most part, changes are characterized by a

general shift from open pine (Pinus ponderosa Laws.) to

closed pine and/or Douglas-fir (Pseudotsuga menziesii

[Mirb.] Franco) stands in dry ecosystems (Gast and

others 1991, Mutch and others 1993, Wickman 1992).

In moist forests, the change has been from tall,

moderately closed pine/larch (Pinus monticola Dougl.

ex. D. Don/Larix occidentalis Nutt.) to relatively short,

closed grand fir/hemlock/cedar (Abies grandis [Dougl.

ex. D. Don.]/Tsuga heterophlla [Raf.] Sarg./Thuja

plicata Donn) stands (Byler and others 1994, Harvey

and others 1994b, O'Laughlin and others 1993, Moeur

1992, Monnig and Byler 1992).

Common to these reports is an observed shift toward

late successional species in all ecosystems and a view

that insect-pathogen activities are generally increasing

in many (about half [Quigley and others 1996]) of the

resulting stands (Hessburg and others 1994, Lehmkuhl

and others 1994, Sampson and others 1994).

However, since half still exhibit relatively low pest

activities, it has lead to much "active discussion" of the

"state of health" and function (positive vs negative

effects) of endemic "pests" in these forests (Haak and

Byler 1993, Harvey 1994, Martin 1988, Showalter

1994).

Variable fire regimes, successional processes and

endemic insect and pathogen activities generally

worked in concert to produce the stable and resilient

forest structures and species compositions of the past

(Harvey 1994, Harvey and others 1992, Martin 1988).

However, it is clear that exotics, like white pine blister

rust (Cronartium ribicola J.C. Fisch.), are producing

destabilizing impacts in most interior forests (Byler and

others 1994, Harvey and others 1994b, Monnig and

Byler 1992, Hoff and Hagle 1990)!

1Alan E. Harvey, Supervisory Plant Pathologist for the Rocky Mountain Research Station, Forestry Sciences Laboratory, 1221 S.

Main St. at Moscow, Idaho 83843-4211.

2Russell T. Graham, Research Forester, Rocky Mountain Research Station, Forestry Sciences Laboratory, 1221 S. Main St. at

Moscow, Idaho 83843-4211.

3Geral I. Mcdonald, Research Plant Pathologist, Rocky Mountain Research Station, Forestry Sciences Laboratory, 1221 S. Main St.

at Moscow, Idaho 83843-4211.

137
CURRENT TRENDS

In effect, the changes in vegetation in both dry and

moist ecosystems of the Inland West have compressed

successional processes. By a factor of at least 10 in

some locations! In dry ecosystems, where ponderosa

pine may (or may not) give way to Douglas-fir in 300-

400 years with historically frequent fire regimes, the

change occurred in only 40 years on the Boise Basin

Experimental forest in southern Idaho (see Sloan in

Harvey and others 1994b). Similarly, In western white

pine dominated ecosystems with mixed and variable

fire regimes, seral species may (or may not, depending

on fire history) give way to climax species in 200-300

years. That conversion took place in just 30 years at

Deception Creek Experimental Forest in northern Idaho

(Moeur 1992).

This time compression process is usually most evident

in dry forests where lack of frequent fires, stand

conversions, and high densities are obvious (Covington

and others 1994). But lack of fire is also a major force

for change in moist forests where effects are more

subtle and have been confused to some extent by

aggressive harvesting of seral species plus the impacts

of the blister rust fungus the last century (Quigley and

others 1996, Monnig and Byler 1992).

VEGETATION PARAMETERS

Associated with the rapid change from serai to climax

dominated stands are significant alterations of the

above-ground, structural geometry of individual trees

and, thus, the stands and landscapes that support

them. Perhaps the most striking, at least in dry

forests, is the increased number of stems relative to

historical forests (Covington and others 1994, Gast and

others 1991, Baker 1988). However, other changes

are no less significant or less important, if visually less

striking. Some are even more widespread and are

often typical of both dry and moist forest

environments. For example, canopy height and shape,

foliar dispersion, density (flammability) and nutrient

content are all changed in both dry and moist

ecosystems.

Ponderosa pine, western white pine and western larch

all tend to be tall and self prune well, even in

moderately dense stands. They also carry large

branches high in their crowns and have good foliar

dispersion (Minore 1979). Foliage and small branch

wood from western white pine and western larch is of

particularly low bulk density (and nutrient content)

compared to most climax species (Minore 1979).

Consequently, stands of pines (except lodgepole pine

[Pinus contorta Dougl. ex Laud.) and western larch

carry their canopies well above surface fuels, tend not

to have ladder fuels, do not carry crown fires well, and

generally have low nutrient content. Thus, canopy

nutrient stores are relatively well protected from fire.

In comparison, intermediate-age Douglas-fir, true firs,

western hemlock and western red cedar are not tall, do

not self prune well (especially when young), carry large

branches low in their canopies (strong conical shape),

do not have good foliar dispersion, and have relatively

high bulk densities with resulting high nutrient

concentrations in foliage and small branchwood (Minore

1979, Brown 1978, Van Wagner 1977). So, stands

dominated by these species often have ladder fuels,

carry crown fires well and have a generally high

nutrient content in tissues susceptible to loss from fire.

Below-ground there are similar changes. Pines and

larch tend to be deep rooted while the firs, hemlock

and cedar all tend to be relatively shallow rooted with

high levels of feeder root development (and

mycorrhizae) in the high nutrient content, shallow soil

organic layers (Gale and Grigal 1987, Harvey and

others 1986, Minore 1979). Douglas-fir also shows

these trends when compared to ponderosa pine but

less so than with other climax species (Minore 1979).

Thus, nutrients and nutrient turnover tend to be

dispersed vertically in soils dominated by seral species

and concentrated in shallow soil horizons with climax

species where fire cycles approximate historical

conditions. Feeder roots of all plants tend to

accumulate in the shallow, high resource organic

horizons (Marschner and Marschner 1996, Robinson

1994). So, soil surface nutrient stores and feeder roots

in climax species dominated stands (or seral stands if

surface organics accumulate from fire exclusion) can be

more at risk to wildfire, or other disturbance, than

those in historic pine and/or larch dominated stands.

This risk can be magnified, especially on infertile sites,

because in dense stands climax species may be much

more demanding of resources (nutrients and water)

than the historically open stands of seral species and,

in some species combinations, they may also be less

tolerant of short supply (Minore 1979).

Even more subtle, but no less important, are

differences in the overall genetic nature of these

forests. Western white pine, ponderosa pine, western

larch and western red cedar all tend to be broadly

adapted to a wide range of sites and environments, the

others tend to be more finely tuned within their

respective ranges of sites and environments (Rehfeldt

1994). Western white pine is particularly well suited to

variation in site and environment.

Sress based on site resource limitations, competition

and short- or long-term environmental variation, is

therefore much more likely in climax species stands

138
(except for western red cedar) than with seral species

stands. Similarly, tolerance to disturbances within the

stand, or with climatic variations or trends that affect

the stand externally, can be expected to cause rapid

responses {stress) with climax and low or no responses

in stands dominated by seral species.

MICROBIAL INTERACTIONS

It should not be surprising that seral species are far

more tolerant of endemic insects and pathogens than

climax species (Harvey 1994, Harvey and others

1994a) or that climax stands should tend toward high

insect- and pathogen-mediated mortality, i.e.,

aggressive and continuous removal of highly stressed

individual trees by endemic "pests," perhaps most

notably by aggressive root pathogens (Harvey 1994,

McDonald 1991). Thus, activities of endemic

organisms tend to "rogue," climax species and stressed

trees, encouraging domination of historical ecosystems

by serai species with intermixed "best adapted"

individuals representing climax species.

Exotic pests are an exception. Despite a wide latitude

for tolerating variation in sites, environments and

activities driven by endemic microbes, especially root

pathogens, western white pine is heavily damaged by

the imported white pine blister rust. In a very real

sense, arrival of the blister rust fungus changed the

rules in white pine dominated ecosystems (Monnig and

Byler 1992). Microbe-mediated selection is now

discouraging rather than supporting historical forest

compositions on western white pine sites. Also, since

insects and pathogens are now operating in

ecosystems dominated by susceptible species under

stress, they generate momentum and are likely to

become much less selective in their actions than in

historical forests (Harvey 1994, Lehmkuhl and others

1994).

SOIL MICROBE-MEDIATED NUTRIENT

CYCLING

Most striking of the changes in microbial actions in pine

vs fir forests are differences in vertical distribution of

ectomycorrhizal short roots in dry ponderosa pine

compared to dry Douglas-fir or moister ecosystems

(Harvey and others 1986). Such differences are not

evident with western white pine compared to western

hemlock, despite the relatively deeper root system of

the pine (fig. 1). Ponderosa pine carries a much greater

proportion of its ectomycorrhizal activities in mineral

soil vs shallow organic horizons. Douglas-fir, western

white pine and western hemlock dominated

ecosystems all have most of their ectomycorhizae

concentrated in surface organic horizons.

Thus, ectomycorrhizal-mediated nutrient acquisition and

turnover are relatively well protected from wildfire

damage only in relatively natural ponderosa pine

dominated systems. This seems quite appropriate

considering the participation of frequent fire in the

nutrient cycling processes of historical ponderosa pine

ecosystems.

However, with the shift to fir-type forests, wildfires

can be expected to have a much larger impact on

ectomycorrhizal-mediated, post-fire acquisition and

cycling of nutrients. It may also increase the likelihood

of uncoupling any continuity of current with preceding

vegetative communities by altering critical soil microbial

legacies, especially ectomycorrhizae (Amaranthus and

Perry 1994).

Another soil-based change in pine vs fir ecosystems is

the relative distribution of nutrient storage, especially

nitrogen. There is a gradual shift in percent of soil

nitrogen reserves and organic matter in mineral layers

of pine forests to surface organic horizons in fir or

hemlock forests (fig. 2). Similarly, the microbe-driven

non-symbiotic nitrogen fixation process fueled by

carbon from organic horizons reflects these same

changes (Jurgensen and others 1997). As with roots

and mycorrhizae, vertical distribution of nitrogen

storage and fixation are compressed upwards. Again,

dry pine forests tend to have critical nutrient-related

processes better protected from surface disturbances

than fir or hemlock forests.

Accumulation of both above- and below-ground

biomass from roots, canopies and boles in dense firtype forests is greatly accelerating activities of

decomposers by increasing and changing basic

substrates, both live and dead. An inescapable dictum

of interior forests is that biological decomposition

139

Figure 1 —Comparisons in distribution of ectomycorrhizal

activity in mineral versus organic soil horizons of pine and

fir/hemlock-dominated ecosystems from dry and moist forests

typical of the Inland West (Harvey and others 1986).
Figure 2 —Comparisons of nitrogen distribution in mineral

versus organic soil horizons of ponderosa pine, Douglas-fir and

hemlock dominated forests typical of the Inland West

(Jurgensen and others 1997).

(respiration-driven) is more limited than biological

production (photosynthesis-driven). So, rapid

accumulations are inevitable (Harvey 1994, Harvey and

others 1994a).

The net result can be accumulations of materials that

differ substantially from those of historical pine-type

forests. Increasing canopy-derived litter and more

prolific formation and death of fine roots (and

mycorrhizae) will change the type and can greatly

increase the quantity of organic matter at the soil

surface (Edmonds 1991, Minore 1979).

Associated with changes in litter type and quantity is a

likely change in soil surface chemistry, including

allelopathic substances with potential to alter a wide

variety of microbial activities, including

ectomycorrhizae (Rose and others 1983). A change of

dominant tree species also provides both live and dead

stem wood that is far more vulnerable to internal decay

and defect, and the increased mortality that

accompanies activities of the causal organisms

(Wagener and Davidson 1954).

Will this provide an increased supply of important

coarse woody debris which can add to site fertility and

wildlife habitat? Likely not, at least in many cases.

Most of the wood being added to interior forests (grand

fir, white fir [Abies concolor Gord.&Glendl. Lindl. Ex

Hildebr.], subalpine fir [Abies lasiocarpa Hook. Nutt.],

western hemlock, western red cedar) is prone to whiterot. Douglas-fir and western white pine are exceptions!

White rot end products are dispersed in soil relatively

rapidly (Larsen and others 1979). Brown rotted wood

on the other hand (i.e., the end products from

decomposition of pine, larch and Douglas-fir) is

persistent in forest soils for at least several rotations

(hundreds of years [Harvey and others 1987]).

High accumulations of woody debris have been

measured in north Idaho's most productive western

white pine sites, i.e., western hemlock/clintonia habitat

type (Jurgensen and others 1997, Graham and others

1994). Although this represents one of the most

productive ecosystems in interior forests, seemingly

high accumulations may indicate blister rust-driven

mortality occurring over the last 50 years or so in these

forests (over a 90% loss of the type [Monnig and Byler

1992]) have made an unwelcome, if long-term addition

to north Idaho soils and soil-related processes (fig. 3).

Figure 3 —Ranges of coarse woody debris measured on

forested sites of the Interior West. Sites identified by state

and habitat type. Note the heavy volumes located in the

western hemlock/clintonia habitat type of northern Idaho

(Graham and others 1994). This habitat type is representative

of what were the best western white growing sites in the

region.

EFFECTS ON NUTRIENT

CONSERVATION AND CYCLING

With lowering of base height and increasing density of

canopies, plus increasing canopy-and fine root-derived

litter, there are accompanying changes in both the

location of nutrients and the microbial-mediated

processes that affect them. These changes effectively

drive nutrients and nutrient cycling process up from

deep in the soil, or down from high in the canopies, to

the soil surface (litter and humus). Both are highly

vulnerable to disturbance-related effects, especially site

preparation-driven physical removal or wildfire-driven

volatilization (Jurgensen and others 1997).

With increases in wood from climax species there is

also an increase in white rotted woody products that

also ultimately contribute primarily to surface litter and

humus horizons. With accumulations of Douglas-fir

and pine wood there is an increase in nutrients

sequestered in a recalcitrant organic horizon. Such an

effect can be good or bad, depending on the nature

140
and history of the site. Site nitrogen and potassium

stores and cycling are of particular concern (Jurgensen

and others 1997, Moore and others 1991, Mika and

Moore 1990)

Thus, the roles of microbes can be larger and different

than in historical forests. As a result, nutrients are

likely to be more concentrated near soil surfaces in

virtually all forest ecosystems in the interior west than

in the past. Where forests are changing composition

from a few large stems of ponderosa pine to many

smaller stems of Douglas-fir, this process is likely

sequestering substantial nutrient resources into a

recalcitrant organic material slow to release them and

susceptible to fire or harvesting-related loss (Jurgensen

and others 1997, Mann and others 1988). A situation

likely to cause problems in both the short- and longterm.

Where woody debris is accumulating in forests once

dominated by live western white pine, any gain in soil

organic matter is likely to be offset by the general loss

of resistance to endemic root pathogens. This is not

likely a good tradeoff with already fertile soils! Such

changes emphasize the microbe-fire codependency in

regulating nutrient cycling processes in interior forest

soils and how important it is to understand how they

interact when dealing with the productivity and

sustainability issues affecting them (Harvey and others

1994a, Olsen 1981).

OVERALL EFFECT ON THE "NATURE" OF

INTERIOR FORESTS

The strong interaction between above- and belowground processes of historical versus most current

forests highlights the importance of activities that may

not originate in but will affect the soil (fig. 4). White

pine blister rust and lack of fire serve as conspicuous

examples. However, many soil-based, microbial-driven

processes are directly impacted. Among these are both

saprophytic and internal or live stem decay processes

(and end products), root pathogen mortality,

mycorrhizal activities and non-symbiotic nitrogen

fixation.

Linked with inevitable fire related (or lack thereof)

changes, these microbial processes are quite capable of

participating significantly in creating very different

future forests, both above- and below-ground. In many

cases the new forests are not likely to have some of

the best features of those being replaced.

Historical "pine type," largely seral forests were likely

well matched to soil resources, relatively resistant to

detrimental fire effects, well adapted to wide ranges of

site and short-term climate variation, subject to modest

Figure 4 —A graphical schematic representation of resource

compression, as represented by above-ground canopy

characteristics (height, density, weight, nutrient content) and

below-ground root and soil characteristics (root depth,

density, ectomycorrhizal activities and soil nitrogen content),

resulting from the conversion of "pine type" to "fir-type"

forests currently taking place in the Inland West.

(largely beneficial) insect and pathogen mortality,

generally tolerant of long-term climate change and

could be considered long-lived and relatively stable.

In contrast, current "fir type," climax forests are

probably not well matched to soil resources. They are

also not as resistant to detrimental fire effects or well

adapted to the wide range of site and climate variation.

They are often subject to very high insect and

pathogen mortality, and they can not be considered

either long-lived or stable, especially if faced with

significant long-term climate change.

Sustainability of historic rates of carbon fixation, soil

stability and related amenities (including wildlife,

diversity and other features) in many current forests are

in doubt and, in general, experiences with such forests

are not encouraging.

OBJECTIVE-BASED "SOLUTIONS"

Means to counter some of the undesirable features of

current forests with "forest health" problems have been

widely cussed and discussed. There are differences of

opinion as to whether there are problems and what, if

anything, should be done about them. Because there

have been no generally accepted concepts of forest

health (see discussions in this symposium), there are

widely divergent opinions on what effects

management, fire or no management might create. In

addition, most of the "general" solutions thus far

proposed and implemented can either help or hurt,

depending on details of history, current condition and

future trends in affected forests. Such solutions

141
usually attempt to apply a simple fix to a variety of

sometimes complex problems.

For example, if lack of fire is the problem, introduce

fire. But current fuel accumulations and distributions

place both vegetative and soil resources at very high

risk. Then salvage logging can reduce the fuels to

levels commensurate with appropriate fire effects. But

resulting roads and soil impacts can be detrimental.

Thinning, with or perhaps without prescribed fire, must

then be the answer. But thinning root disease

susceptible species is likely to set off an epidemic of

root disease activities, upset natural selection

processes or perhaps worsen fire effects if slash is not

appropriately dealt with. How about clearcuts with

heavy fuel removal and species conversion? Excessive

fuel removal could put site nutrient stores in jeopardy.

And the list continues!

The fact is, scientifically or politically, we must address

site specific problems with customized, site specific

and objective oriented solutions. For the most part

such specific solutions need not be extremely difficult,

just well considered and carried out. Appropriate, site

specific silviculture! In some cases intervention may

seem expensive, but if long-term environmental costs

are factored in, it may well be much less costly than

doing nothing (Oliver and others 1994a, 1994b, Mutch

and others 1993) One thing seems certain, without

active intervention many, perhaps most solutions for

problems now facing interior forests will get more

difficult, more expensive and the current trends more

potentially damaging as time goes on (Sampson and

others 1994).

QUESTIONS FROM THE FLOOR

Q: You mentioned an interaction between Nitrogen &

Potassium. Is that just an issue of which is more

limiting in a system or is there a biochemical

interaction... if yes can you give a quick sketch of what

it is?

A: Nitrogen-potassium interaction. Idaho forest

nutrition coop has discovered when they add N it often

sets off an Armillaria "explosion" or increased damage

by defoliaters but, if they add K to the N, the Armillaria

or other pest effects are much reduced. In many inland

sites, the 2nd most limiting nutrient is K! Proposed

explanation...inducing growth with N under shortage of

K limits trees ability to form secondary defense

chemicals thus making tree more susceptible to pests.

Since K is intimately involved in production of phenolics

and precursers the proposed interaction is at least

interesting and quite plausible.

Q: Which has a greater adverse effect on soil health?

A high severity fire resulting from unnaturally high

dense forest stands or repeated commercial thinning

entries on an order of 1 entry per 10-1 5 years with

ground - based logging systems (e.g., mechanical

harvesters) resulting in widespread and chronic soil

disturbance (compaction and displacement)?

A: Excessively hot fires versus repeated entry

(compaction/displacement) and forest soil health.

Obviously, both have the potential to cause great

damage. As described, repeated entries that cause

"widespread" compaction and displacement could be

worse than fire because restoration will be slow at best

and impossible at worst (for example loss of ash cap)!

Neither should be acceptable.

Q: Your research on coarse woody debris (CWD) levels

being used in "Eastside" effort as targets - we (in Blue

Mts) believe these (high tonnage) levels not sustainable

or realistic if/or with more natural fire cycles -

comments/help!

A: I don't think eastside Oregon and Washington are

any more difficult to retain CWD than southwest

Montana, southern Idaho or Arizona D-fir or P-pine.

Five tons CWD/acre is almost nothing, two or even

three times that should be possible, in at least some

situations. So what is the problem? Main point to our

recommendation is be aware of it's importance and do

the best job you can to keep it in the system, without

creating a fire hazard. Simply put, do not destroy or

discourage formation of such a valuable commodity in

the ecosystem. You might even be able to get more in

the system safely (with the added productivity) than

the natural system normally supports. I think this might

be doable in some systems, certain Douglas-fir and

subalpine fir forests come to mind as interesting

possibilities.

LITERATURE CITED

Amaranthus, M.P.; Perry, D.A. 1994. The functioning

of ectomycorrhizal fungi in the field: linkages in

space and time. Plant and Soil. 159:133-140.

Baker, W.L. 1988. Effects of settlement and fire

suppression on landscape structure. Ecology.

73:1879-1887.

Brown, J.K. 1983. The "unnatural fuel buildup" issue.

In: Lotan, J.E.; Kilgore, B.M.; Fischer, W.C.; and

others (coords.). Proceedings-symposium and

workshop on wilderness fire; 1983 November 15-

18; Missoula, MT. General Technical Report INT182. U.S. Department of Agriculture, Forest

142
Service, Intermountain Forest and Range Experiment

Station: Ogden, UT: 127-128.

Brown, J.K. 1978. Weight and density of crowns of

Rocky Mountain conifers. Res. Pap. INT-197. U.S.

Department of Agriculture, Forest Service,

Intermountain Forest and Range Experiment Station:

Odgen, UT: 5p.

Byler, J.W.; Krebill, R.G.; Hagel, R.K.; and others.

1994. Health of cedar-hemlock-western white pine

forests. In: Baumgartner, D.M; Lotan, J.E.; Tonn,

J.R. (comps.). Symposium Proceedings, Interior

cedar-hemlock-white pine forests: ecology and

management. Washington State University

Cooperative Extension, Pullman, WA: 107-117.

Covington, W.W.; Everett, R.L.; Steele, R.; and others.

1994. Historical and anticipated changes in forest

ecosystems of the inland west of the United States.

Journal of Sustainable Forestry. 2:13-64.

Edmonds, R.L. 1991. Organic matter decomposition

in Western United States forests. In: Harvey, A.E.;

Neuenschwander, L.F., comps. Proceedingsmanagement and productivity of western-montane

forest soils; 1990 April 10-12; Boise, ID. General

Technical Report INT-280. U.S. Department of

Agriculture, Forest Service, Intermountain Research

Station: Ogden, UT: 118-128.

Gale, M.R.; Grigal, D.F. 1987. Vertical root

distribution of northern tree species in relation to

successional status. Canadian Journal of Forest

Research. 17:829-834.

Gast, W.R.; Scott, D.W.; Schmitt, C; and others.

1991. Blue Mountains forest health report-new

perspectives in forest health. Special Report. U.S.

Department of Agriculture, Forest Service, Pacific

Northwest Region: Portland, OR.

Graham, R.T.; Harvey, A.E.; Jurgensen, M.F.; and

others. 1994. Recommendations for managing

coarse woody debris in forests of the Rocky

Mountains, INT-RP-477, U.S. Department of

Agriculture, Forest Service, Intermountain Research

Station:Ogden, UT: 13 pp.

Haak, R.A.; Byler, J.W. 1993. Insects and pathogens,

regulators of forest ecosystems. Journal of Forestry.

91:32-37.

Harvey, A.E. 1994. Integrated roles for insects,

diseases and decomposers in fire dominated forests

of the Inland Western United States: Past present

and future forest health. Journal of Sustainable

Forestry. 2:211-220.

Harvey, A.E.; McDonald, G.I.; Jurgensen, M.F.; and

others. 1994a. Microbes: driving forces for longterm ecological processes in the inland northwest's

cedar-hemlock-white pine forests. In: Baumgartner,

D.M.; Lotan, J.E.; Tonn, J.R. (coords.). Interior

cedar-hemlock-white pine forests: ecology and

management, Washington State University,

Department of Natural Resources, Pullman, WA:

157-163.

Harvey, A.E.; Hessburg, P.F.; Byler, J.W.; and others.

1994b. Health declines in interior forests:

symptoms and solutions. In: Baumgartner, D.M;

Everett, R. (coords.). Ecosystem management in

western interior forests. Washington State

University, Cooperative Extension, Pullman, WA:

163-170.

Harvey, A.E.; McDonald, G.I.; Jurgensen, M.F. 1992.

Relationships between fire, pathogens, and longterm productivity in northwestern forests. In:

Kaufman, J.B.; and others (coords.). Fire in Pacific

Northwest ecosystems: exploring emerging issues.

1992 Jan 21-23; Portland: Oregon State University,

Corvallis, OR: 16-22.

Harvey, A.E.; Jurgensen, M.F.; Larsen, M.F.; Graham,

R.T. 1987. Decaying organic materials and soil

quality in the Inland Northwest: a management

opportunity. Gen. Tech. Rep. INT-225. U.S.

Department of Agriculture, Intermountain Research

Station: Ogden, UT: 15p.

Harvey, A.E.; Jurgensen, M.F.; Larsen, M.J.; and

others. 1986. Distribution of active

ectomycorrhizal short roots in forest soils of the

inland northwest: effects of site and disturbance.

Res. Pap. INT-374, U.S. Department of Agriculture,

Intermountain Research Station: Ogden, UT: 8p.

Hessburg, P.F.; Mitchell, R.G; Filip, G.M. 1994.

Historical and current roles of insects and pathogens

in eastern Oregon and Washington forested

landscapes. General Technical Report PNW-GTR327. U.S. Department of Agriculture, Forest

Service, Pacific Northwest Research Station:

Portland, OR: 72 p.

Hoff, R.J.; Hagle, S.K. 1990. Diseases of whitebark

pine with special emphasis on white pine blister

rust. In: Schmidt, W.C.; McDonald, K.J. (comps.).

Proceedings-Symposium on whitebark pine

ecosystems: ecology and management of a highmountain resource. U.S. Department of Agriculture,

Forest Service, Intermountain Research Station:

Ogden:UT: pp. 179-190.

Jurgensen, M.F.; Harvey, A.E.; Graham, R.T.; and

others. 1997. Impacts of timber harvesting on soil

143
organic matter, nitrogen, productivity and health of

Inland Northwest forests. Forest Science. 43:234-251.

Larsen, M.J.; Harvey, A.E; Jurgensen, M.F. 1979.

Residue decay processes and associated

environmental functions in northern Rocky Mountain

forests. In: Environmental consequences of timber

harvesting in Rocky Mountain coniferous forests,

Symposium Proceedings. Gen. Tech. Rep. INT-90.

U.S. Department of Agriculture, Forest Service,

Intermountain Forest and Range Experiment Station:

Ogden, UT: 157-174.

Lehmkuhl, J.F.; Hessburg, P.F.; Everett, R.L.; and

others. 1994. Historical and current forest

landscapes of eastern Oregon and Washington. Part

I: vegetation pattern and insect and disease hazards.

General Technical Report PNW-GTR-328. U.S.

Department of Agriculture, Forest Service, Pacific

Northwest Research Station: Portland, OR: 88 p.

Mann, L.K.; Johnson, D.W.; West, D.C.; and others.

1988. Effects of whole-tree and stem-only clearcutting on post harvest hydrologic losses, nutrient

capital, and site productivity. Forest Science.

34:412-428.

Marschner, G.E.; Marschner, H. 1996. Nutrient and

water uptake by roots of forest trees. Z.

Pflanzenernahr. Bodenk. 159:11-21.

Martin, R. 1988. Interactions among fire, arthropods,

and diseases in a healthy forest. In: Healthy forests,

healthy world. Proceedings of the 1988 Society of

American Foresters national convention; 1988

October 16-18; Rochester, NY. Society of American

Foresters: Washington, DC: 87-91.

McDonald, G.I. 1991. Connecting forest productivity

to behavior of soil-borne diseases. In: Harvey A.E.;

Neuenschwander, L.F. (comps.). Proceedings -

Management and productivity of western montane

forest soils, U.S. Department of Agriculture, Forest

Service, Intermountain Research Station: Ogden, Ut:

129-144.

Mika, P.G.; Moore. J.A. 1990. Foliar potassium status

explains Douglas-fir response to fertilization in the

Inland Northwest. Water, Air and Soil Pollution.

54:477-491.

Minore, D. 1979. Comparative autecological

characteristics of northwestern tree species—a

literature review. Gen. Tech. Rep. PNW-87, U.S.

Department of Agriculture, Forest Service, Pacific

Northwest Forest and Range Experiment Station:

Portland, OR: 72 p.

Moeur, M. 1992. Baseline demographics of late

successional western hemlock/western redcedar

stands in northern Idaho Research Natural Areas.

Research Paper INT-456. U.S. Department of

Agriculture, Forest Service, Intermountain Research

Station: Ogden, UT: 16 p.

Monnig, G.; Byler, J. 1992. Forest health and

ecological integrity in the Northern Rockies. Forest

Pest Management Report 92-7. U.S. Department of

Agriculture, Forest Service, Northern Region:

Missoula, MT: 7 p.

Moore, J.A.; Mika, P.G.; Vanderploeg, J.L. 1991.

Nitrogen fertilizer response of Rocky Mountain

Douglas-fir by geographic area across the Inland

Northwest. Western Journal of Applied Forestry.

64:94-98.

Mutch, R.W.; Arno, S.F.; Brown, J.K.; and others.

1993. Forest health in the Blue Mountains; a

management strategy for fire-adapted ecosystems.

General Technical Report PNW-310. U.S.

Department of Agriculture, Forest Service, Pacific

Northwest Research Station: Portland, OR: 14p.

O'Laughlin, J.O. 1993. Forest health conditions in

Idaho. Report 11. University of Idaho, Idaho Forest,

Wildlife and Range Experiment Station, Idaho

Forest, Wildlife and Range Policy Analysis Group:

Moscow, ID: 244 p.

Oliver, CD.; Ferguson, D.; Harvey, A.E.; and others.

1994a. Managing ecosystems for forest health: an

approach and the effects on uses and values.

Journal of Sustainable Forestry. 2:113-133.

Oliver, CD.; Irwin, I.I.; Knapp, W.H. 1994b. Eastside

forest management practices: historical overview,

extent of their applications, and their effects on

sustainability of ecosystems. General Technical

Report PNW-GTR-324. U.S. Department of

Agriculture, Forest Service, Pacific Northwest

Research Station: Portland, OR: 73 p.

Olsen, J.S. 1981. Carbon balance in relation to fire

regimes. In: Mooney, H.A.; and others (coords.).

Fire regimes and ecosystem properties. General

Technical Report WO-26. U.S. Department of

Agriculture, Forest Service: Washington, DC: 377-

378.

Quigley, T.M.; Haynes, R.W.; Graham, R.T.; and

others. 1996. Integrated scientific assessment for

ecosystem management in the interior Columbia

River Basins and portions of the Klamath and Great

Basins. U.S. Department of Agriculture, Forest

Service, Pacific Northwest Research Station and

144
Bureau of Land Management, Gen. Tech. Rep.

PNW-GTR-382: Portland, OR: 303pp.

Rehfeldt, G.E. 1994. Evolutionary genetics, the

biological species, and ecology of the interior cedarhemlock forests. In: Baumgartner, D.M.; Lotan, J.E;

Tonn, J.R. (comp.). Interior cedar-hemlock-white

pine forests: ecology and management. Washington

State University, Cooperative Extension: Pullman,

WA: 91-100.

Robinson, D. 1994. The responses of plants to

nonuniform supplies of nutrients. New Phytologist.

127:635-674.

Rose, S.L.; Perry, D.A.; Pilz, D.; Schoeneberger, M.M.

1983. Allelopathic effects of litter on the growth

and colonization of mycorrhizal fungi. Journal of

Chemical Ecology. 9:1153-1162.

Sampson, R.M.; Adams, D.L.; Hamilton, S.; and others.

1994. Assessing forest ecosystem health in the

Inland West. Journal of Sustainable Forestry. 2:3-

10.

Showalter, T.D. 1994. An ecosystem-centered view

of insect and disease effects on forest health. In:

DeBano, L.F.; Covington W.W. (coords.).

Sustainable ecological systems: implementing an

ecological approach to land management. General

Technical Report RM-247. U.S. Department of

Agriculture, Forest Service, Rocky Mountain Forest

and Range Experiment Station: Fort Collins, CO:

183-195.

Van Wagner, R.C. 1977. Conditions for the start and

spread of crown fire. Canadian Journal of Botany.

7:23-24.

Wagener, W.W.; Davidson, R.W. 1954. Heart rots in

living trees. Botanical Review. 20:61-134.

Wickman, B.E. 1992. Forest health in the Blue

Mountains: the influence of insects and disease.

General Technical Report PNW-GTR-295. U.S.

Department of Agriculture, Forest Service, Pacific

Northwest Research Station: Portland, OR: 15 p.

145
OUT OF SIGHT, BUT NOT OUT OF MIND: SOIL MICROORGANISMS IN MOISTURELIMITED FORESTS

Matt D. Busse1

ABSTRACT

Microorganisms play a critical role in the maintenance and productivity of forest ecosystems by their regulation of soil

nutrient availability and organic matter decomposition. Regardless, it is uncommon for managers to consider the

microbial community and its processes when assessing forest management options. The reasons for this incongruity are

of little surprise. Microorganisms lack the charisma attributed to key plant and wildlife species, and thus are "out of

sight, out of mind." More detrimental, however, is our incomplete understanding of how microorganisms respond to

natural and management-induced disturbances. Whether changes in microbial processes following disturbance are

sufficient to alter the sustainability of forest ecosystems is unclear, thus providing limited input for decision making by

forest managers. Information from studies of moisture-limited forests on the eastern slopes of the Cascade Mountains

indicates that (1) microbial communities are well adapted to the conditions of water stress, (2) long-term decline in

microbial biomass can result from continuous silvicultural treatment, and (3) microbial activity is limited by available soil

carbon content. Management options which maintain or enhance site organic matter are recommended as the best

practice to sustain microbial processes.

Keywords: soil processes, microbial community, ponderosa pine, moisture stress, forest management practices

INTRODUCTION

Should land managers consider the composition or

function of soil microorganisms when selecting forest

management options? Common sense might argue

yes. After all, the microbial community is responsible

for a multitude of essential functions, from organic

matter decomposition and nutrient cycling dynamics to

enhancement of soil structure and degradation of

xenobiotic compounds. Partial or complete disruption

of these processes would, without question, place

irreparable stress on most terrestrial ecosystems. Unfortunately, common sense is not always a practical

option.

Forest managers must make operational decisions that

impact ecosystem health and stability, yet rarely do

they have the time or knowledge to account for every

resource component. Further complicating their

decision-making process are the seemingly constant

demands from diverse social, political, economic, and

environmental interest groups. In considering the responsibilities facing forest managers, Kimmins (1994)

stated, "it is neither possible, nor is it necessary, to

consider all of the processes and components of a

forest ecosystem in order to make useful predictions

about the long-term consequences of forest management." Although few would argue against this viewpoint, it fails to offer us any direction for the selection

of characteristics vital to ecosystem sustainability

versus those considered superfluous. For example, are

microbial communities sufficiently resilient following

natural and human disturbances to warrant their disregard (are they superfluous from a management perspective?), or could such an indifference result in a

long-term degradation of site quality? If indeed we

accept Kimmins statement as an "operational reality",

then the challenge is to identify, without bias, which

processes and components are most important to the

sustainability of a given ecosystem.

While few would suggest that microorganisms are

superfluous to the function of forest ecosystems, some

ground work is required before endorsing their consideration in forest management decisions. Specifically,

there must be evidence that (1) management practices

or natural disturbances alter the function of microbial

communities, (2) resulting changes in microbial function following disturbance produce detrimental, longterm changes in forest ecosystems, or (3) key species

or functional groups (a set of species that have similar

effects on a specific ecosystem-level process) lack

resilience and are irrevocably lost following disturbance.

Herein lies the problem. While there are numerous sitespecific examples of the impact of disturbance on

microbial communities (for examples, see Atlas and

others 1991, Fritze and others 1993, Staddon and

others 1997, Visser and Parkinson 1989), the evi1Matt D. Busse is a Research Microbiologist with the U.S. Department of Agriculture, Forest Service, Pacific Southwest Research

Station, 2400 Washington Ave., Redding, CA 96001.

146
dence to support or refute these principles, particularly

(2) and (3), is far from complete.

Therefore, in this paper I will argue that attempts to

incorporate our knowledge of microbial processes or

community composition in management decisions,

while good intentioned, are impractical and premature.

Scientific and theoretical applications of microbial ecology have yet to catch-up with the disciplines of plant

and wildlife ecology; we still lack knowledge of the

diversity of organisms present in forest soils, how

active they are, how they interact with other organisms, and how resilient they are to management practices (Madsen 1996, Wardle and Giller 1996). I will

also argue, however, that ignorance is not bliss, and

that soil organisms should not be overlooked by forest

managers. To the contrary, sustainable forest management requires explicit understanding of both above- and

below-ground processes. Certainly soil microorganisms

are difficult to relate to and lack a certain degree of

charisma associated with more popular organisms.

Nevertheless, awareness by managers of their role in

forest ecosystems and protection of their habitat needs

(primarily site organic matter) is emphasized.

The forests on the dry, east side of the Cascade Mountains provide a unique backdrop for the discussion of

the role of microbial communities. These forests are

less biologically complex and maintain slower rates of

biogeochemical cycling in comparison with the forests

in western Washington, Oregon, and California as a

consequence of their stress and disturbance regimes. I

will focus on three topics as they pertain to the inland

forests, with emphasis given to the ponderosa pine

forests of central Oregon:

1. How well do microbial communities adapt to stress

and human-caused disturbance? Whether their

prevalence (population size and activity) fits within

the demands of the system, as suggested by Zak

and others (1994), will be questioned. In addition,

two case studies will be presented to emphasize

the differences in long-term versus short-term responses of microorganisms to common silvicultural

disturbances (thinning, harvesting method, prescribed fire, fertilization, and vegetation control).

2. Gaps in our knowledge of forest soil biology: what

are the causes and what are the potential solutions?

3. What role can forest management play to insure a

functioning microbial community.

MICROBIAL RESPONSES TO STRESS

AND DISTURBANCE GRADIENTS

External stress (moisture, temperature, or nutrient

limitations for example) and disturbance (flooding,

wildfire, or silvicultural treatment for example) are of

fundamental importance in defining ecosystem composition, structure, and processes (Grime 1979, Kimmins

1996, Wardle and Giller 1996). This is particularly

evident in forests east of the Cascade crest where

stress and disturbance gradients strongly regulate ecosystem productivity. Low annual precipitation, cool

temperatures, poor nutrient supply, and wildfire all

contribute to restrict growth conditions. A key question to ask, therefore, is whether soil processes can

function effectively in these harsh conditions, or, conversely, whether the microbial community is unable to

adapt and thus becomes a limiting factor to forest

health and productivity.

In attempting to deal with a very complex subject, I will

limit my discussion to the effects of moisture stress

and silvicultural disturbance. Further, microbial processes will be viewed from a synecological perspective

(combining bacteria, fungi, algae, and viruses as a

whole), ignoring the concern of Ohtonen and others

(1997) that stress should be discussed at the species

level to address differences in stress tolerance between

soil organisms. My aim in this section is to show that

soil microorganisms are arguably the most tolerant

group of organisms in dealing with stress and disturbance.

Adaptation to Water Stress

There is no immunity to the debilitating effects of moisture stress. Soil microorganisms respond to water

stress with a decline in total activity and biomass

(Sommers and others 1981), as do most organisms.

As an example, monthly measurements of microbial

respiration (an index of total activity) decreased linearly

with increasing water stress in soil collected from central Oregon ponderosa pine forests (fig. 1). Microbial

biomass (not presented) showed a similar, although

less dramatic, decline as the soils dried. These responses should come as little surprise. Decreasing soil

water content results in reduced motility of microorganisms (with the exception of fungi and actinomycetes

that have a hyphal network to span air spaces), as well

as reduced nutrient diffusion, and increased energy

requirements to maintain cellular integrity.

Most theoretical evidence, however, suggests that

microorganisms can adapt or become tolerant of stress

more easily than higher organisms (Atlas 1984, Wardle

and Giller 1996, Ohtonen and others 1997). This

premise is attributed to several inherent characteristics

of microorganisms. Most important is the tremendous

147
Figure 1—Relationship between microbial activity and water

stress in mineral soil from the central Oregon pumice zone.

Water potentials of -0.01 and -1.5 MPa represent field

capacity and permanent wilting point, respectively, for the

pumice zone soil. Samples were collected monthly from May

through October 1990 at eight replicate ponderosa pine

stands.

genetic diversity and malleability of soil microorganisms. Bacteria alone have as high as 4,000 different

genomes per gram of soil (Torsvik and others 1990).

Short generation times in comparison to higher organisms also provides a mechanism for rapid genetic transfer, leading to environmentally responsive mutations.

In addition, microbial communities are physiologically

versatile; no other group is capable of utilizing such a

wide range of compounds for their energy and nutrient

requirements (virtually no natural or manmade compound can escape their degradative abilities). Finally,

microorganisms are ubiquitous in natural environments,

possess a large surface area for nutrient and water

uptake, and have a high rate of metabolic activity, all

characteristics that promote a rapid recovery following

the alleviation of stress.

A good example of microbial adaptation to moisture

stress is illustrated in figure 2. The optimum soil moisture conditions for microbial biomass in this study

depended on the geographical origin of the soil. The

microbial community adapted to arid conditions was

actually stimulated by the onset of water stress, while

the community from a subalpine soil declined rapidly in

size. A related strategy was identified by Atlas and

others (1991). They found that the addition of chemical pollutants resulted in a reduction in the genetic

diversity of soil microorganisms. The remaining population, however, had greater versatility in carbon substrate utilization and tolerance to external conditions

(temperature, salt, and pH), indicating that a generalist

population had been selected for which was better

adapted to the conditions of stress.

Figure 2 —Differential adaptation of microbial communities

to water stress. The arid soil is from an Artemisia tridentata

dominated site and the subalpine soil is from an Abies

lasiocarpa site, both in northern Utah. Microbial biomass

was measured as ATP concentration. Adapted from Knight

and Skujins (1981).

The ability of microbial communities in central Oregon

soil to tolerate moisture stress can also be inferred by

comparing their population size and activity with other

terrestrial ecosystems. Zak and others (1994) provide

such an opportunity with their reported estimates of

microbial biomass for various North America ecosystems. A partial summary of their results is presented in

figure 3 (sites 1-7), along with the average microbial

biomass that I measured in central Oregon during the

1990 growing season (site 8). I would like to emphasize two points here. First, microbial biomass in the

ponderosa pine forests of central Oregon is intermediate to the poorly productive, desert grassland (site 1)

and the highly productive forests in the Pacific Northwest (site 6) and North Carolina (site 7), as expected.

Second, Zak's group found a strong relationship existed

between microbial biomass and annual net primary

productivity (ANPP). This synchronicity suggests that

microbial and plant communities respond in a similar

fashion to meet the carry capacity of a given ecosystem. Using their regression equation that predicts

microbial biomass as a function of ANPP, the predicted

microbial biomass in central Oregon is 315 kg/ha

(based on the ANPP value from Gholz (1982)), well

below my measured value of 507 kg/ha. Actual microbial biomass in central Oregon, therefore, exceeds its

predicted carrying capacity, and supplies further evidence that microbial populations are not a limiting

component in these forests.

A similar conclusion can be drawn by comparing rates

of microbial processes in central Oregon with those

from the highly productive forests on the westside of

the Cascades Mountains. For example, the litter decay

rate in central Oregon (Monleon and Cromack 1996) is

about three-fold lower than the rate in western Washington (Edmonds 1980). This also approximates the

relative differences in symbiotic nitrogen fixation (Busse

in press) and ANPP (Gholz 1982) between the two

148
Figure 3 —Comparison of microbial biomass from a central

Oregon ponderosa pine forest (site 8) with other sites across

North America. Microbial biomass values for sites 1-7 were

extrapolated from Zak and others (1994).

regions. Therefore, although water stress reduces the

absolute rates of processes, the microbial community is

able to function within the demands of the system.

Impacts of Silvicultural Treatments

Silvicultural manipulation plays an important role in the

design of many temperate and tropical forests (Attiwill

1994, Kimmins 1996). Treatments such as harvesting,

site preparation, prescribed fire, vegetation control, and

others can alter site biogeochemistry (accumulation,

loss, and availability of carbon and nutrients), biodiversity, microclimate, wildfire risk, and long-term productivity. Knowledge of how ecosystems respond to

silvicultural treatments, in turn, is required if we are to

understand the processes that determine sustainable

forests.

Several pitfalls are encountered in discussing the impact of disturbance on forest ecosystems. Topping the

list is the unpredictable nature of disturbance. For

instance, changes following silvicultural disturbance

can range from undetectable to large-scale modifications in forest composition, structure, and function

(Kimmins 1994). As a result, attempts to make generalities regarding ecosystem responses to disturbance

are often ill-advised. Prescribed fire can serve as a

prime example: increases in soil nitrogen availability

after fire can either be high (Covington and Sackett

1992), intermediate (Monleon and others 1997), or low

(Simon 1990) depending on the burn conditions.

A conceptual diagram of the impact of disturbance on

ecosystem integrity is presented in figure 4. Ecosystem integrity declines with increasing disturbance until

either mitigation is required or sufficient time is needed

to insure natural recovery. Mitigation of soil compaction on federal lands when bulk density increases more

than 20 percent above pre-disturbance values is a

classic example of this principle. The diagram also

points out a troublesome fact: there is a poorly defined

transition zone between acceptable and unacceptable

levels of disturbance that highlights the unpredictable

natural of disturbance ecology. As a result, hard targets for mitigation standards are often the result of

educated guessing. What is acceptable or unacceptable will remain a subjective issue until we increase our

understanding of the mechanisms responsible for longterm changes in ecosystem sustainability.

Figure 4—Conceptual diagram of the relationship between

disturbance and ecosystem integrity.

The information that follows is derived from two studies of microbial responses to silvicultural practices in

moisture-limited forests. Admittedly, these studies do

not represent a thorough analysis of the subject. Instead my intentions are to (1) question whether common management practices alter microbial characteristics sufficiently to require mitigation, and (2) emphasize

the importance of long-term experiments to avoid misinterpretation of initial findings. The first study, which

began in 1959 on the Pringle Falls Experimental Forest

in central Oregon, is a long-term comparison of the

effects of understory vegetation control on ponderosa

pine growth and soil quality. The second study,

shorter in duration, compares the effects of common

silvicultural practices (thinning, prescribed fire, organic

matter removal, and fertilization) on microorganisms

and vegetation in central Oregon forests.

Experiment #1 : Long-term effects of vegetation

control —The initial objective of this study was to

determine whether competition from understory vegetation reduces ponderosa pine growth (Barrett 1982).

The hypothesis was that continual removal of competing vegetation would result in greater tree growth due

to improved water and nutrient status. Two treatments (understory vegetation present, understory vegetation absent) were compared at five levels of initial

thinning of a young pine stand. Barrett's hypothesis

was confirmed after 20 years; understory vegetation

significantly reduced tree growth. Surprisingly however, continued measurements of tree growth have

found that the benefits of understory removal were

short-lived, and that from 1976 through 1994 there

149
were no differences between treatments in the rate of

tree growth (table 1; Busse and others 1996). A similar response was also found for a productive pine plantation on the Challenge Experimental Forest in California (Bill Oliver, personal communication). The takehome message is simple: use caution when interpreting

findings from short-term studies out of respect for the

complexity of forest ecosystems.

Table 1 —Long-term response of a ponderosa pine forest in central

Oregon to continuous vegetation control.

Component

Tree growth, m3/ha/yr

1959 - 1975

1976 - 1994

Mineral Soil

Organic C, g/kg

Total N, g/kg

Microbial biomass, ug/g

0 horizon

C, Mg/ha

N, kg/ha

Microbial biomass, kg/ha

Understory vegetation

Absent

2.36

4.41

28.3

1.1

778

11.5

146

406

Present

1.84

4.52

35.3

1.6

9 4 4

14.2

193

304

Percent

Change

- 20

2

17

17

21

2 3

32

34

ing methods (whole tree removal, bole only removal, no

removal, and no thinning; see Shea (1993) for treatment and site descriptions). The primary hypothesis is

that the level of organic matter retention following

harvesting or burning is an important determinant of

site quality and sustainability.

The short-term response by the microbial community to

selected treatments is presented in table 2. Unlike the

long-term experiment at Pringle Falls (which has the

same soil type and climate), there has been virtually no

measurable response from the microbial community to

silvicultural treatment in the initial six-year period.

Prescribed fire resulted in a slight decline in microbial

biomass, however this was offset by an increase in

carbon use efficiency (qCO2) by the remaining population. Differences in microbial characteristics between

the four methods of harvesting treatments were also

nominal (not shown), suggesting that organic matter

(forest floor) retention has minimal impact on shortterm microbial responses. Table 3 also points to the

relative insensitivity of microorganisms to management

practices in the Bend LTSP study. Overstory and

understory vegetation responded positively to fertilization, while microbial biomass was unaffected. This

supports my contention that microorganisms are the

least sensitive group of forest biota to stress or disturbance.

A summary of selected soil properties from the experiment is also presented in table 1. Presence of

understory vegetation produced substantial increases in

soil C, N, and microbial biomass. This was likely the

result of increased organic matter input from litterfall

and root turnover by understory shrubs. Since the total

nutrient content of the system was not altered by

either treatment (leaching losses were minimal and only

a minor nitrogen input from symbiotic nitrogen fixation

was evident), my assumption is that the presence of

understory vegetation produced a reapportionment of

nutrients into a more active biological pool, and resulted in increased microbial biomass. Whether the

improvement in soil quality with understory vegetation

present might have contributed to the long-term

overstory response is an untested, yet intriguing hypothesis.

Experiment #2: Short-term responses to

silvicultural treatments —The Deschutes National

Forest in central Oregon responded to a mountain pine

beetle outbreak in the early- to mid-1980's by targeting

about 100,000 acres of second-growth ponderosa pine

for thinning. The Bend Long-term Site Productivity

Study (LTSP) was developed to monitor the effects of

this thinning program and to evaluate the effects of

several other silvicultural practices on ecosystem productivity. The main treatments in this study include

thinning, prescribed fire, fertilization, and four harvestKNOWLEDGE GAPS

Developing practical applications for forestry has been

a challenging and often frustrating goal of soil microbiologists for most of this century. There are notable

exceptions of course, such as the study of forest pathology, mycorrhizae (for plant nutrition or edible mushroom production), and symbiotic nitrogen fixation.

Nevertheless, the struggle to provide 'hands-on' information continues. To understand this dilemma requires

an understanding of the limitations to the study of soil

microbiology.

The difficulties in studying soil organisms are unique.

Bluntly, soil is a lousy medium for conducting ecological studies. It is opaque, heterogeneous, and has a

tremendously complex physical and chemical structure.

Adding insult to injury, most microorganisms are too

damn small to be identified without the use of time

consuming methods. As a result, theories of organism

succession, population dynamics, and disturbance

ecology remain at infancy in comparison with those of

plant ecology.

Methodological limitations abound in the study of soil

microbiology. In the 1920's for example, Selman

Waksman commented that "the lack of proper methods

for a microbiological investigation of soils is widely

150
Table 2 —Short-term response of soil quality indicators to disturbance in central Oregon

ponderosa pine forests. Soil C, microbial biomass, and qCO2 were measured in the fifth growing

season after treatment; litter decomposition was measured for a 2-year period beginning the

second growing season after treatment. The metabolic quotient (qCO2) is an index of carbon use

efficiency by microorganisms.

Treatment

Control

Prescribed burn

Fertilize

Soil C

g/kg

2.3 aa

2.3 a

2.4 a

Microbial

Biomass

ug/g

372 a

316 b

389 a

qCO2

resp/biomass

048 a

0.55 b

0.44 a

Litter decay rate

yr 1

0.181 a

0.183 a

0.178 a

a

 Means within a column followed by the same letter are not significantly different at the 0.05

level.

Table 3— Growth responses to fertilization (nitrogen, phosphorus, and sulfur) in central Oregon

pine forests.

Lifeform

Ponderosa pine, m3/ha/yr

Shrub biomass, kg/ha

Herbaceous biomass, kg/ha

Microbial biomass, mg/kg

Control

3.1 ba

971 b

18 b

372 a

Fertilized

4.9 a

1420 a

324 a

389 a

Percent

Increase

61

46

1700

4

a

 Means within a row followed by the same letter are not significantly different at the 0.05

level.

recognized" (Waksman 1925). Unfortunately, a similar

quote was also made in 1997: "the practical difficulties

encountered in identifying soil organisms seem to

be insurmountable at the moment" (Ohtonen and oth

ers 1997). The difficulties in developing useful methods has been partly responsible for our inability to

answer the following fundamental questions outlined

by Madsen (1996):

1. What microorganisms are present in terrestrial

ecosystems?

2. What are the microorganisms actually doing?

3. How can we make accurate measurements of in

situ microbial processes that are integrated in time

and space?

It is no wonder that practical applications have eluded

soil microbiologists. The struggle to understand the

basic ecology of soil organisms still faces an uphill

battle.

Another limitation in the study of soil microbiology is

the relatively large number of microbial species found in

nature (table 4). According to Hawksworth (1991) the

best estimate is that about 10 percent of bacteria and

only 5 percent of fungi on earth have been identified.

This fact alone makes it practically impossible to determine whether the taxonomic diversity of soil organisms

is important to forest processes, or whether external

factors such as stress and disturbance impact soil

diversity. It is also responsible for what Kennedy and

Smith (1995) criticize as the 'black box' approach to

soil microbiology. That is an approach in which only

the inputs and outputs of processes are measured,

without any understanding of the nature of the organisms directly responsible for the processes.

Is help on the way? Perhaps. Improvements in methods of detection have been dramatic during the past

ten years. In particular, microbiologists now have an

arsenal of molecular techniques available for taxonomic

and ecological studies (Head and others 1998).

Bornemann and Triplett (1997), for example, used

rDNA sequencing to identify changes in bacterial spe151
cies following the conversion of a mature tropical forest

to pastureland. Changes in microbial functional groups

in response to forest management can also be identified now using carbon-substrate utilization patterns

(Garland and Mills 1991), fatty acid signatures (Zelles

and others 1992 ), or ribosomal RNA methods (Muyzer

and others 1993). Whether these methods will deliver

soil microbiology to a level that has direct applications

for forest managers remains to be seen.

Bacteria

Fungi

Algae

Viruses

Number of species

Described

3,000

69,000

40,000

5,000

Estimated

30,000

1,500,000

60,000

130,000

A ROLE FOR FOREST MANAGERS?

What can managers do to insure a functioning microbial

population in dry inland forests? Education is the first

step; it is important to recognize the role of soil organisms in forest ecosystems, and to recognize that moisture limitations do not override the importance of microbial processes such as nutrient cycling and decomposition. Granted, these processes do not function at

the impressive rates found in more productive forests;

they are still an inherent requirement of any ecosystem

however. Recognition of this fact can then become a

factor in the critical-thinking process that occurs during

the design of forest management options.

The next step is to acknowledge that microbiologists

will not be able to provide managers with many practical solutions to important soil biology questions. At

least not in the immediate future. For example, concepts such as standards and guidelines for microbial

processes, identification of keystone species (or functional groups) for given habitat types, and inexpensive,

user-friendly methods are still a source of frustration for

the scientific community. This frustration will remain,

unfortunately, until the knowledge gap in our understanding of population dynamics and processes are

narrowed. Until then, we will continue to use the

nefarious 'black box' approach to describe soil processes.

The best approach that managers can take to insure

the sustainability of microbial processes, therefore, is

to protect key soil resources, particularly forest floor

and mineral soil organic matter, that serve as habitat

for microorganisms (fig. 5). Following either natural or

human-caused disturbance, changes in the availability

of soil resources dictate microbial population dynamics.

Managing for microorganisms thus becomes an indirect

process of selecting options which insure the stability

of the soil resource. I have singled out organic matter

as the key soil resource because of its role in fulfilling

the habitat needs of most microorganisms for energy,

nutrients, and water.

Figure 5 —Importance of conserving soil resources. Changes

in resource availability, particularly site organic matter, have a

direct impact on microbial communities and ecosystem

sustainability.

The premise that microorganisms in moisture-limited

forest soils are limited by organic matter is verified in

figure 6. Only those treatments that received an input

of available carbon responded with a flush of microbial

activity. This observation is in agreement with the

conclusions of several authors who suggest soil organic

matter is directly linked to ecosystem sustainability

(Jurgensen and others 1997, Kimmins 1994, Morris

and others 1997). In fact, Jurgensen and others

(1997), in a discussion of dry inland forests, emphasize

that "the maintenance of adequate soil organic matter

levels is critical for sustaining forest health and productivity" based on its many biological, physical, and

chemical functions. Obviously, managers can't overlook the added function of organic matter (forest floor)

as potential fuel for wildfires in dry forests. There is no

easy answer to the question of how much decaying

material (wood, branches, litter) should remain on a

given site. Instead, it requires an assessment at a local

scale, with input from soil, fire, and silviculture specialists.

152
Figure 6 —Carbon limitation in central Oregon soil. Microbial

respiration was measured after the addition of available

carbon (glucose) and/or nitrogen (ammonium nitrate) at

optimum concentrations.

CONCLUSION

Picture a forest without a soil microbial community.

With a build-up of organic material incapable of decomposing and no continuous supply of plant available

nutrients, the forest would be unsustainable to say the

least. Yes, this is a ridiculous concept, but it serves as

a method to visualize the importance of microorganisms. They are out of sight for sure, but should not be

out of mind or taken for granted.

Several concepts were introduced with regard to the

function of microorganisms in moisture-limited forests.

Foremost, the microbial community appears more tolerant of stress and disturbance in comparison with other

forest biota. This does not mean microorganisms are

unresponsive to silvicultural manipulation however.

Results from studies of ponderosa pine forests in central Oregon show a decline in microbial biomass following continuous vegetation control, and underscore the

importance of long-term experiments to avoid overinterpretation of short-term findings. Currently there is

limited "hands-on" information of microbial processes

to assist forest managers as a consequence of our

inadequate understanding of microbial ecology. An

indirect approach is recommended, therefore, to insure

the sustainability of microbial processes: selection of

forest management options that protect their habitat

needs for site organic matter.

LITERATURE CITED

Atlas, Ronald M. 1984. Use of microbial diversity

measurements to assess environmental stress. In:

Klug, M.J.; Reddy, C.A. (eds.). Current perspectives in microbial ecology. Washington, D.C.:

American Society for Microbiology 540-545.

Atlas, Ronald M.; Horowitz, Ami; Krichevsky, Micah;

Bej, Asim K. 1991. Responseof microbial populations to environmental disturbance. Microbial Ecology 22:249-256.

Attiwill, Peter M. 1994. The disturbance of forest

ecosystems: the ecological basis for conservative

management. Forest Ecology and Management

63:247-300.

Barrett, James W. 1982. Twenty-year growth of

ponderosa pine saplings thinned to five spacings in

central Oregon. Res. Note PNW-301. Portland, OR:

Pacific Northwest Experiment Station, Forest Service, U.S. Department of Agriculture; 18 p.

Borneman, James; Triplett, Eric W. 1997. Molecular

microbial diversity in soils from eastern Amazonia:

evidence for unusual microorganisms and microbial

population shifts associated with deforestation.

Applied and Environmental Microbiology 63:647-

653.

Busse, M.D. [in press]. Ecological significance of nitrogen fixation by actinorhizal shrubs in interior

forests of California and Oregon. In: Powers, R.F.;

Hauxwell, D.; Nakamura, G. (eds.). Forest soil

biology and forest management. Gen. Tech. Rep.

Albany, CA: Pacific Southwest Research Station,

Forest Service, U.S. Department of Agriculture.

Busse, M.D.; Cochran, P.H.; Barrett, J.W. 1996.

Changes in ponderosa pine site productivity following removal of understory vegetation. Soil Science

Society of America Journal 60:1614-1621.

Covington, W.W.; Sackett, S.S. 1992. Soil mineral

nitrogen changes following prescribed burning in

ponderosa pine. Forest Ecology and Management

54:175-191.

Edmonds, Robert L. 1980. Litter decomposition and

nutrient release in Douglas-fir, red alder, western

hemlock, and Pacific silver fir ecosystems in western Washington. Canadian Journal of Forest Research 10:327-337.

Fritze, Hannu; Pennanen, Taina; Pietikainen. 1993.

Recovery of soil microbial biomass and activity

from prescribed burning. Canadian Journal of Forest Research 23:1286-1290.

Garland, J.L.; Mills, A.L. 1991. Classification and

characterization of heterotrophic microbial communities on the basis of community-level sole-carbonsource utilization. Applied and Environmental Microbiology 57:2351-2359.

153
Gholz, Henry L. 1982. Environmental limits on aboveground net primary production, leaf area, and biomass in vegetation zones of the Pacific Northwest.

Ecology 63:469-481.

Grime, J.P. 1979. Plant strategies and vegetation

processes. New York: John Wiley and Sons.

Hawksworth, D.L. 1991. The fungal dimension of

biodiversity: magnitude, significance, and conservation. Mycological Research 95:641-655.

Head, I.M; Saunders, J.R.; Pickup, R.W. 1998. Microbial evolution, diversity, and ecology: a decade of

ribosomal RNA analysis of uncultivated microorganisms. Microbial Ecology. 35:1-21.

Jurgensen, M.F.; Harvey, A.E.; Graham, R.T.; PageDumroese, D.S.; Tonn, J.R.; Larsen, M.J.; Jain,

T.B. 1997. Impacts of timber harvesting on soil

organic matter, nitrogen, productivity, and health

of Inland Northwest forests. Forest Science

43:34-251.

Kennedy, A.C.; Smith, K.L. 1995. Soil microbial diversity and the sustainability of agricultural soils.

Plant and Soil 170:75-86.

Kimmins, J.P. 1994. Identifying key processes affecting long-term site productivity. In: Dyck, W.J.;

Cole, D.W.; Comerford, N.B. (eds.). Impacts of

forest harvesting on long-term site productivity.

New York: Chapman and Hall.

Kimmins, J.P. 1996. Importance of soil and role of

ecosystem disturbance for sustained productivity

of cool temperate and boreal forests. Soil Science

Society of America Journal 60:1643-1654.

Knight, W.G.; Skujins, J. 1981. ATP concentration

and soil respiration at reduced water potentials in

arid soils. Soil Science Society of America Journal

45:657-660.

Madsen, Eugene L. 1996. A critical analysis of methods for determining the composition and

biogeochemical activities of soil microbial communities in situ. In: Stotzky, G; Bollag, Jean-Marc

(eds.). Soil biochemistry, Volume 9. New York:

arcel Dekker, Inc.

Monleon, Vincente J.; Cromack, Kermit Jr. 1996.

Long-term effects of prescribed underburning on

litter decomposition and nutrient release in ponderosa pine stands in central Oregon. Forest Ecology

and Management 81:143-152.

Monleon, Vincente J.; Cromack, Kermit Jr.; Landsberg,

Johanna D. 1997. Short- and long-term effects of

prescribed underburning on nitrogen availability in

ponderosa pine stands in central Oregon. Canadian

Journal of Forest Research 7: 369-378.

Morris, Dave M.; Kimmins, J.P. (Hamish); Duckert, Dan

R. 1997. The use of soil organic matter as a criterion of the relative sustainability of forest management alternatives: a modelling approach using

FORECAST. Forest Ecology and Management

94:61-78.

Muyzer, G.; de Waal, E.C.; Uitterrlinden, A.G. 1993.

Profiling of complex microbial populations by denaturing gel electrophoresis analysis of polymerase

chain reaction-amplified genes coding for 1 6S

rRNA. Applied and Environmental Microbiology

59:695-700.

Ohtonen, R.; Aikio, S.; Vare, H. 1997. Ecological

theories in soil biology. Soil Biology and Biochemistry 29:1613-1619.

Shea, Ronald W. 1993. Effects of prescribed fire and

silvicultural activities on fuel mass and nitrogen

redistribution in Pinus ponderosa ecosystems of

central Oregon. M.S. Thesis, Oregon State Univ.,

Corvallis, OR.

Simon, Steven A. 1990. Fire effects from prescribed

underburning in central Oregon ponderosa pine

plant communities: first and second growing season after burning. Unpublished draft on file,

Fremont National Forest, Lakeview, OR.

Sommers, L.E.; Gilmour, CM. ; Wildung, R.E.; Beck,

S.M. 1981. The effect of water potential on decomposition processes in soils. In: Parr, J.F.; Gardner, W.R.; Elliot, L.F. (eds.). Water potential relations in soil microbiology. Madison, Wl: Soil Science Society of America.

Straddon, W.J.; Duchesne, L.C.; Trevors, J.T. 1997.

Microbial diversity and community structure of

postdisturbance forest soils as determined by solecarbon-source utilization patterns. Microbial Ecology 34:125-130.

Torsvik, V.; Goksoyr, J.; Daae, F.L. 1990. High diversity of DNA of soil bacteria. Applied and Environmental Microbiology 56:782-787.

Visser, S.; Parkinson, D. 1989. Microbial respiration

and biomass in soil of a lodgepole pine stand acidified with elemental sulphur. Canadian Journal of

Forest Research 19:955-961.

Waksman, S.A. 1925. Soil microbiology in 1924: an

attempt at an analysis and a synthesis. Soil Science 19:201-249.

154
Wardle, D.A.; Giller, K.E. 1996. The quest for a contemporary ecological dimension to soil biology.

Soil Biology and Biochemistry 28:1549-1554.

Zak, Donald R.; Tilman, David; Parmenter, Robert R.;

Rice, Charles W.; Fisher, Frederick M.; Vose,

James; Milchunas, Daniel; Martin, C. Wayne.

1994. Plant production and soil microorganisms in

155

late-successional ecosystems: a continental-scale

study. Ecology 75:2333-2347.

Zelles. L; Bai, Q.Y.; Beck, T.; Beese, F. 1992. Signature fatty acids in phospholipids and lipopolysaccharides as indicators of microbial biomass and

community structure in agricultural soils. Soil Biology and Biochemistry. 24:317-323.
MANAGING ROOT DISEASES BY MANAGING STANDS

By Walter G. Thies1

ABSTRACT

Root diseases are important natural disturbance agents in forest ecosystems in the Pacific Northwest. The common

root diseases are caused by fungi. Root disease fungi inhabit roots, rather than soil. This helps them to avoid

competition for food with saprophytic soil organisms. At landscape scales and over long time frames, the impacts of

root diseases are considerable. Understanding the ecological roles of diseases, their impacts on resources, management

influences on them, and their effect on management is crucial to meeting ecosystem management objectives.

Keywords: root diseases, forest pathology, fungi, white rot, brown rot

INTRODUCTION

Root diseases are important natural disturbance agents

in forest ecosystems of the Pacific Northwest. Root

disease biology, incidence, severity, and impact must

be understood and accounted for if ecosystem

management objectives are to be met. Root diseases

exert a profound influence on forest structure,

composition, function, and yield. Root diseases are

important gap formers; they create openings of varied

sizes in the forest, depending upon the pathogen(s) and

hosts present. These gaps are then colonized by a

variety of forb, shrub, and tree species depending on

site characteristics and seed sources. Whether root

disease is perpetuated on the site is a function of the

relative susceptibilities of the ingrowth species. If

highly susceptible species continue to colonize root

disease pockets, their likelihood of reaching large sizes

is low. If immune species occupy root disease pockets,

inoculum dies out and eventually susceptible species

may grow on the site again without being influenced by

root disease. If intermediately susceptible species

colonize pockets, root disease inoculum will be

perpetuated on the site and influence vegetation

development according to relative species

susceptibilities. This may mean reduced growth,

increased levels of butt decay, and increased chance of

windthrow and mortality. Basically, root diseases

influence species composition by selectively killing

some species and allowing others to survive. Species

diversity may increase or decrease depending upon

location. Production of special forest products such as

berries, boughs, and mushrooms may be increased.

Root diseases influence structure by reducing the

likelihood that some trees will achieve large sizes, or at

the very least, slowing the process. Stocking levels

may be reduced in discrete areas or across the stand

depending on the distribution of inoculum and tree

species present. Root diseases create standing dead

trees that may be important for wildlife habitat. The

very nature of the decay associated with some root

diseases would suggest that many snags will be shortlived, that is, they may remain standing for only a short

time compared to trees killed by other agents. Root

diseases also create down woody material that is

important for habitat, water holding capacity, and

nutrient cycling.

Accepting the effects of root disease may be

consistent with meeting resource objectives in the

immediate future; mortality and spread rates, in

general, are low and impacts are not usually dramatic in

the short-term. However, root disease impacts over

time and space must be considered in planning.

Defining the pathogen(s), inoculum concentration and

distribution, and population of host species of an area

could be critical. There is a legacy of large acreages

where susceptible species were planted on sites with

high inoculum levels. Whether or not these stands are

entered for thinning treatments, are interplanted with

resistant species, or are left alone, will depend on the

desired future condition of the sites. For example, if

the objective is to grow large trees as quickly as

possible, either to achieve late successional stand

structural characteristics or to satisfy harvest volumes,

plantations with high levels of laminated root rot that

are predominantly Douglas-fir may not be capable of

meeting goals.

Knowledge of the occurrence and level of root disease

in an area is required to make reasonable predictions of

root disease impacts. Surveys, stand examinations,

and resource inventories must include collecting a

minimal level of information on root disease. This

includes the identity of the root diseases present, tree

species composition, size, and stocking, the proportion

1Walt Thies is a forest pathologist, USDA Forest Service, Pacific Northwest Research Station, Forestry Sciences Laboratory, 3200

Jefferson Way, Corvallis, OR 97331.

156
and species of trees affected, how the root disease is

distributed in the area, and an estimate of area

affected. With this information, a clear picture can be

drawn of current root disease conditions, enabling

prioritization of treatments. Future impacts can also be

predicted with these data using tools such as the

Western Root Disease Model Extension to the Forest

Vegetation Simulator (Frankel 1998).

COMMON TRAITS OF ROOT DISEASES

It is not within the scope of this paper to provide

details about the many root diseases in Pacific

Northwest forests. However, there are some

generalizations that apply to most pathogens that will

help managers better understand diseases:

1. The common root diseases are caused by fungi.

2. Root disease fungi live belowground most of their

lives but they are root, not soil, inhabitors. The

special ability of these organisms to penetrate and

colonize healthy root systems enables them to avoid

competition for food with saprophytic soil

organisms.

3. Root disease fungi cause dysfunction in root

systems by killing cambial tissue, disrupting the

transport system, or consuming cellulose, thus

causing loss of structural root integrity.

4. Most cause white rots although a few may cause

brown rots. White rot fungi can breakdown both

cellulose and lignin, thus causing pits. Brown rot

fungi breakdown the cellulose leaving the brown

lignin untouched.

5. Most attack and invade only living tissue but once

the tree is dead they become saprophytic and can

remain within the dead root system for a long time,

in many cases decades.

6. Because they occupy the root system, which can be

fairly massive, these fungi have a good food base

and a high inoculum potential with which to attack

other hosts.

FACTORS IMPORTANT TO ROOT

DISEASE MANAGEMENT

To start planning, a manager needs to know three

general factors about a root disease found on a

particular site: host species, fungal survival or inoculum

longevity, and means of spread. I won't spend more

time on host species except to say that most diseases

are fairly host specific or have a well-defined host

range, avoid particular hosts and the disease will not

develop.

Fungal Survival

Many root disease fungi can survive for decades after

the host tree has died. Perhaps the longest survivor is

the pathogen causing laminated root rot.

In very large stumps this fungus can survive for 50 to

100 years. Longevity of the pathogen inside roots may

be attributable in large part to zone lines commonly

formed by this and many other fungi in colonized

wood. Zone lines are formed as protective barriers to

unfavorable conditions and other decay fungi or

antagonistic organisms; they are visible as black lines

when decayed wood is cut. The process by which a

pathogenic fungus converts from a long-term

saprophytic existence inside an old root, to a

pathogenic form on the surface of a root able to initiate

an infection on a new host, is not well understood.

Thus, any stump or root with an internal viable

pathogen should be viewed as a potential source of

inoculum and a threat to adjacent host trees. The

regional exception to long-term survival of root

pathogens is black stain root disease; the causal fungus

dies within 1 to 2 years of tree death. However, this

disease is spread by insects and management

strategies aim to minimize spread by timing stand entry

to life cycle of the insects.

Methods of Spread

Selection of an appropriate management strategy for

root diseases depends in large part on the method of

spread of the fungus involved. The following are the

four common methods of spread, and an example

disease.

1. Root-to-root: Laminated root rot —Virtually all spread

of Phellinus weirii is by mycelia on or within roots

with little if any by windblown spores. The fungus

does not grow through the soil. The fungus persists

on the site in the large roots and stumps of dead or

cut trees and appears to effectively wall itself off in

the infested wood by a protective hyphal sheath,

seen as dark zone lines in cut wood. The pathogen

continues to live saprophytically in dead roots for 50

years or more. Infection in a young stand begins

when roots of young trees contact residual infested

stumps and roots from the preceding stand. The

infection spreads between living trees through root

contact. As roots decay, a tree dies standing or is

robbed of structural support and is windthrown.

Nearly all root diseases can spread from an infected

tree to a healthy tree when their roots are in

contact.

2. Insect: Black stain root disease —Black stain root

disease is caused by the fungus Leptographium

157
wageneri (perfect stage Ophiostoma wagenen,

rarely found). This disease is unique among the root

diseases affecting forests of the Pacific Northwest

because it is spread by root feeding insects, and it is

a vascular wilt disease that does not actually cause

a decay of root or stem tissue. Although long

distance spread involves insect vectors, once

introduced into a tree the fungus can move to new

trees across root contacts or by growing up to 15

cm through the soil. Tree to tree spread is not

dependent on hosts of low vigor, and centers may

expand radially at a rate of nearly 5 feet per year.

Insect spread involves root-feeding bark beetles and

weevils. Fruiting bodies of the fungus form in the

egg-laying galleries of the insects, and the sticky

spores that are produced adhere to the insects and

are dispersed by them. The spore-carrying bark

beetles and weevils feed and breed in roots of lowvigor trees so occurrence of black stain root disease

centers in disturbed areas probably reflects vector

preference for stressed or injured trees.

3. Water: Port Orford cedar root disease —Port-Orfordcedar root disease, caused by Phytophthora

lateralis, an introduced fungus, is a severe threat to

Port-Orford-cedar on sites favorable for the

pathogen. The pathogen is an Oomycete, or "water

mold" and is spread by water-borne spores

(zoospores) in either water or contaminated soil.

Port-Orford-cedar grows in a narrow coastal strip in

southern Oregon and northern California. It grows

on a wide variety of sites including streambanks,

bogs, coastal sand dunes, deep productive soils,

and dry sites. Port-Orford-cedar needs a consistent

supply of water and is an important species in

riparian ecosystems. The pathogen is spread mainly

by people moving infected seedlings and infested

soil to disease-free sites. Livestock and wildlife may

move the pathogen in soil on their feet but that is

believed to be a minor factor in most situations. A

much greater threat is the movement of soil on tires

and machinery from an infested area to a

noninfested area during road building or logging

activities. Zoospores of the fungus, which can

swim a short distance, move in water films in the

soil, and are attracted to the uninjured root tips of

cedar. They are totally dependent on free water for

infection and spread and cannot survive in dry soil.

Natural spread is virtually all downslope or

downstream in water moved by gravity.

4. Air currents: Annosus root disease —Annosus root

disease is caused by the fungus Heterobasidion

annosum (previously Fomes annosus) which can

infect its hosts in two ways: by mycelial growth

across root contacts and by windblown spores.

Spore infection of freshly cut stumps is the primary

way that new infection centers start. In coastal

stands, stumps can become infected any time of the

year. After growing vegetatively down through the

stump and root system, the fungus can move to

adjacent trees across root contacts, but does not

grow through the soil. Infection centers can expand

at one foot per year. The fungus can remain viable

in large stumps for at least 50 to 60 years but in

small (less than 6 inch dib) stumps the fungus dies

out within a few years. The fungus can also infect

wood exposed when living trees are wounded.

When this fungus colonizes and injury the rate of

spread is two to three times greater upward than

downward.

DISEASE MANAGEMENT PLANNING

Although no universal system exists for integrating root

disease management into land management decisions,

some variation of the following decision ladder may be

useful for including root disease management in

selecting appropriate management scenarios for Pacific

Northwest forests.

1. ESTABLISH MANAGEMENT OBJECTIVES: The

choice of options will likely vary depending on the

objectives. Examples include variations on fiber

production, wildlife habitat enhancement, alternative

forest products, or some combination of these.

2. GET A HISTORICAL PERSPECTIVE: Determine the

historical appearance of the stand: tree species (not

what was planted), relative stand structure,

composition, and age. Determine the diseases that

were and are present, and their distribution. Equally

important for some diseases is the availability or

probable future availability of inoculum from nearby

stands.

3. PROPOSE MANAGEMENT STRATEGIES: Use

available models to evaluate the effects of the

proposed strategies over time. An example is the

western root disease model which can currently be

used to address laminated root rot, annosus root

disease, and armillaria root disease and their

interactions with bark beetles. Simulations designed

around one of these three root diseases can be used

to approximate others. Stands can be grown to

evaluate various cultural treatments and "play" with

options. In the absence of models, use local data to

project the impacts of various scenarios. Growing a

forest takes a long time, and it is important to

consider the cumulative effects of various actions

and inactions.

4. INTEGRATE: Integrate the decision for managing

disease with the needs for managing other

resources on the site. Try to integrate the activities

on the site with those planned for adjacent sites.

158
Through this integration, attempt to project impacts to

the next larger scale. For example, try to project

how the actions planned for this stand will, through

time, impact other activities in the watershed.

Managers should expect that there will be some failures

to meet desired future conditions. The best laid plans

have some built-in uncertainty. In general, we can't

control diseases. This is especially true of root

diseases given their hidden inoculum. And, where

controls are available, they are generally too expensive

to use on large areas. When we try to manage

diseases we are really engaging in risk reduction.

Managing forests is a very long-term enterprise, and

changes may occur that are well beyond our reasonable

planning horizon. Examples include changes in local

weather conditions or the unexpected appearance of

other disturbance agents which alter the outcome of

our applied management strategy.

TAKE HOME MESSAGES

1. Pathogens are a natural and perhaps essential part

of the forest ecosystem.

2. The presence of root disease does not necessarily

mean that the soil or forest is unhealthy.

3. Root diseases contribute to plant and animal

diversity.

4. Losses are a social not a biological term.

5. Take measures against introducing exotic diseases.

6. Include root disease management considerations

early in the planning process.

SUMMARY AND CONCLUSIONS

Forest tree diseases have profound influence on stand

structure, composition, and function in Pacific

Northwest forests. They are among the most

important disturbance agents operating in the forest

ecosystems of the Pacific Northwest, even though their

effects are often not spectacular at any given point in

time. At landscape scales and over long time frames,

the impacts of root diseases are considerable.

Understanding the ecological roles of diseases, their

historical and current distributions, their impacts on

resources, how management influences them, and how

they affect management, is crucial to meeting

ecosystem management objectives. Unless diseases

are included as a planning consideration, it is unlikely

that ecosystem management strategies will be

successful in achieving desired future conditions.

QUESTIONS FROM THE FLOOR

Q: "What insights do you have on effects of subsoiling

and disease spread and/or effects of compaction and root

damage."

A: Considering the amount of subsoiling that is

apparently being done, we know precious little about the

effects. I can suggest a few general observations or

comments:

1. It is unlikely that there will be much movement of

inoculum as a result of subsoiling.

2. Subsoiling equipment should stay as far from living

trees as practical.

3. Any subsoiling is likely to injure roots. In some cases

armillaria root disease has been associated with

subsoiling activities. The injuries may present

opportunities for annosus root disease to get started,

especially in true firs. But, in general, it is unlikely

that subsoiling is going to cause any catastrophic

increase in root disease.

4. Hard data is not available but we would expect that

subsoiling in the vicinity (within 200 meters) of black

stain root disease centers would attract the beetle

vectors of the fungus and initiate new centers.

5. My recommendation is that every plan for subsoiling

should carry with it a plan and a commitment of funds

for a monitoring effort to determine which if any

diseases increase. In general that has not been done

and is an essential part of evaluating the efficacy of

subsoiling. There are many pathologists available who

could assist in developing monitoring plans.

6. In 1993 a study was initiated in Region Five to

examine the effects of subsoiling on Jeffrey pine on

the Milford Ranger District, Plumas national Forest, in

northern California. The study and some results are

reported in two papers: (a) Otrosina, W. J.; Sung, S.;

White, L. M. 1996. Effects of subsoiling on lateral

roots, sucrose metabolizing enzymes, and soil

ergosterol in two Jeffrey pine stands. Tree Physiology

16:1009-1013. (b) Kliejunas, J. T.; Otrosina, W. J.

1997. Progress report: Effects of subsoiling study,

Milford Ranger District, Plumas National Forest.

USDA Forest Service, Pacific Southwest Region,

Forest Pest Management Report Number R97-01

10 pp.

159
SELECTED REFERENCES

Boyce J.S. 1961. Forest Pathology. McGraw Hill Book

Company, New York, Toronto, London. 572 p.

Filip, G.M.; Kanaskie, A.; Campbell, A., III. 1995.

Forest disease ecology and management in Oregon.

Manual 9. Corvallis, OR: Oregon State University

Extension Service. 60 p.

Frankel, S.J. (tech. coordinator). 1998. User's guide

to the Western Root Disease Model, Version 3.0.

Gen. Tech. Rep. PSW-GTR-1 65. Albany, CA: U.S.

Department of Agriculture, Forest Service, Pacific

Southwest Research Station. 120 p.

Hadfield, J.S.; Goheen, D.J.; Filip, G.M.; and others.

1986. Root diseases in Oregon and Washington

conifers. Portland, OR: U.S. Department of

Agriculture, Forest Service, Pacific Northwest

Region. 27p.

Hansen, E.M.; Goheen, D.J.; Hessburg, P.F.; Witcosky,

J.J.; SchowalterJ.D. 1988. Biology and

management of black stain root disease in

Douglas-fir. In: Leptographium Root Diseases of

Conifers. Harrington, T.C.; Cobb, F.W. Jr. (eds.).

APS Press, The American Phytopathological Society,

St. Paul, Minnesota, p. 63-80.

Hansen, E.M.; Lewis, K.J. (eds.). 1997. Compendium

of conifer diseases. St. Paul, MN: The American

Phytopathological Society. 101 p.

Hawksworth, F.G.; Wiens, D. 1996. Dwarf

mistletoes: biology, pathology, and systematics.

Agricultural Handbook 709. Washington, D.C.: U.S.

Department of Agriculture, Forest Service. 410p.

Hessburg, P.F.; Goheen, D.J.; Bega, R.V. 1995. Black

stain root disease of conifers. USDA Forest

Service. Forest Insect and Disease Leaflet 145. 9 p.

Otrosina, W.J.; Scharpf, R.F. (technical coordinators).

1989. Research and management of annosus root

disease (Heterobasidion annosum) in western North

America. USDA Forest Service, Pacific Southwest

Forest and Range Experiment Station. General

Technical Report PSW-116. 177 p.

Scharpf, R.F. 1993. Diseases of Pacific coast

conifers. Agricultural Handbook 521. Albany, CA:

U.S. Department of Agriculture, Forest Service,

Pacific Southwest Research Station. 199 p.

Thies, W.G.; Sturrock, R.N. 1995. Laminated root rot

in western North America. Gen. Tech. Rept. PNWGTR-349. Portland, OR: U.S. Department of

Agriculture, Forest Service, Pacific Northwest

Research Station. 32 p. In cooperation with: Natural

Resources Canada, Canadian Forest Service, Pacific

Forestry Center.

Zobel, D.B.; Roth, L.F.; Hawk, G.M. 1985. Ecology,

pathology, and management of Port-Orford-cedar

(Chamaecyparus lawsoniana). USDA Forest Service,

Pacific Northwest Forest and Range Experiment

Station, Gen. Tech. Rep. PNW-184. Portland, OR.

161 p.

160
ROLE OF SOIL ORGANISMS IN RESTORATION

AND APPLICATIONS TO MANAGEMENT
STRUCTURE AND FUNCTION OF BIOLOGICAL SOIL CRUSTS

Jayne Belnap1, John Williams2 and Julie Kaltenecker3

ABSTRACT

Biological soil crusts are found throughout the world. They are generally dominated by cyanobacteria, lichens, and

mosses. When moistened, cyanobacteria move through soils, leaving sticky sheaths behind. Sheaths connect loose soil

particles, contributing soil resistance to wind and water erosion. Abandoned sheaths are found up to 10 cm down in

soils, providing stability at depth. Moistened sheaths absorb ten times their volume of water. Crusted surfaces

roughened by frostheaving slow rainfall runoff, thus increasing infiltration.

Crusts can reduce germination of exotic annual grasses. They contribute nitrogen to desert ecosystems, where

nitrogen often limits productivity. Plants growing in crusted areas often have higher levels of many nutrients than

plants in non-crusted areas.

Many land uses are incompatible with crustal integrity. Cyanobacterial fibers have little resistance to compressional

stress. Crushed crusts contribute less nitrogen to soils. Impacted soils are susceptible to both wind and water erosion,

leading to soil loss and burial of nearby crusts.

Recovery is generally very slow. Rates depend on climate; size, timing, and intensity of disturbance; and soil type. In

areas with less stable soils and less rainfall, full recovery may take over 250 years, even with small disturbances. Soils

with more inherent stability or areas with greater rainfall will recover more rapidly.

Keywords: cryptobiotic soil crusts, microphytic, microbiotic, soil erosion, germination, nutrient cycles, deserts

INTRODUCTION

In arid and semi-arid lands throughout the world, the

cover of vegetation is generally sparse or absent. Open

spaces between the higher plants are not bare of

autotrophic life, but usually covered by a community of

highly specialized organisms. This soil surface floral

community consists of cyanobacteria, green algae,

lichens, mosses, microfungi, and other bacteria.

Cyanobacterial and microfungal filaments weave

throughout the top few millimeters of soil, gluing loose

soil particles together, forming a biological crust which

stabilizes and protects soil surfaces from erosive forces

(Belnap and Gardner 1993). These crusts occur in all

hot, cool, and cold arid and semi-arid regions. They

may constitute up to 70% of the living cover (Belnap

1994) and have only recently been recognized as

having a major influence on terrestrial ecosystems.

These communities are also referred to as cryptogamic,

cryptobiotic, microbiotic or microphytic soil crusts

(Harper and Marble 1988).

Globally, this consortium of soil biota has many

similarities in species composition, structure and

function, in spite of occurring in unconnected and

seemingly dissimilar environments. Crusts are found in

an astonishing variety of higher plant communities

throughout the world, ranging from deserts to tundra.

WHAT BIOLOGICAL CRUSTS ARE NOT:

PHYSICAL SOIL CRUSTS

Non-biotic soil surface crusts are a major structural

feature in many arid regions. Their properties and

manner of formation have been studied for many years,

primarily because of their detrimental effects to

agriculture crops. These crusts are transient soilsurface layers that are structurally different than

material immediately beneath. Rainfall onto

unprotected soil surfaces is what forms most physical

crusts. Surface aggregates are broken apart by

raindrops. Smaller particles wash into spaces between

larger particles, thus clogging soil pores and reducing

infiltration rate by as much as 90%. In general, rain-

'Jayne Belnap is a Research Ecologist for the US Geological Survey, Biological Resources Division, Forest and Rangeland Ecosystem

Science Center, Canyonlands Field Station, 2282 S. West Resource Blvd., Moab, UT 84532.

2

John Williams is a hydrologist for the USDA-ARS in Pendleton, OR 97801-0370.

3

Julie Kaltenecker is a biologist for the Idaho State Office, Bureau of Land Management, 1387 S. Vinnell Way, Boise ID 83709.

161
formed crusts are less than 5 mm thick. This layer is

often harder than the rest of the soil because it is drier

and compounds such as salts, lime and silica have been

deposited at the surface as water evaporates. With

large pores absent, these crusts increase water runoff

and soil erosion, and prevent the emergence of

seedlings.

Crusts may form on soil of almost any texture except

coarse sandy soils containing very low silt and clay

(Lemos and Lutz 1957). Conditions that make surface

soil especially susceptible to crusting are low organic

matter, high silt and high sodium. These characteristics

are all related to soils with low structural stability and

soil aggregates easily broken down under the impact of

water. Soil aggregate structure is also reduced by the

hooves of grazing animals or machinery. Such

compacted surfaces have reduced infiltration rates and

increased surface runoff and function hydrologically in

a manner similar to raindrop-induced crusts.

Using intensive grazing to break up these crusts has

short-lived results, as the soil surface is resealed after

the first minutes of an intense rainstorm. Management

systems that promote soil surface protection (through

plant and biological soil crust cover) and increases in

soil organic matter produce the only lasting solution to

physical soil crust reduction (Blackburn 1983).

BIOLOGICAL SOIL CRUST DISTRIBUTION

Many of the dominant cyanobacteria and soil lichens

have a cosmopolitan distribution, and crustal

communities can have similar growth forms. For

instance, the same soil lichens dominate soils crusts of

both the Sonoran and Great Basin deserts of the

western United States and areas of South Africa and

Australia, while vascular plant species and climate

varies greatly between these regions. Crusts are

generally flat, and sometimes covered with a layer of

pebbles in hot deserts of the world, such as the

Atacama, Sonoran, Chihuahuan and Australian deserts.

Where freezing temperatures are common, frost

heaving often results in crusts with pedicels up to 10

cm high, such as seen on the Colorado Plateau, Great

Basin and the Arctic.

In the western US, overall biological crust cover is

greatest in low elevation inland sites (<1000 m)

compared to mid-elevation sites (1000-2500 m). This

results from reduced vascular plant cover, leaving more

surface available for colonization and growth of crustal

organisms. Crustal organisms have reduced water and

nutrient needs relative to vascular plants, and can

withstand the harsh growing conditions found in plant

interspaces. The positive relationship between

biological crust cover and available soil surfaces has

been demonstrated in other studies (Anderson and

others 1982a; Eldridge 1993a; Harper and Marble

1988; Johansen 1993; Rogers 1972).

Crusts in cool deserts are dominated by the nonheterocystic cyanobacterium Microcoleus vaginatus

(Belnap and Gardner 1993). M. vaginatus is a large,

highly mobile filamentous species that lacks

heterocysts, the structurally differentiated, oxygenexcluding cells where cyanobacterial nitrogen fixation

generally takes place. In contrast, most crusts in warm

deserts are dominated by heterocystic cyanobacteria

such as Lyngbya and Nostoc, with no single species

dominating the microflora. These species are much

smaller and less mobile than M. vaginatus. Common

co-dominant lichens in both warm and cold deserts are

Collema sp. and Placodium sp.

Likelihood of well-developed lichen crusts is increased

with increasing moisture by factors that increase soil

surface stability. Factors that increase soil surface

stability are 1) more closely-spaced plants; 2) flatter

areas (depositional rather than erosional surfaces); 3)

limited surface rocks, roots and/or light plant litter to

slow water and or wind and to keep hooves off

surface; 4) soils with more inherent stability (silt/clay

>sandy>shrink-swell clay); and 5) more stable

microhabitats (under shrubs, away from small washes).

More stable soils support greater cover and more varied

populations of cyanobacteria, lichens and mosses.

Very sandy soils and high shrink-swell clays (such as

Mancos shale) have low inherent stability, and as a

result, generally only large filamentous cyanobacteria

that are highly mobile can survive in these soils (such

as M. vaginatus). However, if the very sandy soils are

sufficiently stabilized by these cyanobacteria, other

smaller species can then colonize them (such as

Scytonema and Nostoc) and lichen and mosses may be

found under vascular plants, where some protection

from sediment burial is provided. In soils such as

Mancos shales, some mosses and lichens can be found

occasionally on northern slopes. Extremely shallow

soils colonized by M. vaginatus are also often stable

enough to support most soil species of lichens and

cyanobacteria found in a given area. As soil stability

increases, these rich communities of cyanobacteria,

mosses and lichens become more widespread, covering

all surfaces not occupied by vascular plants or rocks.

Time since disturbance also influences the composition

of crusts. The composition of biological soil crusts

changes through time in a manner analogous to

transitions occurring generally during terrestrial plant

succession (Johansen and others 1984). As is typical

of primary succession of vascular plants, nitrogen-fixing

organisms appear first. The species appearing earliest

in the crustal succession are cyanobacteria and algae

(Anderson and Rushforth 1976; Johansen and others

1981, 1984; Harper and Marble 1988). The ubiquitous

162
cyanobacterium, M. vaginatus, can colonize within just

days of soil disturbance if soils are wet. Mosses and

lichens colonize later, with the lichens Collema and

Catepyrenium generally the first to appear.

MICROSTRUCTURE

Lichens and mosses are easily seen without aid of

magnification. However, much of the structure and

function of crusts is dependent on cyanobacteria, green

algae, and microfungi. All these are often too small to

be seen without a microscope. In most desert soils,

cyanobacteria contribute the most to crust

microstructure.

Cyanobacterial filaments confer structural integrity to

the soils in which they occur. When wetted, the

sheath of filamentous cyanobacteria swell, expelling

the living filaments and leaving behind empty sheath

material. These filaments often string sand and clay

particles together, much like fibers in fiberglass.

Depending on environmental conditions and soil

textures, cyanobacterial sheaths may be found at

depths of 10 cm below the soil surface (Belnap and

Gardner 1993). As aeolian and water-borne materials

are trapped in the polysaccharide sheaths of

cyanobacteria on the soil surface, old sheaths are

gradually buried. Thus, influence on water-holding

capacity and soil stability may extend far below the

depth to which light can penetrate, unless sheaths are

crushed. If sheath-soil connections are broken by

trampling or vehicles, these sheaths are no longer living

and therefore cannot be repaired.

Measurements of the bumpy surface topography of

undisturbed crusts show that few bumps are greater

than 8 mm across, with cyanobacterial filaments

ramifying throughout the micropedicels. The volume of

cyanobacterial filaments per unit soil surface coverage

is thus increased greatly in areas where

microtopography of the soil surface is very complex,

while still preserving photosynthetic capabilities. The

roughened soil surface also provides habitat for

microfauna, which are responsible for humus

production (Gayel and Shtina 1974; Danin 1978).

ECOLOGICAL ROLES

Carbon Fixation

Biological soil crusts are an important source of fixed

carbon on sparsely vegetated areas common

throughout the West (Beymer and Klopatek 1991).

While vascular plants provide organic matter to soils

directly underneath them, large interspaces between

plants have little opportunity to receive such input.

Carbon contributed by soil crusts help keep plant

interspaces fertile, and thus provide energy sources for

other microbial populations.

The dominant components of biological soil crusts are

photosynthetic organisms, and require sunlight. When

soils are dry, the bulk of the cyanobacterial biomass is

at 0.2-0.5 mm, with bundles found down to 4 mm

where sufficient light for net carbon gain is available

but UV exposure is reduced. When soils are

moistened, M. vaginatus, which lacks UV-protective

pigments, can be seen migrating to the surface for

short time periods. Species with UV-screening

pigments, are commonly found on the soil surface

(Garcia-Pichel and Belnap 1996).

Recently, ecophysiologists have studied photosynthetic

patterns in the lichen species present in biological soil

crusts (Lange and others 1997, 1998). They have

found that soil lichens have several different strategies

to gain carbon. While photosynthesis is activated at

low water content for all species, species varied in their

response to higher water content. Some species, such

as Collema, showed dramatic drops in carbon fixation

at water contents over 60-80%. Other species

showed little, if any, depression at high water

contents. Water-holding capacity varies between

lichens, which will also influence time of carbon

fixation for a given species. Photosynthetic rates are

also dependent on temperature (Lange 1980; Lange

and others 1997; Nash and others 1982a, b; Rychert

and others 1978). Most crustal species increase

photosynthetic rates with increasing temperatures up

to about 26-28 C, after which rates decline.

Differential responses to timing and amount of

precipitation and temperature may mean species

composition will shift in response to predicted climate

changes. Many areas that are currently winter-rain

dominated are predicted to receive increases in summer

rain. Effects on species composition may be currently

reflected in the different composition of biological soil

crusts in the Sonoran (summer rain-dominated) vs the

Mojave, Colorado, Colorado Plateau, and Great Basin

deserts (winter-rain dominated). Winter-rain deserts

have a rich flora of soil lichens and the cyanobacterial

flora is heavily dominated by Microcoleus. In contrast,

the summer rain-dominated Sonoran desert has only a

small subset of the lichens found in winter-rain deserts,

while the cyanobacterial flora is a complex community

co-dominated by up to 10 species.

Nitrogen Fixation

Nitrogen concentrations are known to be low in desert

ecosystems relative to other ecosystems. Total

atmospheric input over the past 10,000 years has been

conservatively estimated about 3 kg N m 2 (ignoring

cyanobacteria inputs), with 77% lost through wind

163
erosion, ammonia volatilization, nitrification, and

denitrification (Peterjohn and Schlesinger 1990).

Extensive surveys in cold deserts have revealed only a

few nitrogen-fixing plants (Farnsworth and others

1976; Wullstein 1989). Since nitrogen can limit net

primary productivity in many desert ecosystems,

normal nitrogen cycles are critical to the fertility of

semi-arid soils and in preventing desertification (Dregne

1983).

Cyanobacteria and cyanobacterial-containing soil

lichens can be an important source of both fixed

nitrogen for plants and soils in desert ecosystems

(Evans and Ehleringer 1993; Belnap 1995). Most

biological soil crusts in the western US are dominated

by nitrogen-fixing cyanobacteria and lichens (M.

vaginatus, Scytonema myochrous, Nostoc sp.) and the

soil lichen Collema. All these species are capable of

both light and dark nitrogen fixation (Paerl 1990).

Nitrogen inputs from biological soil crusts have been

estimated from 1 to 100 kg ha-1 annually (Harper and

Marble 1988). The lowest estimates are almost ten

times atmospheric input estimates by Peterjohn and

Schlesinger (1990). Nitrogen fixation is highly

dependent on past and present water and light regimes,

as well as species composition (Rychert and others

1978; Belnap 1996), with maximum fixation at

approximately 26°C and above 20% moisture. Past

disturbance history is also a critical factor in

determining fixation rates (Belnap 1995, 1996).

Five to 88% of N fixed by Nostoc has been shown to

leak into the surrounding substrate (Belnap and others

1997; Magee and Burris 1954). Nitrogen leaked from

these organisms is available to nearby vascular plants

(Mayland and Macintosh 1966). Vascular plants

growing in biologically crusted areas show higher tissue

concentrations of nitrogen when compared to plants in

uncrusted soils (Harper and Pendleton 1993; Belnap

1994, 1995; Belnap and Harper 1995). Since rainfall

and dewfall events in desert areas are often too small

to promote plant growth, but do stimulate microbial

community activity, biological soil crusts can be the

dominant nitrogen source for the ecosystem. As with

carbon, crusts contribute nitrogen to soils both under

plants and in plant interspaces, thereby counteracting

the tendency of nutrients to concentrate around

perennial plants.

Effects on Vascular Plants

Germination and establishment—The presence of

crusts can affect vascular plant germination and

establishment. While small cracks and crannies on the

soil surface are sufficient for small-seeded plants to

lodge and germinate, most large-seeded plants need

some cover by soil or vascular plant litter. In deserts

where plant litter is often limiting in interspaces, large

native seeds generally have self-burial mechanisms

(such as hygroscopic awns) or through rodent caching.

Plants adapted to loose moving soils (such as sand

dunes) or deep litter (forests) accomplish this passively.

However, exotic species may lack such adaptations.

As crusts stabilize soils, germination can be inhibited in

sites with well-developed crusts and low plant litter, as

was recently demonstrated for the annual exotic grass

Bromus tectorum in both the field and laboratory

(Belnap, Kaltenecker unpublished data). Once the

seeds germinate, biological soil crusts show no barrier

to seedling root penetration (Belnap, Pendleton and

Meyer unpublished data).

Seedling germination per se has not been shown to

limit species density or presence in desert plant

communities. Harper and Marble (1988) and Eckert and

others (1986) showed seedling establishment and

survival greater on crust-dominated soils than on soils

lacking crusts. Experiments done with both fine and

coarse-textured soils demonstrate that seedling

establishment and survival for both forbs and grasses

can be much higher in crusted areas when compared to

areas where the crust had been removed (Belnap 1993;

Harper and St. Clair 1985; St. Clair and others 1984;

Lesica and Shelley 1992). No study has shown a

negative relationship between crusts and vascular plant

growth. Rather, many studies worldwide suggest that

vascular plant cover is most often controlled by water

and/or nutrient availability rather than other site factors

(Dunkerley and Brown 1995; Mabbutt and Fanning

1987; Tongway and Ludwig 1990).

Some authors have suggested that there is a negative

relationship between cover of biological crusts and

vascular plants. However, in a recent study (Belnap

and others, in review), little correlation was seen

between vascular plants and crustal cover at coastal

and mid-elevation sites. At more arid low-elevation

sites, correlations between vascular plant cover and

cover of crustal components was generally positive.

This relationship suggests that the presence of plants

aids the survival of crustal components, perhaps due to

microclimate conditions associated with perennial

vegetation (such as decreased soil surface

temperatures and increased surface moisture due to

shading). Rather than crusts inhibiting plants, it

appears more likely that crust occur in spaces where

vascular plants cannot grow and keep them out.

Numerous other authors have reported crusts either do

not compete with vascular plants (Anderson and others

1982a; Beymer and Klopatek 1992; Jeffries and

Klopatek 1987; Kleiner and Harper 1972; 1977a, b) or

actually enhance vascular plant cover (Carleton 1990;

Dadlich and others 1969; Graetz and Tongway 1986;

Ladyman and Muldavin 1994; Ladyman and others

1994; Lesica and Shelley 1992; Muncher and others

1988; Rosentreter 1986).

164
Nutrient levels in vascular plants —Plants growing

on crusted soil often show higher concentrations and/or

greater total accumulation of various essential nutrients

when compared to plants growing in adjacent,

uncrusted soils. In southeast Utah, leaf tissue nitrogen

in annual, biennial and perennial species was 9-31%

higher in crusted areas. Dry weight were greater as

well (Belnap 1995; Belnap and Harper 1995). This was

verified with greenhouse experiments (Harper and

Pendleton 1993). Other authors have obtained similar

results with other species (Brotherson and Rushforth

1983; Shields and Durrell 1964).

Several mechanisms have been postulated to explain

this effect. Crusts accumulate nutrient rich fine soil

and organic matter (Fryberger and others 1988;

Campbell and others 1989; and Gillette and

Dobrowolski 1993). Material that blows across

biological soil crusts can become trapped, either

accumulating within low pockets in the

microtopography or stuck by exudate on algal sheaths.

Algal sheaths are phototropic, and if not buried too

deeply, will push through loose soil and organic matter,

trapping or entangling the soil in the process (Campbell

and others 1989). Danin and others (1989) proposed a

positive feedback loop based on this process in which

nutrient-rich microsites develop, leading to germination

and growth of vascular plants, which in turn contribute

to a reduction in erodibility and an increase in the

accumulation of washed or wind blown soil.

Cyanobacterial sheath material is often coated with

negatively-charged clay particles. Positively charged

macro-nutrients bind to these particles, and are thus

prevented from leaching from the soil profile (Belnap

and Gardner 1993). These clay particles are more

nutrient-rich than sand (Black 1968). Lange (1974)

demonstrated that compounds in the gelatinous sheath

material of half the cyanobacteria species studied were

able to chelate elements essential for their growth

(e.g., iron, copper, molybdenum, zinc, cobalt, and

manganese). Four of the five genera shown to possess

this ability (Anabaena, Anacystis, Lyngbya, and

Nostoc) are represented by common species in the

biological crusts of western North American deserts

(Shields and Durrell 1964). It is also possible that

nutrient differences are a result of a thermal effect, as

dark crusts would be warmer that lighter uncrusted

soils, and uptake of nutrients would occur at a higher

rate. Herbivores and other consumers benefit directly

from the enhanced nutrient status of these ecosystems

(Belnap and Harper 1995). Indirect effects include

positive correlations between soil mycorrhizae and

microarthopod populations with the presence of welldeveloped biological soil crusts (Harper and Pendleton

1993; Belnap unpublished data).

Water Relations

The effect of biological soil crusts on soil water

relations is highly variable between different regions,

soils and climatic regimes. Crustal development (e.g.,

cyanobacterial, lichen, moss), climatic regimes, the

amount of surface roughness, time since destructive

disturbance, soil texture, and soil structure can all

heavily influence hydrologic cycles at a given site. Soil

texture is especially important and can override any

effect of biological soil crusts. For instance, soils with

high shrink-swell clays have low infiltration rates, and

sandy soils have high infiltration rates, regardless of the

biological soil crusts present.

Research results conducted under a variety of soil and

climate conditions around the world show the variable,

and interactive, effects of biological soil crusts and soil

properties. While the presence of the mucilaginous

cyanobacteria can decrease soil permeability, increased

surface roughness can increase water pooling and

residence time. Consequently, in cool and cold deserts

where frost-heaving is common and biological soil

crusts greatly increase soil surface roughness, the

presence of biological soil crusts may increases the

amount and depth of rainfall infiltration. Under

sagebrush mounds in a clayey-skeletal, montmorillonitic

soil in southern Idaho, sites with a high cover of

biological soil crusts in coppice and moss-grass

microsites had significantly higher infiltration rates than

adjacent bare or vesicular dominated soils (Blackburn

and others1990; Johnson and Gordon 1986; Pierson

and others 1 994a and b; Johnson and Rosentreter

1988). (However, this result was complicated by the

presence of more plant litter under the shrubs than in

plant interspaces, which may have been responsible for

this increase). In a Utah sandy loam soil, infiltration

rates increased with increased lichen-dominated crusts

(Loope and Gifford 1972), and 3-year-old algal crusts

decreased time to ponding, although infiltration rates,

once runoff began, were the same with or without the

crust (Williams and others 1995b).

In warm deserts, where frost-heaving is not present

and biological soil crusts are flat, infiltration rates

appear to depend more on soil type and climatic

regimes, and so are site-specific. Reduced infiltration

rates as a result of biological soil crust cover have been

reported for sandy soils in Arizona (Brotherson and

Rushforth 1983), Australian sandy soils (Bond 1964;

Eldridge 1993a,b; Rogers 1977; Stanley 1983; and

Graetz and Tongway 1986), and Israeli sandy soil (Yair

1990). In New South Wales, Australia, infiltration rate

increased on a transect through a sandy-loam soil as

grazing decreased and biological soil crust cover

increased. However, comparison with adjacent

ungrazed soils with a continuous cover of lichens and

mosses showed reduced infiltration (Eldridge 1993b).

Infiltration studies conducted on sandy soils of

165
Oklahoma and Kansas (Booth 1941) and silt-loams and

clayey silt loams in Arizona (Faust 1970, 1971) failed

to find any influence of biological soil crusts on

infiltration rates.

Conversely, components of biological soil crusts rapidly

swell (Shields and Durrell 1964) up to 13 times the dry

volume (Campbell 1977), potentially closing flow

pathways through soils. Scanning electron microscopy

shows sufficient openings in sandy loam soils for water

flow (Belnap and Gardner 1993) but concentration of

silt and clay particles, combined with swelling will

restrict infiltration even in sandy soils when wet

(Verrecchia and others 1995). The hydrophobic nature

of some microphytes have been speculated to

contribute to soil surface sealing (Bond 1964);

however, other authors have experimental evidence

that shows these organisms to be highly hydrophilic

(Kidron personal communication).

The influence of biological soil crust on effective

hydraulic conductivity (Ke: the rate at which water

moves through the soil profile under field conditions) is

also variable. Lichen-dominated crusts in Utah

decreased Ke (Loope and Gifford 1972). Lichen and

algal crusts in Australia were associated with high Ke

(Greene and others 1990). Algal crusts in Utah and

non-disturbed Nevada crusts had no discernible

influence on Ke (Williams 1993; Dobrowolski 1994).

Though overall infiltration of precipitation is critical for

plant growth, where water infiltrates can also be

critical in maintaining plant community structure.

Recent work done on banded vegetation has shown

that water infiltration and runoff patterns can be

important in maintaining vegetative community

structure in hyper-arid zones. Banded vegetation

occurs in very flat areas where rainfall is too low to

support homogenous distribution of vegetation;

instead, plants are concentrated in bands that occur

perpendicular to the flow of water. These bands

capture water, nutrients and fine soil particles. The

bigger the bands, the more effective they are in

resource capture; however, their overall size is limited

by rainfall and nutrients. Biological soil crusts cover

inter-band soils. When these inter-band biological soil

crusts are disrupted, water infiltration increases

between vegetated areas. This results in less water

reaching the vegetated bands, causing large die-offs.

This was also seen in Israel, where vegetation died

when water infiltration was increased in plant

interspaces (Zaady, personal communication).

The effect of biological soil crusts on soil moisture is

also variable. Soils under biological soil crusts showed

deeper water penetration into the profile and greater

availability during drought (Brotherson and Rushforth

1983; Abrahams and others 1988). The ability of the

crust to seal the soil surface and reduce evaporation

due to high clay and silt concentrations in the crusts

has been repeatedly proposed (Danin 1978; Brotherson

and Rushforth 1983; Williams and others 1995b) and

recently supported by research specifically designed to

address the issue (Verrecchia and others 1995).

However, this can vary. In Utah and Mexico, soil

moisture was less under disturbed crusts than intact

crusts (Harper and Marble 1988; Meyer and GarciaMoya 1989). Increased soil temperature, through the

absorption of solar energy by black crusts, may

increase soil moisture evaporation rates (Harper and

Marble 1988).

As can be seen above, the impact of crusts on

hydrologic cycles is highly site-specific. While much of

this variation is probably due to soil and crust

differences, this is difficult to assess due to lack of

standardized data collection methods and better

descriptions of soil, crust and climate characteristics at

each study site.

Soil Stabilization

Wind and water can be major erosive forces in deserts,

as sparse vegetation leaves large soil spaces

unprotected by plant litter or vegetative cover (Goudie

1978). These interspace soils in deserts is most often

stabilized by rocks or biological soil crusts. Biological

soil crusts play an important role in both preventing soil

loss and facilitating soil accretion. They contribute to

aggregate stability, surface and subsurface organic

matter, soil structure and soil surface microtopography

as a result of physical and chemical attributes (Booth

1941; Fletcher and Martin 1948; and Shields and

Durrell 1964). Contribution to aggregate stability can

be visually assessed on both a micro- and macro-scale

(Belnap 1994; Belnap and Gardner 1993).

As seen in the microstructure chapter, polysaccharides

extruded by the cyanobacteria and green algae, in

combination with lichen and moss "rootlets", entrap

and bind soil particles together, increasing the size of

soil aggregates. As soil aggregates get larger, they are

heavier, have a greater surface area, and are therefore

more difficult for wind or water to move. In both wet

and dry soils, these structures can be seen firmly gluing

soil particles together at up to 7 cm below the soil

surface. The presence of biological soil crusts enable

otherwise loose sandy soils to stay in place on steep

slopes, and stabilize pockets of very shallow soil (Bond

and Harris 1964; Marathe 1972; Gayel and Shtina

1974; Danin and Yaalon 1980; Schulten 1985; Graetz

and Tongway 1986; Campbell and others 1989; Danin

and others 1989; Belnap and Gardner 1993). Rough

microtopography also slows both wind and water

erosion by creating a still air boundary layer to protect

against wind erosion. Providing detention dams creates

pooling and slowed flows, thus giving sediment time to

settle out of the water and reducing sediment loss

166
(Brotherson and Rushforth 1983; Alexander and Calvo

1990). The degree to which different types of crusts

protect the soil surface from both wind and water

erosion is: bare soil < algal crust < lichen/moss crust

(Belnap and Gillette 1997, 1998; Kinnell and others

1990; Eldridge and Greene 1994; Tchoupopnou 1989).

Biological soil crusts are unambiguously effective in

reducing wind erosion of soil (Belnap and Gillette 1997,

1998).

EFFECTS OF DISTURBANCE

Many uses of deserts result in impacts to the biological

soil crusts. The largest impacts come from driving offroad with vehicles, including military and civilian

vehicles; trampling by livestock and people; and various

mining activities. Effects of these activities are

especially noticeable at sites with highly erodible soils

with high topographic relief.

Species Composition

Surface disturbance also results in changes in species

composition of soil crusts. While multiple species of

soil lichens and mosses, as well as 4 or more species

of cyanobacteria can be found in untrampled areas on

most soil types, no lichens and only one species of

cyanobacteria are generally found in directly-adjacent

trampled areas (Belnap 1995). Changes in species

composition has implications for both soil stability and

nutrient inputs.

Microtopography

Trampled surfaces are generally flat, whereas

untrampled interspace surfaces can support

cyanobacterial pedicels from 1-7 cm high (depending

on climate of the site). Flattened surfaces do not slow

water or wind velocity, nor does sediment have an

opportunity to settle out; thus, more sediment is lost

from trampled sites than untrampled sites. Water

residence time on smooth surfaces is shorter, and

water infiltration reduced (Harper and Marble 1988).

Soil Stability

Different soils have different intrinsic soil strengths that

vary with moisture content. Arid soils with little

tendency to form inorganic aggregates (e.g., sandy

soils) are more susceptible to stresses when dry, while

soils with inorganic crusting are more susceptible to

impacts when soils are wet. Trampling greatly

accelerates desertification processes through increased

soil loss and water runoff. Soils in arid regions are

often highly erodible, and soil formation is extremely

slow, taking 5,000 to 10,000 years (Dregne 1983).

Crustal components are brittle when dry and the

connections they make between sand grains easily

crushed. Therefore, compressional disturbances, such

as those generated by animal, people and vehicular

traffic, can severely affect the ability of crusts to

stabilize soils (Belnap and Gillette 1997, 1998; Williams

and others 1995a). Damage to the underlying

cyanobacterial sheath material is non-repairable, since

living cyanobacteria are no longer present at these

depths to regenerate filament and sheath materials.

Water erosion from plots with biological soil crusts,

undisturbed for two to three years, was significantly

less than from plots without crusts in Utah and Idaho

(Williams and others 1995b; Loope and Gifford 1972;

Blackburn and others1990; Pierson and others1994a,

b). Lower erosion rates were not simply due to

decreased runoff, strongly suggesting that the reduced

erodibility was attributable to conditions unique to the

coppice and moss-grass microsites. Globally, many

authors have reported that the presence of biological

soil crusts reduces soil susceptibility to water erosion

through reduced raindrop erosion and sediment loss

from sites (Alexander and Calvo 1990; Beymer and

Klopatek 1992; Eldridge 1993; Eldridge and Greene

1994; Foth 1978; Harper and Marble 1988; Ladyman

and Muldavin 1996).

All studies have shown that crust cover reduces wind

erosion on all soil types by requiring much higher wind

speeds to initiate soil particle movement (Belnap and

Gillette 1997, 1998; McKenna-Neuman and others

1996; Williams and others1995a). Resistance to wind

erosion parallels biological crust development. Welldeveloped crusts (with lichens and mosses) on all soils

tested conferred the greatest erosion resistance, when

compared with less-developed crust. Compressional

disturbances to the crusts greatly decrease resistance

to wind erosion for all soil types, regardless of the

disturbance regime or soil type, as cyanobacteria and

lichens are brittle when dry and crush easily. Vehicle

tracks result in greater damage than hoof prints on a

given soil type. After 10 years of recovery, sandy soils

tested in southeast Utah were still susceptible to wind

erosion at commonly-occurring wind speeds (Belnap

and Gillette 1997, 1998).

Decreases in the wind resistance of soils is directly

associated with increased sediment movement (Leys

1990; Williams and others 1995a). Nearby biological

soil crusts can be buried by blowing sediment, resulting

in the death of the photosynthetic organisms (Belnap

1995, 1996). Because over 75% of the

photosynthetic biomass, and almost all photosynthetic

productivity, is from organisms in the top 3 mm of

these soils, very small soil losses can dramatically

reduce site fertility and further reduce soil surface

stability. In addition, many plants have relatively

167
inflexible rooting depths, and often cannot adapt to

rapidly changing soil depths.

Nutrient Cycles

Nitrogenase activity can be dramatically reduced in soil

crusts after experimentally-applied disturbance (Belnap

and others 1994; Belnap 1995, 1996). Disturbance by

human feet, mountain bikes, four-wheel drive trucks,

tracked vehicles (tanks), and raking all resulted in an

immediate 40-80% reduction in nitrogenase activity.

After 6-9 months, nitrogenase activity had dropped still

lower in the disturbed areas than the initial

measurements, with treatments showing an 80-100%

reduction in nitrogenase activity.

Crust disturbance can result in large decreases in soil

nitrogen through a combination of reduced input and

elevated losses (Peterjohn and Schlesinger 1990).

Current long-term studies demonstrate a 42% decrease

in soil nitrogen and 34% decrease in plant tissue

nitrogen 25 years following release from grazing. The

greatest long-term impact of disturbance may be on the

soil microbial pool: plant-available nitrogen can

decrease almost 80% following disturbance. This has

large implications for ecosystems that are dependent

on biological crusts for nitrogen, such as the Colorado

Plateau (Evans and Ehleringer 1993; Evans and Belnap

1 999). Reduced fertility of systems is one of the most

problematic aspects of desertification (Dregne 1983).

Albedo

Albedo is also of concern in semi-arid and arid systems.

When trampled surfaces were compared to untrampled

surfaces, there was up to a 50% increase in

reflectance from 0.25 to 2.5 μm (Belnap 1995). This

represents a change in the surface energy flux of

approximately 40 watts/m2. Large acreages of

trampled areas can lead to changes in regional climate

patterns in many regions (Sagan and others 1979).

Changes in surface albedo will result in changes in soil

temperatures. Trampled surfaces have significantly

different surface temperatures than untrampled

surfaces. For example, trampled surfaces in southeast

Utah were found to be 23°C cooler than adjacent

crusted surfaces. In the winter, surface temperatures of

well-developed crusts were up to 14°C higher than

ambient air temperature (Belnap 1995). Surface

temperatures can be very important in desert systems.

Nitrogenase activity is heavily temperature dependent,

with lower temperatures resulting in lowered activity

levels (Rychert and others 1978). Altered soil

temperatures can be expected to affect microbial

activity, plant nutrient uptake rates and soil water

evaporation rates. Soil temperatures have been shown

to affect seed germination time and seedling growth

rates for vascular plants. Timing of these events is

often critical in deserts, and relatively small delays can

reduce species fitness and seedling establishment

which may eventually affect community structure

(Bush and Van Auken 1991). Food and other resources

are often partitioned among ants, arthropods and small

mammals on the basis of surface temperaturecontrolled foraging times (Doyen and Tschinkel 1974;

Crawford 1991; Wallwork 1982). Many small desert

animals are weak burrowers, and soil surface

microclimates are of great importance to their survival

(Larmuth 1978). Consequently, altering surface

temperatures can affect nutrient availability and

community structure for many desert organisms, thus

increasing susceptibility to desertification.

Fire

High intensity fire will burn biological crusts, resulting

in reduction of visible cover, biomass, and species

diversity (Callison and others 1985; Greene and others

1990; Johansen and others 1993). The extent of

damage depends on the type of plant community in

which the crust occurs, the distribution of fuel and thus

fire intensities (Johansen and others 1993). Arid and

semiarid communities historically had patchy plant

distributions, resulting in discontinuous fuel with

biological crusts between (Whisenant 1990). Biological

crusts provide little fuel to carry a fire through

interspaces, thereby acting as "refugia" to slow the

spread of fire and decrease its intensity (Rosentreter

1986). Unburned islands of vascular vegetation and

biological crust provide propagules for reestablishment

of the community in burned areas. Johansen and

others (1993) observed that the structural matrix of the

crust was left intact following fire, maintaining stability

against erosive forces during the recovery period for

vascular plants and the biological crust.

Exotic annual grasses, primarily Bromus sp., have

invaded semiarid and arid landscapes throughout

western North America, homogenizing fuel distribution

and drastically altering fire regimes (Whisenant 1990;

Peters and Bunting 1994). Increases in both fuel

amount and continuity have resulted in large,

continuous fires. For example, natural fire return

intervals for Artemisia tridentata ssp. wyomingensis

communities on the western Snake River Plain are

greater than 60 years (Whisenant 1990). These have

been modified to intervals of five years or less, with

some areas burning annually. Salt-desert shrub

communities rarely burned under pristine conditions. In

addition, fire seasons have become longer than the

historical average due to the presence of dry fuel from

late spring until the onset of cool weather and fall

rains. Biological crusts are lost from the community if

fire return intervals are shorter that the period required

for the crusts to recover (Greene and others 1990;

Whisenant 1990). Lichens and mosses are lost if

annual grasslands are allowed to dominate. Increases in

168
both live plant density and litter amount result in little

open space, repressing the recovery of a diverse

biological crust (Kaltenecker 1997).

Timing and Intensity of Disturbance

Only a few studies discuss season of use and attendant

impacts on biological soil crusts. In addition, past

disturbance regimes are generally not reported. Marble

and Harper (1989) found heavy grazing treatments (17

sheep days/acre, applied over 1-2 days) in early winter

did not significantly affect biological soil cover and

composition or vascular plant cover and composition.

When the grazing season was extended to late winter,

both biological soil crust cover and species richness

declined significantly. Differences are attributed to

insufficient soil moisture in late winter/early summer to

permit regrowth of biological soil species. This

reduction in biological soil crust cover coincides with

Lusby's (1979) findings that extending the grazing

season from February 1 5 to May 1 5 resulted in

increased runoff/erosion from depletion of biological

soil crusts. Continuous grazing has been shown to be

deleterious to biological soil crusts. Jefferies and

Klopatek (1987) showed a near complete destruction

of the crust on a site heavily grazed year-long,

compared to similar pristine and light-moderate winter

gazing on sandy blackbrush sites in southern Utah and

northern Arizona. Brotherson and Rushforth (1983)

showed large reductions of both biological soil crust

and vascular vegetation on continuous moderate to

heavily grazed sites in northern Arizona.

Likewise, intensifying physical impacts through short

duration grazing strategies is also deleterious to

biological soil crusts. While this is reported to be

beneficial to vascular plant communities in some

ecosystems dominated by summer moisture such as in

Zimbabwe (Savory and Parsons 1 980) and the Great

Plains region of the U.S., it is not beneficial to either

native vascular plants or biological soil crusts in areas

where precipitation is predominantly received in the

winter, such as the Mojave, Great Basin, Colorado

Plateau or Columbia River Basin (Johansen 1986).

Platou and Tueller (1985) suggest that the natural

grazing system that evolved in the Great Plains closely

resembles high-intensity short duration grazing, while

the natural grazing system in the shrub-steppe is more

similar to rest-rotation grazing in the winter only (Mack

and Thompson 1982; Parmeter and Van Devender

1995; Hormay 1970).

crust communities. Surveys in these plant

communities show that the rich perennial moss/lichen

community has generally been replaced with annual

mosses and cyanobacteria. The mechanisms by which

the presence of annual grasses negatively affects the

biological soil crusts is not clear, but could include a

decrease in available soil surfaces (via increased cover

of vascular plant and plant litter); increased soil

disturbance by small rodents responding to an increase

in seed availability; increased fire frequency; increased

soil turnover by increased populations of soil fauna;

and/or increased soil disturbance by plant surface

roots.

REGIONAL THREATS

Many regional factors influence ecological processes in

deserts. Atmospheric nitrogen deposition from power

plants and/or car exhausts, increased UV-B radiation

from ozone reductions, and changes in land-use

patterns are likely to cause large changes in the

integrity and sustainability of ecosystems in the

southwestern US. Inherent low resource availability

and species diversity of these ecosystems (Verstraete

and Schwartz 1991) make them especially susceptible

to even small changes in critical ecosystem processes

that determine species composition and primary

productivity. However, large changes in nitrogen and

carbon dynamics may occur because components of

biological crusts may be very sensitive to these global

changes (Garcia-Pichel and Bebout 1996; Beymer and

Klopatek 1991; Evans and Belnap 1999).

Arid ecosystems may be more severely impacted by

increased atmospheric nitrogen deposition than more

mesic ecosystems. Historic rates of nitrogen input and

overall nitrogen availability are low so even small

changes in rates of deposition represent a significant

change in the nitrogen cycle. Nitrogen derived via

long-distance pollution transport from major urban

regions and from power plants may contribute

significant amounts of N to ecosystems still recovering

from more than a century of grazing impacts. Nitrogen

deposition could in part offset the reduced N inputs

resulting from plant invasion and previously-impacted or

recovering cryptobiotic crusts; alternatively

anthropogenic N deposition may aggravate N loss

through increased ammonia volatilization and

denitrification.

EXOTIC PLANTS

Introduced annuals such as Bromus tectorum

(cheatgrass) and Taeniatherum asperum (medusahead)

appear to impose long-term threats to biological soil

RECOVERY FROM DISTURBANCE

Natural Recovery Rates

The recovery of biological crusts and vascular plants

appear to occur in concert, rather than as individual

169
"steps" in the recovery process (Danin and others

1989; Johansen and others 1993; Kaltenecker 1997).

The biological crust stabilizes the soil surface and

increases nutrient availability to enhance

reestablishment vascular plants (Belnap and Harper

1995; Williams and others 1995). Vascular plants also

protect the soil surface. In addition they provide

shading and curb wind speed at the surface, providing

conditions conducive to development of the biological

crust (Danin and others 1989). As the community

develops, the shaded areas under plant canopies and

open interspaces between plants allow diversification in

the crust. Microsites created by the uneven

topography of the crust serve as safe sites for vascular

plant seeds.

Species composition-Recovery rates of biological

soil crusts depend on the type and extent of

disturbance, the availability of nearby inoculation

material, as well as on the temperature and moisture

regimes that follow disturbance events. Recovery time

is faster when crustal material is not removed, as

pieces of remaining organisms remain to re-inoculate

recovering surfaces. Therefore, although most damage

is done with the initial impact, recovery will be faster if

disturbances are not repeated. Timing of the

disturbance is also important. Damage is less severe

when crusts are wet. In addition, if damage occurs

when rain is imminent, then crustal organisms have

opportunity to reattach themselves before being blown

away or buried. However, if disturbances occur before

a long dry period, reattachment is not possible and

much crustal material may be lost or too deeply buried

for recovery. Size of disturbance can be important,

especially if crustal material has been lost from the

disturbed site. As inoculant must come from adjoining

areas, the size of the perimeter area relative to the

internal surface area of the disturbance can heavily

influence recovery rates. In addition, recovery is

slower if soils in adjacent areas are destabilized.

Sediments from these areas can either bury adjacent

crusts, leading to their death, or provide material for

"sandblasting" nearby surfaces, thus increasing wind

erosion rates and slowing recovery (Belnap 1995;

McKenna-Neumann and others 1996).

Cyanobacteria or green algae recover first. Microcoleus

is generally the first species to appear. Cyanobacteria

are mobile, and can often move up through disturbed

sediments to reach needed light levels for

photosynthesis, while slow-growing lichens and

mosses are incapable of such movement. Instead, they

require stable soil surfaces for growth, and colonization

of these components generally takes place after

surfaces have been stabilized by cyanobacteria.

Collema, a nitrogen-fixing lichen, is generally the first

lichen to appear.

Cyanobacterial and green algal communities can

recover quickly, especially in regions where effective

precipitation is relatively high. As these organisms are

metabolically active only when wet, the recovery

process is more rapid in regions where soil surface

moisture lasts for a relatively longer period of time.

For instance, green algal communities recovered within

two years of a summer wildfire in the lower Columbia

Basin (Johansen and others 1993). This relatively rapid

recovery is credited to cooler, wetter conditions relative

to the more southerly locations. This includes areas

that receive a majority of annual precipitation as coolseason moisture. Other influential factors might

include mild temperatures, with the majority of

moisture occurring as rain rather than snowfall. Sites

with fine-textured soils such as silt loams retain

surface-soil moisture for a longer period than do coarsetextured, sandy or gravelly soils.

Estimates of time for visually-assessed recovery have

varied from 5 to 100 years (Anderson and others

1982b; Jeffries and Klopatek 1987; Callison and others

1985; Cole 1990). However, Belnap (1993) showed

that many components of recovery can not be

assessed visually. Assuming linear recovery rates,

recovery was estimated to be 35-65 years for

cyanobacterial biomass, 45-85 years for lichen cover,

and 250 years for moss cover in scalped 0.25 m2 plots

surrounded by well-developed crusts. Since recovery

time is dependent on presence of nearby inoculant,

larger disturbed areas will take longer to recover.

Several studies have demonstrated that inoculation can

hasten recovery (Tidemann and others 1980; Ashley

and Rushforth 1984; Belnap 1993).

Nitrogen-fixation-On the Colorado Plateau, recovery

of nitrogenase activity levels in experimentallydisturbed areas is slow. In areas where crusts were

removed, no nitrogenase activity was detectable after 9

years. N content of soils was much lower when

compared to adjacent control plots. In areas disturbed

with 4 wheel drives, no recovery could be documented

after 2 years (Belnap 1996). Thirty years after release

from grazing, soil and plant N and nitrogenase activity

levels were found to be significantly lower when

compared to an area that was never grazed (Evans and

Belnap 1999). Similar results were found for

disturbances in the Mojave desert. Analysis oj

recovery rates for dated disturbances show 2000 years

may be required to restore N inputs at old tank tracks

and old ghost towns. Recovery of nitrogenase activity

in experimentally-disturbed areas in the Great Basin and

southern New Mexico show much faster recovery

trajectories of 20-30 years (Belnap and Herrick

unpublished data).

Albedo-Restoration of normal surface albedos and

temperatures will depend on the restoration of

cyanobacteria, lichens and mosses. While

170
cyanobacteria form a dark matrix in which other

components are embedded, dark mosses and lichens

contribute up to 40% of the cover in an undisturbed

crust (Belnap 1993). Consequently, recovery of

surface albedos will depend on climate and soils.

Fire-Time required for post-fire recovery of the

biological crust depends on a number of factors,

including the size and intensity of the fire, the

composition of the crust (i.e., which types of

organisms are dominant), general climate of the area

and weather immediately following the fire, soil

characteristics, pre-fire condition of the biological crust

and plant community, and post-fire disturbances

(Belnap 1993; Johansen and others 1993). Small or

low intensity fires result in a mosaic of burned and

unburned vegetation, thereby providing vascular plant

and biological crust propagules to colonize the burned

areas. Mosses dominate biological crusts in the

northern Great Basin within two to five years following

fire if the community is not converted to annual

grasses (J. Kaltenecker, unpublished data). Although

the lichens that codominate are slow-growing and may

take several decades to reach pre-fire cover values,

considerable diversity is apparent within about one

decade. Recovery may be enhanced in areas that burn

late in the dry season, as opposed to early-season fires.

Historically, fires occurred in late summer when the

native bunchgrasses were dry enough to burn.

Consequently, the period of time between burning and

cooler, moister weather during which resprouting of

vascular vegetation and growth of the biological crust

occur was probably less than two months long. Since

invasion of exotic annual grasses, fire seasons have

increased in length by one to two months. The longer

interim before conditions conducive to regeneration

increases the vulnerability of the soil surface to

destructive impacts: trampling or crushing by humans,

livestock, or vehicles and the erosive forces of wind or

brief, high intensity summer rain storms.

Enhanced Recovery Rates

The use of inoculants to speed up recovery of crusts

works well (St. Clair and others 1986; Lewin 1977;

Tidemann and others 1980; Ashley and Rushforth

1984). In an experiment reported from SE Utah, all

measured responses were significantly enhanced by

inoculation (Belnap 1993, 1995, 1996).

EVOLUTIONARY HISTORY OF

DISTURBANCE

Soil and plant characteristics of low and mid-elevation

Colorado Plateau ecosystems suggests that they

probably evolved with low levels of soil surface

disturbance by ungulates. These characteristics include

limited surface water, sparse vegetation, the presence

of biological soil crust which are easily disrupted by

trampling, and the dependence of these ecosystems on

nitrogen provided by the biological soil crusts (Evans

and Ehleringer 1993; Evans and Belnap 1999). Dung

beetles, present globally in other systems with large

ungulate populations, are lacking (Mack and Thompson

1982). Limited surface water would have kept

ungulate populations small and generally limited to

winter use of lower elevations, as is seen today

(Parmeter and Van Devender 1995). Winter use results

in lower impacts to biological crust (Marble and Harper

1989) as soils are wet or soon to be wet. Dominant

bunchgrasses that lack adaptations to grazing such as

tillering, secondary compounds, or high tissue silica

content (Mack and Thompson 1982; Martin 1975;

Stebbins 1981). In addition, shallow soils and limited

precipitation limits the distribution of burrowing

vertebrate and invertebrate species. Thus, these

systems may depend more heavily on soil surface

integrity for natural ecosystem functioning than other

regions. As a result, these deserts may be more

negatively affected by soil surface disturbances than

deserts that evolved with higher levels of surface

disturbance.

CONCLUSION

Unfortunately, many activities of man are incompatible

with the presence and well-being of biological soil

crusts. The cyanobacterial fibers that confer such

tensile strength to these crusts are no match for the

compressional stresses placed on them by increased

anthropogenic activities. Crushed crusts contribute

less nitrogen and organic matter to the ecosystem.

Impacted soils are left highly susceptible to both wind

and water erosion. Raindrop erosion is increased, and

overland water flows carry detached material away.

Soil erosion in arid lands is a major threat worldwide.

Beasley and others (1984) estimated that in rangeland

of the United States alone, 3.6 million ha has some

degree of accelerated wind erosion. Relatively

undisturbed biological soil crusts can contribute a great

deal of stability to otherwise highly erodible soils.

Unlike vascular plant cover, crustal cover is not

reduced in drought, and unlike rain crusts, these

organic crusts are present year-round. Consequently,

they offer stability over time and in adverse conditions

that is often lacking in other soil surface protectors.

Unfortunately, disturbed crusts now cover vast areas in

the western United States as a result of ever-increasing

recreational and commercial uses of these semi-arid

and arid areas. Therefore, the tremendous land area

currently being impacted may lead to significant

increases in regional and global soil erosion rates.

171
QUESTIONS FROM THE SYMPOSIUM

FLOOR

Q: Are herds of wild horses, donkeys, and burros

compatible with the stability of semi-arid landscapes?

A: This depends on population numbers, densities,

timing of use, and intensity of use of both native and

introduced large mammals. Before the advent of wells,

irrigation and other water developments, availability of

free surface water, along with predators and forage

availability, naturally limited large mammal populations,

as well as season of use in dry areas. We have

disrupted these natural limitations through water

development, predator control, conversion of

grasslands to shrublands, irrigation and other activities.

Coinsequently, soil surfaces in arid and semi-arid

systems are, in general, receiving much greater use

than they have in the past. Increased densities and

year-round use will compromise the stability of these

surfaces.

Q: Our current situation is millions of acres where

crusts have been disturbed by past 100-150 years of

land use. Society still demands products and land use.

Is it possible to restore crusts on a large scale and still

meet societies demands for land use?

A: It is possible to restore crusts on a large scale.

However, it is also important to realize that there are

many areas that have excellent crust development, and

conservation is much easier and cheaper than

restoration. Land use does not have to be incompatible

with healthy crusts. It is more an issue of social

values, and we need to reassess what values are most

important to us as a society.

Q: When saying "disturbance" was/is there any

distinction in levels/intensities? i.e. any vs. high

disturbance; truck track spinning tires vs. motorcycle or

bicycle tires, etc.

A: Most damage occurs with the first disturbance.

Repeated disturbance will increase compaction, which

has other consequences for hydrologic cycles, soil food

webs, and vascular plants. Continous disturbance,

such as wheeled vehicles produce, are more

problematic than discrete disturbances, such as

footprints, if water is channeled along the disturbance.

Q: So why did the cheatgrass establish in Virginia Park

with the optimal crust conditions present?

A: Cheatgrass requires humidity to germinate. In this

dry, arid region, humidity is achieved by very minimal

burial by soil or organic matter, which does not

generally happen in a heavily crusted, native bunch

grassland, where soils are stabilized and surface litter is

minimal. However, rainfall in the fall of 1994 was very

unusual here. Instead of our normal rainfall pattern, it

rained almost daily for several weeks, and was very

cloudy and cool. We speculate that the increased

humidity allowed the cheatgrass to germinate. Through

experimentation, we have replicated this in the lab.

Q: On radical disturbances - what can we do to

enhance crust colonization?

A: Inoculation works really well on all disturbances.

The downside is that you need to disturb other crusts

to obtain material. We try to stay in touch with

proposed construction projects and get there before the

bulldozers to salvage crusts.

Q: Will colonization occur mostly through wind blown

spores or some other mode?

A: Cyanobacteria colonize mostly through wind-blown

material. Lichens and mosses vary with species, but

my observations are that reproduction is mostly

vegetative, when pieces from adjoining areas break off

and roll in.

Q: Can we artificially transport - transplant? Please

refer to Jane's ecology cookbook.

A: Yes. It is simple. You can scalp off the top 2 mm

of soil. If collected when dry and kept dry, crustal

organisms will stay alive for years.

LITERATURE CITED

Abrahams, A.D.; Parsons, A.J.; Luk, K.S. 1988.

Hydrologic and sediment responses to simulated

rainfall on desert hillslopes in southern Arizona.

Catena 15:105-117.

Alexander, R.W.; Calvo, A. 1990. The Influence of

lichens on slope processes in some Spanish

badlands. In: Thornes, J.B. (ed.). Vegetation and

Erosion. John Wiley and Sons, Ltd. Chichester,

England. Pages 385-398.

Anantani, Y.S.; Marathe, D.V.. 1972. Soil aggregating

effects of some algae occurring in the soils of Kutch

and Rajasthan. Journal of the University of

Bombay 41:94-100.

Anderson, D.C.; Rushforth, S.R. 1976. The

cryptogamic flora of desert soil crusts in southern

Utah. Nova Hedwigia 28:691-729.

Anderson, D.C.; Harper, K.T.; Holmgren, R.C. 1982a.

Factors influencing development of cryptogamic

172
crusts in Utah deserts. Journal of Range

Management 35:180-185.

Anderson, D.C.; Harper, K.T.; Rushforth, S.R. 1982b.

Recovery of cryptogamic soil crusts from grazing in

Utah winter ranges. Journal of Range Management

35:355-359.

Ashley, J.; Rushforth, S.R.. 1984. Growth of soil

algae on topsoil and processed oil shale from the

Uintah Basin, Utah, USA. Reclamation and

Revegetation Research 3:49-63.

Beasley, R.P.; Gregory,M.; McCarty, T.R. 1984.

Erosion and Sediment Pollution Control, 2nd ed,

Iowa State University Press.

Belnap, J. 1993. Recovery rates of cryptobiotic soil

crusts: assessment of artificial inoculant and

methods of evaluation. Great Basin Naturalist

53:89-95.

Belnap, J. 1994. Potential role of cryptobiotic soil

crust in semiarid rangelands. In Monsen, S.B.;

Kitchen, S.G. (eds.), Proceedings, Ecology and

Management of Annual Rangelands. USDA Forest

Service, General Technical Report INT-GTR-313.

Pages 179-185.

Belnap, J. 1995. Surface disturbances: their role in

accelerating desertification. Environmental

Monitoring and Assessment 37:39-57.

Belnap, J. 1996. Soil surface disturbances in cold

deserts: effects on nitrogenase activity in

cyanobacterial-lichen soil crusts. Biology and

Fertility of Soils 23:362-367.

Belnap, J.; Gardner, J.S. 1993. Soil microstructure of

the Colorado Plateau: the role of the

cyanobacterium Microcoleus vaginatus. Great Basin

Naturalist 53:40-47.

Belnap, J.; Gillette, D.A. 1997. Disturbance of

biological soil crusts: impacts on potential wind

erodibility of sandy desert soils in SE Utah, USA.

Land Degradation and Development 8:355-362.

Belnap, J.; Gillette, D.A. 1998. Vulnerability of desert

soil surfaces to wind erosion: impacts of soil

texture anddisturbance. Journal of Arid

Environments 39:133-142.

Belnap, J.; Harper, K.T. 1995. Influence of

cryptobiotic soil crusts on elemental content of

tissue in two desert seed plants. Arid Soil Research

and Rehabilitation 9:107-115.

Belnap, J.; Harper, K.T.; Warren, S.D. 1994. Surface

disturbance of cryptobiotic soil crusts: nitrogenase

activity, chlorophyll content and chlorophyll

degradation. Arid Soil Research and Rehabilitation

8:1-8.

Belnap, J.; Barr, D.; Garcia-Pichel, F. 1997. N leakage

in cyanobacterial-lichen soil crusts (abs.). Ecological

Bulletin 78:250.

Belnap, J.; Sanford, R.L.; Lungu, L. 1996. Biological

soil crusts: ecological roles and response to fire in

Miombo Woodlands of Zimbabwe. Transactions of

the Zimbabwe Scientific Association 70:14-20.

Beymer, R.J.; Klopatek, J.M. 1991. Potential

contribution of carbon by microphytic crusts in

pinyon-juniper woodlands. Arid Soil Research and

Rehabilitation 5:187-198.

Beymer, R.J.; Klopatek, J.M. 1992. Effects of grazing

on cryptogamic crusts in pinyon-juniper woodlands

in Grand Canyon National Park. American Midland

Naturalist 127:139-148.

Black, C.A. 1968. Soil-plant relationships, 2nd

edition. John Wiley and Sons, Inc. New York. 790

pp.

Blackburn, W.H. 1983. Influence of brush control on

hydrologic characteristics of range watersheds. In:

McDaniel, K.C. (ed.). Proceedings of the Brush

Management Symposium, Texas Tech Press,

Lubbock. USA. Pp. 73-88.

Bond, R.D. 1964. The influence of the microflora on

the physical properties of soils. II. Field studies on

water repellent sands. Australian Journal of Soil

Research 2:123-131.

Bond, R.D; Harris, J.R. 1964. The influence of the

microflora on physical properties of soils. I. Effects

associated with filamentous algae and fungi.

Australian Journal of Soil Research 2:111-1 22.

Booth, W.E. 1941. Algae as pioneers in plant

succession and their importance in erosion control.

Ecology 22:38-46.

Brotherson, J.D.; Rushforth, S.R. 1983. Influence of

cryptogamic crusts on moisture relationships of soils

in Navajo National Monument, Arizona. Great Basin

Naturalist 43:73-78.

Bush, J.K.; Van Auken, O.W. 1991. Importance of

time of germination and soil depth on growth of

Prosopis glandulosa seedling in the presence of a C4

grass. American Journal of Botany 78:1732-1739.

173
Callison, J.; Brotherson, J.D.; Bowns, J.E. 1985. The

effects of fire on the blackbrush (Coleogyne

ramosissima) community of southwest Utah.

Journal of Range Management 38:535-538.

Campbell, S.E. 1977. Desert crust of Utah: an aridity

adapted algal mat community. 16th Algal

Symposium, 30 April 1977, Woods Hole,

Massachusetts. 11 Pp.

Campbell, S.E.; Seeler, J.S.; Golubic, S. 1989. Desert

crust formation and soil stabilization. Arid Soil

Research and Rehabilitation 3:217-228.

Carleton, T.J. 1990. Variation in terricolous bryophyte

and macro lichen vegetation along primary gradients

in Canadian boreal forests. Journal of Vegetation

Science 1:585-594.

Cole, D.N. 1991. Trampling disturbance and recovery

of cryptogamic soil crusts in Grand Canyon National

Park. Great Basin Naturalist 50:321-325.

Crawford, C.S. 1991. The community ecology of

macroarthropod detritivores. In: Polis, G. (ed.).

Ecology of Desert Communities. University of

Arizona Press. Tucson, Arizona.

Dadlich, K.S.; Varma, A.K.; Venkataraman, G.S. 1969.

The effect of Calothrix inoculation on vegetable

crops. Plant and Soil 31:377-379.

Danin, A. 1978. Plant species diversity and plant

succession in a sandy area in the Northern Negev.

Flora 167:409-422.

Doyen, J.T.; Tschinkel, W.F. 1974. Population size,

microgeographic distribution and habitat separation

in some tenebrionid beetles. Annals of the

Entomological Society of America 67:617-626.

Dregne, H.E. 1983. Desertification of Arid Lands.

Harwood Academic Publishers, New York. Pp. 1-

15.

Dunkerley, D.L.; Brown, K.J. 1995. Runoff and runon

areas in a patterned chenopod shrubland, arid

western New South Wales, Australia:

characteristics and origin. Journal of Arid

Environments 30:41-55.

Eckert, Jr., R.E.,; Peterson, F.F.; Meurisse, M.S.;

Stephens, J.L. 1986. Effects of soil-surface

morphology on emergence and survival of seedlings

in big sagebrush communities. Journal of Range

Management 39(5):414-420.

Eldridge, D.J. 1993a. Cryptogams, vascular plants

and soil hydrological relations: some preliminary

results from the semiarid woodlands of eastern

Australia. Great Basin Naturalist 53(1):48-58.

Eldridge, D.J. 1993b. Cryptogam cover and soil

surface condition: effects on hydrology on a

semiarid woodland soil. Arid Soil Research and

Rehabilitation 7:203-217.

Eldridge, D.J.; Greene, R.S.B. 1994. Assessment of

sediment yield by splash erosion on a semi-arid soil

with varying cryptogam cover. Journal of Arid

Environments 26:221-232.

Danin, A.; Barbour, M.B. 1982. Microsuccession of

cryptogams and phanerogams in the Dead Sea Area,

Israel. Flora 172:173-179.

Danin, A.; Bar-or, Y.; Dor, I.; Yisraeli, T. 1989. The

role of cyanobacteria in stabilization of sand dunes

in Southern Israel. Ecologia Mediterranea XV:55-

64.

Danin, A.; Yaalon, D.H. 1980. Trapping of silt and

clay by lichens and bryophytes in the desert

environment of the Dead Sea region. Bat Sheva

Seminar on Approaches and Methods in

Paleoclimatic Research with Emphasis on Aridic

Areas, Jerusalem. 32 Pp.

Dobrowolski, J.P. 1994. In situ estimation of

effective hydraulic conductivity to improve erosion

modeling for rangeland conditions. In: Variability of

Rangeland Water Erosion Processes. SSSA

Publication 38, Soil Science Society of America,

677 S. Segoe Rd., Madison, Wl 53711, USA. Pp.

83-91.

Evans, R.D.; Belnap, J. 1999. Long-term

consequences of disturbance on nitrogen dynamics

in an arid ecosystem. Ecology. In press.

Evans, R.D.; Ehleringer, J.R. 1993. A break in the

nitrogen cycle in aridlands? Evidence from N15 of

soils. Oecologia 94:314-317.

Farnsworth, R.B.; Romney, E.M.; Wallace, A. 1976.

Implications of symbiotic nitrogen fixation by desert

plants. Great Basin Naturalist 36(1):65-80.

Faust, W.F. 1970. The effect of algal-mold crusts on

the hydrologic processes of infiltration, runoff, and

soil erosion under simulated rainfall conditions.

M.S. Thesis, University of Arizona, Tucson, AZ. 60

Pp.

Faust, W.F. 1971. Blue-green algal effects on some

hydrologic processes at the soil surface. In:

Hydrology and Water Resources in Arizona and the

Southwest, Proceedings from the 1971 meetings of

the Arizona Section, American Water Resource

174
Association, and Hydrology Section of the Arizona

Academy of Sciences, Tempe, AZ. Pp. 99-105.

Fletcher, J.E.; Martin, W.P. 1948. Some effects of

algae and molds in the raincrust of desert soils.

Ecology 29:95-100.

Foth, H.D. 1978. Fundamentals of Soil Science. 6th

edition. John Wiley and Sons. London, New York.

Fryberger, S.C.; Schenk, C.J.; Krystinik, L.F. 1988.

Stokes surfaces and the effects of near surface

groundwater-table on aeolian deposition.

Sedimentol. 35:21-41.

Garcia-Pichel, F.; Bebout, B.M. 1996. The penetration

of ultraviolet radiation into shallow water sediments:

high exposure for photosynthetic communities.

Marine Ecology Progress Series 131:257-262.

Garcia-Pichel, F.; Belnap, J. 1996. The

microenvironments and microscale productivity of

cyanobacterial desert crusts. Journal of Phycology

32:774-782.

Gayel, A.G.; Shtina, E.A. 1974. Algae on the sands of

arid regions and their role in soil formation. Soviet

Soil Science 6:311-319.

Gillette, D.A.; Dobrowolski, J.P. 1993. Soil crust

formation by dust deposition at Shaartuz, Tadzhik,

S.S.R. Atmospheric Environment 27A:251 9-2525.

Goudie, A.S. 1978. Dust storms and their

geomorphological implications. Journal of Arid

Environments 1:291-310.

Graetz, R.D.; Tongway, D.J.U. 1986. Influence of

grazing management on vegetation, soil structure

and nutrient distribution and the infiltration of

applied rainfall in a semi-arid chenopod shrubland.

Australian Journal of Ecology 11:347-360.

Greene, R.S.B.; Chartes, C.J.; Hodgkinson, K.C. 1990.

The effects of fire on the soil in a degraded semiarid woodland. I. Cryptogam cover and physical

and micromorphological properties. Australian

Journal of Soil Research 28:755-777.

Harper, K.T.; Marble, J.R. 1988. A role for

nonvascular plants in management of arid and

semiarid rangeland. In: Tueller, P.T. (ed.).

Vegetation Science Applications for Rangeland

Analysis and Management. Kluwer Academic

Publishers, Dordrecht. Pp 135-169.

Harper, K.T.; St. Clair, L.L. 1985. Cryptogamic soil

crusts on arid and semiarid rangelands in Utah:

effects of seedling establishment and soil stability.

Final report, Bureau of Land Management, Utah

State Office, Salt Lake City.

Harper, K.T., Pendleton, R.L. 1993. Cyanobacteria

and cyanolichens: can they enhance availability of

essential minerals for higher plants? Great Basin

Naturalist 53:59-72.

Hormay, A.L. 1970. Principles of rest-rotation grazing

and multiple-use land management. Forest Service

Training Text 4(2200). U.S. Department of

Agriculture, Forest Service.

Jeffries, D.L.; Klopatek, J.M. 1987. Effects of grazing

on the vegetation of the blackbrush association.

Journal of Range Management 40(5):390-392.

Johansen, J.R. 1993. Cryptogamic crusts of semiarid

and arid lands of North America. Journal of

Phycology29:140-147.

Johansen, J.R. 1986. Importance of cryptogamic soil

crusts to arid rangelands: implications for short

duration grazing. In: Tiedman, J.A. (ed.). Short

Duration Grazing. Washington State University,

Pullman, WA. Pp 127-136.

Johansen, J.R.; Ashley, J.; Rayburn, W.R. 1993. The

effects of rangefire on soil algal crusts in semiarid

shrub-steppe of the Lower Columbia Basin and their

subsequent recovery. Great Basin Naturalist 53:73-

88.

Johansen, J.R.; Rushforth, S.R.; Brotherson, J.D.

1981. Sub-aerial algae of Navajo National

Monument, Arizona. Great Basin Naturalist 41:433-

439.

Johansen, J.R.,; St. Clair, L.L.; Webb, B.L.; Nebeker,

G.T. 1984. Recovery patterns of cryptogamic soil

crusts in desert rangelands following fire

disturbance. Bryologist 87:238-243.

Johnson, C.W.; Gordon, N.D. 1986. Runoff and

erosion from rainfall simulator plots on sagebrush

rangeland. In: Proceedings of the Fourth

Interagency Sedimentation Conference, Las Vegas,

NV. Vol. 1. Pp 3-132 to 3-141.

Johnson, C.W.; Rosentreter, R. 1988. Influence of

cryptogamic ground cover on rangeland soil loss.

In: UNEP Workshop: Uses of Microbiological

Processes in Arid Lands for Desertification Control

and Increased Productivity. ISEB Symposium, 6-12

October 1988, Albuquerque and Santa Fe, NM.

Kaltenecker, J.H. 1997. The recovery of microbiotic

crusts following post-fire rehabilitation on

175
rangelands of the western Snake River Plain. M.S.

thesis, Boise State University, Boise, ID. 99 Pp.

Kappen, L.; Lange, O.L.; Schulze, E.D.; Evenari, M.;

Buschbom, U. 1979. Ecophysiological

investigations on lichens of the Negev Desert. VI.

Annual course of the photosynthetic production of

Ramalina maciformis (Del.) Bory. Flora 168:85-108.

Kinnell, P.I.A.; Chartres, C.J.; Watson, C.L. 1990.

The effects of fire on the soil in a degraded semiarid woodland. Susceptibility of the soil to erosion

by shallow rain-impacted flow. Australian Journal

of Soil Research 28:755-777.

Kleiner, E.F.; Harper, K.T. 1972. Environment and

community organization in grasslands of

Canyonlands National Park. Ecology 53:229-309.

Kleiner, E.F.; Harper, K.T. 1977a. Occurrence of four

major perennial grasses in relation to edaphic factors

in a pristine community. Journal of Range

Management 30(4):286-289.

Kleiner, E.F.; Harper, K.T. 1977b. Soil properties in

relation to cryptogamic groundcover in Canyonlands

National Park. Journal of Range Management

30(3):202-205.

Ladyman, J.A.R.; Muldavin, E. 1994. A study of the

terricolous cryptogam and other ground cover in low

disturbance pinyon-juniper woodlands in New

Mexico. Unpublished report to the USDA Forest

Service. Albuquerque, New Mexico.

Ladyman, J.A.R.; Mudavin, E.; Fletcher, R.; Aldon, E.

1994. An examination of three mesas to compare

and contrast the relationships between terrestrial

cryptogam and vascular plant cover. American

Academy for the Advancement of Science. In:

Proceedings, Southwestern and Rocky Mountain

Division 70th Abstract 57:25.

Lange, W. 1974. Chelating agents and blue-green

algae. Canadian Journal of Microbiology 20:1311-

1321.

Lange, O.L. 1980. Moisture content and C02

exchange of lichens. I. Influence of temperature on

moisture-dependent net photosynthesis and dark

respiration in Ramalina maciformis. Oecologia

45:82-87.

Lange, O.L.; Belnap, J.; Reichenberger, H.; Meyer, A.

1997. Photosynthesis of green algal soil crust

lichens from arid lands in southern Utah, USA: role

of water content on light and temperature responses

of CO2 exchange. Flora 192:1-15.

Larmuth, J. 1978. Temperatures beneath stones used

as daytime retreats by desert animals. Journal of

Arid Environments 1:35-40.

Lemos, P.; Lutz, J.F. 1957. Soil crusting and some

factors affecting it. Soil Science Society of America

Proceedings 21:485-491.

Lesica, P.; Shelley, J.S. 1992. Effects of cryptogamic

soil crust on the population dynamics of Arabis

fecunda (Brassicaceae). American Midland

Naturalist 128:53-60.

Lewin, R.A. 1977. The use of algae as soil

conditioners. CIBCASIO Trans. 3:33-35.

(Publication of Centros de Investigacion de Baja

California, Scripps Institute of Oceanography,

Lajolla, California).

Leys, J.N. 1990. Soil crusts: their effect on wind

erosion. Research Note 1/90. Soil Conservation

Service of New South Wales, Australia.

Loope, W.L.; Gifford, G.F. 1972. Influence of a soil

microfloral crust on select properties of soils under

pinyon-juniper in southeastern Utah. Journal of Soil

Water and Conservation 27:164-167.

Lusby, G.C. 1979. Effects of grazing on runoff and

sediment yield from desert rangeland at Badger

Wash in Western Colorado, 1953-1973. Geological

Survey Water-Supply Paper 1532-1. US Government

Printing Office, Washington, D.C.

McKenna-Neuman, C; Maxwell, CD.; Boulton, J.W.

1996. Wind transport of sand surfaces crusted

with photoautotrophic microorganisms. Catena

27:229-247.

Mabbutt, J.A.; Fanning, P.C. 1987. Vegetation

banding in arid Western Australia. Journal of Arid

Environments 12:41-59.

Mack, R.N.; Thompson, J.N. 1982. Evolution in

steppe with few large, hooved mammals. The

American Naturalist 119:757-773.

Lange, O.L.; Belnap, J.; Reichenberger, H. 1998.

Photosynthesis of the cyanobacterial soil-crust

lichen Collema tenax from arid lands in southern

Utah, USA: role of water content on light and

temperature responses of CO2 exchange. Functional

Ecology 12:195-202.

Magee, W.E.; Burris, R.H. 1954. Fixation of N2 and

utilization of combined nitrogen by Nostoc

muscorum. American Journal of Botany 41:777-

782.

176
Marathe, K.V. 1972. Role of some blue-green algae in

soil aggregation. In: Desikachary, T.V., ed.

Taxonomy and Biology of Blue-green Algae.

University of Madrias, India. Pp 328-331.

Marble, J.R.; Harper, K.T. 1989. Effect of timing of

grazing on soil-surface cryptogamic communities in

Great Basin low-shrub desert: A preliminary report.

Great Basin Naturalist 49:104-107.

Martin, P.S. 1975. Vanishings, and future of the

prairie. Geoscience and Man 10:39-49.

Mayland, H.F.; Mclntosh, T.H. 1966. Availability of

biologically fixed atmospheric nitrogen-15 to higher

plants. Nature 209:421-422.

Meyer, S.E.; Garcia-Moya, E. 1989. Plant community

patterns and soil moisture regime in gypsum

grasslands of north central Mexico. Journal of Arid

Environments 16:147-155.

Muncher, H.J.; Chartres, C.J.; Tongway, D.J.U.;

Greene, R.S.B. 1988. Micromorphology and

significance of the surface crusts of soils in

rangelands near Cobar, Australia. Geoderma

42:227-244.

Nash, III, T.H.; Lange, O.L.; Kappen, L. 1982.

Photosynthetic patterns of Sonoran desert lichens.

II. A multivariate laboratory analysis. Flora

172:419-426.

Paerl, H.W. 1990. Physiological ecology and

regulation of N2 fixation in natural waters.

Advanced Microbial Ecology 2:305-344.

Parmenter, R.R.; Van Devender, T.R. 1995. Diversity,

spatial variation, and functional roles of vertebrates

in the desert grassland. In: McClaran, M.P.; Van

Devender, T.R. (eds.). Desert Grasslands.

University of Arizona Press, Tucson. Pp 196-229.

Peterjohn, W.T.; Schlesinger, W.H. 1990. Nitrogen

loss from deserts in the southwestern United

States. Biogeochemistry 10:67-79.

Peters, E.F.; Bunting, S.C. 1994. Fire conditions preand postoccurrence of annual grasses on the Snake

River Plain. In: Monsen, S.B.; Kitchen, S.G. (eds.).

Proceedings, Ecology and Management of Annual

Rangelands. eneral Technical Report INT-GTR-313.

U.S. Department of Agriculture, Forest Service,

Intermountain Research Station, Ogden, Utah. Pp

31-36.

Pierson, Jr., F.B.; Blackburn, W.H.; Van Vactor, S.S.;

Woods, J.C. 1994a. Partitioning small scale spatial

variability or runoff and erosion on sagebrush

rangeland. Water Resources Bulletin 30(6):1081-

1089.

Pierson, Jr., F.B.; Blackburn, W.H.; Woods, J.C.

1994b. Incorporating small scale spatial variability

into predictions of hydrologic response on sagebrush

rangelands. In: Variability of Rangeland Water

Erosion Processes. SSSA Publication 38, Soil

Science Society of America, 677 S. Segoe Rd.,

Madison, Wl 5371 1, USA. Pp 23-34.

Platou, K.A.; Tueller, P.T. 1985. Evolutionary

implications for grazing management systems.

Rangelands 7:57-61

Rogers, R.W. 1972. Soil surface lichens in arid and

subarid south-eastern Australia. Australian Journal

of Botany 20:301-316.

Rogers, R.W. 1977. Lichens of hot arid and semi-arid

lands. In: Seaward, M.R.D. (ed.). Lichen Ecology.

Academic Press, New York.

Rosentreter, R. 1986. Compositional patterns within a

rabbitbrush (Chrysothamnus) community of the

Idaho Snake River plain. In: Proceedings,

Symposium on the Biology of Artemisia and

Chrysothamnus. General Technical Report INT-200:

USDA Forest Service, Intermountain Research

Station. Ogden, Utah. Pp 213-211.

Rychert, R.C.; Skujins, J.; Sorensen, D.; Porcella, D.

1978. Nitrogen fixation by lichens and free-living

microorganisms in deserts. In: Nitrogen Fixation in

Desert Ecosystems, West, N.E.; Skujins, J. (eds.).

Dowden, Hutchinson and Ross. Stroudsburg,

Pennsylvania. Pp 20-30.

St. Clair, L.L.; Johansen, R.; Webb, B.L. 1986. Rapid

stabilization of fire-disturbed sites using a soil crust

slurry: inoculation studies. Reclamation and

Revegetation Research 4:261-269.

St. Clair, L.L.; Webb, B.L.; Johansen, J.R.; Nebeker,

G.T. 1984. Cryptogamic soil crusts: enhancement

of seedling establishment in disturbed and

undisturbed areas. Reclamation and Revegetation

Research 3:129-136.

Sagan, C; Toon, O.B.; Pollack, J.B. 1979.

Anthropogenic albedo changes and the earth's

climate. Science 206:1 363-1 368.

Savory, A.; Parsons, S.D. 1980. The Savory grazing

method. Rangelands 2:234-237.

Schulten, J.A. 1985. Soil aggregation by cryptogams

of a sand prairie. American Journal of Botany

72:1657-1661.

177
Shields, L.M.; Durrell, L.W. 1964. Algae in relation to

soil fertility. The Botanical Review 30:92-128.

Stanley, R.J. 1983. Soils and vegetation: An

assessment of current status. In: Messer, J.;

Mosley, B. (eds.). What future for Australia's Arid

Land? Australian Conservation Foundation,

Canberra. Pp 8-18.

Stebbins, G.L. 1981. Coevolution of grasses and

herbivores. Annuals of the Missouri Botanical

Garden 68:75-86.

Tchoupopnou, E. 1989. Experimental studies of rain

splash erosion from soil microphytic crusts on Utah

rangelands. M.S. thesis, Utah State University,

Logan, UT.

Tidemann, A.R.; Lopushinsky, W.; Larsen, Jr, H.J.

1980. Plant and soil responses to a commercial

blue-green algae inoculant. Soil Biology and

Biochemistry 12:471-475.

Tongway, D.J.; Ludwig, J.A. 1990. Vegetation and

soil patterning of semi-arid mulga lands of Eastern

Australia. Australian Journal of Ecology 15:23-34.

Verrecchia, E.; Yair, A.; Kidron, G.J.; Verrecchia, K.

1995. Physical properties of the psammophile

cryptogamic crust and their consequences to the

water regime of sandy soils, north-western Negev

Desert, Israel. Journal of Arid Environments

29(4):427-437.

Verstraete and Schwartz. 1991. Desertification and

Global Climate Change. Vegetatio 91:3-13.

Wallwork, J.A. 1982. Desert Soil Fauna. Praeger

Scientific Publishers, London.

Webb, R.H.; Wilshire, H.G. (eds.) 1983.

Environmental effects of off-road vehicles: impacts

and management in arid regions. Springer-Verlag,

New York.

Whisenant, S.G. 1990. Changing fire frequencies on

Idaho's Snake River Plains: ecological and

management implications. In: McArthur, E.D.;

Romney, E.M.; Smith, S.D.; Tueller, P.T. (eds.).

Proceedings, Symposium on Cheatgrass Invasion,

Shrub Die-off, and Other Aspects of Shrub Biology

and Management. General Technical Report INT276. U.S. Department of Agriculture, Forest

Service, Intermountain Research Station, Ogden,

Utah. Pp 4-10.

Williams, J.D. 1993. Influence of microphytic crusts

on selected soil physical and hydrologic properties in

the Hartnet Draw, Capitol Reef National Park, Utah.

Ph.D. Dissertation, Utah State University, Logan,

UT. 170 Pp.

Williams, J.D.; Dobrowolski, J.P.; West, N.E.; Gillette,

D.A. 1995a. Microphytic crust influences on wind

erosion. Transactions of the American Society of

Agricultural Engineers 38:131-137.

Williams, J.D.; Dobrowolski, J.P.; West, N.E. 1995b.

Microphytic crust influence on interrill erosion and

infiltration. Transactions of the American Society of

Agricultural Engineers 38:139-146.

Wullstein, L.H. 1989. Evaluation and significance of

associative dinitrogen fixation for arid soil

rehabilitation. Arid Soil Research Rehabilitation

3:259-265.

Yair, A. 1990. Runoff generation in a sandy area-the

Nizzana sands, western Negev, Israel. Earth

Surface Processes and Landforms 15:597-609.

178
THE ROLE OF SOIL ORGANISMS IN RESTORATION

Michael P. Amaranthus1, Efren Cazares2 and David A. Perry3

ABSTRACT

The soil biological community is inevitably impacted on degraded sites, and much remains to be learned about these

impacts. This myriad of soil organisms and their interactions profoundly affect forest restoration through capture and

uptake of nutrients, nitrogen fixation, protection against pathogens, maintenance of soil structure, and buffering against

moisture stress. The best approach for managers is prevent degradation of the soil biological community. The vast

majority of soil organisms live and work in rhizospheres, humus, rotten wood, and the upper mineral soil which are

powerhouses of soil biological activity and thus are essential substrates for maintaining site resilience. Although there

is no "magic bullet" for enhancing restoration, there is great opportunity to use soil organisms as "tools." Research is

now underway to identify organisms best adapted to specific environmental and biotic conditions and to assess the

potential for "managing" these organisms. Incorporating the concept of the "living soil" in evaluation of site restoration

is part of a new, expanding view of forest ecosystems. Understanding soil organisms and putting them to wise use,

practicing foresters can better not only maintain populations of beneficial organisms on site but also achieve restoration

success.

Keywords: soil restoration, soil organisms, ectomycorrhizae, disturbance, reforestation

INTRODUCTION

In widely varying ecosystems throughout the world,

removal of indigenous vegetation through clearcutting,

mining, urbanization, shifting cultivation, or overgrazing

has led to deterioration in productive capacity, often

accompanied by alterations of plant communities that

appear permanent. Degradation is well documented in

moist tropical forests (Arnason and others 1982), in

the semiarid Miombo woodlands of southern Africa

(Maghembe 1987, personal communication), in the

montane coniferous forests of northern India (Sharma

1983), and in montane forests of North America

(Amaranthus and Perry 1987, 1989). Despite widely

varying community types and environmental conditions,

the loss of soil organisms has been linked to the decline

of these ecosystems.(Perry and others 1989).

Reestablishing beneficial soil organisms has facilitated

restoration of some sites, and is likely to aid restoration

of others as well. The soil biological community is

inevitably impacted on degraded sites, and much

remains to be learned about these impacts. Which

organisms are important to site recovery following

disturbance? How do management activities affect

these populations? What are the rates of recovery?

Can these organisms be "reinoculated" to improve

restoration success? Answers to these important

questions have come slowly, deterred by difficulty in

identifying organisms, enormous variability within and

among sites, and unresolved methodologies (Trappe

and Bollen 1979). However, our growing knowledge of

the workings of this complex belowground forest

system has already enabled us to better predict the

consequences of our actions.

In this paper we (1) briefly review the variety,

abundance, and interactions of organisms in forest

soils, (2) examine the important functions of these

organisms in the context of restoration (3) consider the

factors that adversely affect populations of beneficial

organisms-and thereby restoration-on managed forest

lands, and (4) look at the potential for managing soil

organisms to enhance restoration activities.

VARIETY, ABUNDANCE, AND

INTERACTIONS OF SOIL ORGANISMS

The life forms residing and interacting in forest soil

range from the smallest, most primitive living

organisms all the way up the evolutionary ladder to

mammals. Estimates of the types and numbers of

these organisms, which vary in function and are

strongly interdependent, are staggering. For example,

10-100 million bacteria and actinomycetes, 1000-

100,000 fungal propagules, and several kilometers of

1

 Michael P. Amaranthus is a research biologist, U.S. Department of Agriculture, Forest Service, Pacific Northwest Research

Station, Box 440, Grants Pass, OR 97526.

2Efren Cazares is a research associate, Department of Forest Science, Oregon State University, Corvallis, OR 97331.

3David A. Perry is a professor, Department of Forest Science, Oregon State University, Corvallis, OR 97331.

179
fungal hyphae can be present in a single gram of forest

soil (Bollen 1974). Mites and nematodes can number

from thousands to hundreds of thousands per square

meter of surface soil, and larger soil organisms such as

worms, insects, and animals from 10 to well over 100

(Dindal 1973). The resulting biomass reflected in these

numbers is immense. For example, in the Pacific

Northwest, the soil of an old-growth Douglas-fir

[Pseudotsuga menziesii (Mirb.) Franco] stand can

contain some 4,200 kg/ha (3700 Ib/acre) dry weight of

fungal mycelium and 5,400 kg/ha (4800 Ib/acre) of

mycorrhizal tips alone (Fogel and Hunt 1983).

The balance of forest-soil organisms can shift

dramatically in response to fluctuations of chemical,

environmental, and biotic factors caused by natural

disturbance or management-related activities such as

timber harvest, site preparation, mining, overgrazing,

land clearing and fertilization (Perry and Rose 1983,

Amaranthus and Perry 1987). Although some soil

organisms thrive in the aftermath of disturbance (Pilz

and Perry 1984), those that cannot compete may

decline, lose vigor, or disappear altogether (Perry and

others 1982, Harvey and others 1986, Amaranthus and

Perry 1987). The final equilibrium of soil organisms

may or may not facilitate rapid revegetation of the site.

To assess the potential impact of disturbance, we must

understand where the vast majority of soil organisms

live and work. Humus, rotten wood, and the upper

mineral soil are the powerhouses of soil biological

activity (Harvey and others 1979, 1986) and thus are

essential substrates for maintaining forest productivity.

Densities of fine roots in these substrates and

populations of organisms near these root surfaces are

higher than in surrounding soil (Rambelli 1973). It is in

the rhizosphere, the 2-3 mm zone surrounding root

surfaces, that biological activity is greatest. Fine roots

maneuvering through the soil exude carbohydrates,

amino acids, and other compounds that stimulate the

growth of microflora such as bacteria, actinomycetes,

and fungi, which in turn produce their own compounds

that either stimulate or repel other soil organisms.

Microflora are also prime food for "grazer" herbivores

such as mites, nematodes, and springtails, which

themselves fall prey to carnivores such as centipedes

and spiders. Saprophytic organisms feed on the dead

remains of other microbes and roots accumulating in

the rhizosphere and decompose complex organic

molecules into basic components. Nutrients released

through decomposition, as well as water, are captured

and transported to host plants by specialized fungi that

colonize root tips in a symbiotic association termed

mycorrhizae (Harley and Smith 1983, Read and others

1985). The area immediately surrounding mycorrhizae,

the "mycorrhizosphere" (Rambelli 1973), supports its

own increased levels of microbial activity. Rhizosphere

and mycorrhizosphere populations may vary among

plant species and are further influenced by season and

host maturity (Foster and Marks 1967).

Litter and woody materials are also important centers

of activity. Plant tissue fallen to the forest floor is

attacked by soil organisms called "detritivores". These

organisms, including mites, millipedes, snails, and

beetles, tear and chew holes in fresh litter, increasing

its surface area and utilizing its more soluble

components. The remnants are decomposed by other

soil flora and fauna. Wood fallen to the forest floor is

attacked by beetles and a succession of fungi that

"digest" the cellulose and lignin. Insects, earthworms,

amphibians, reptiles, and mammals likewise

concentrate in and around this woody debris,

enhancing site productivity by mixing and aerating the

soil and by transporting and excreting nutrients and the

spores of mycorrhizal fungi from one microsite to

another (Maser and others 1978).

Woody debris is a dynamic component of the forest

soil. It provides a storehouse of moisture and is an

energy source and refuge for microorganisms critical to

forest restoration. A significantly higher proportion of

mycorrhiza activity occurs in decaying wood compared

to soils (Harvey and others 1979). The relative

importance of woody debris in supporting feeder roots

may be greater on dry sites than on moist sites. During

periods of adequate moisture, humus supports the

highest level of ectomycorrhiza activity; but during

periods of drought, soil wood becomes the most active

site (Harvey and others 1986). Thus, the wood

component appears most important on low-productivity

sites where moisture is limited or where indigenous

levels of woody debris are low. Wood on the forest

floor forms long-lasting, moist microsites that may aid

forest recovery. Following intense wildfire in

southwest Oregon, decaying logs retained 25 times

more moisture than surrounding soil (Amaranthus and

others 1989). Such decaying logs may expedite forest

recovery by providing important refuge for roots and

associated mycorrhizal fungi of pioneering vegetation.

However, accumulated woody debris, if overly

abundant, represents a fuel hazard and increases the

potential for wildfire. So there can be too much as

well as too little woody debris. A balance must be

struck between fuel management guidelines and

protection of large woody debris if forest productivity is

to be maintained.

An appreciation of the importance of the plant-soil

relationship can be acquired from looking at the amount

of energy plants divert to soils. In temperate forests,

from 50 to 80% of the net carbon fixed in

photosynthesis goes to roots and associated organisms

such as mycorrhizal fungi (Fogel and Hunt 1983, Perry

and others 1987). Energy flowing into the soil

supports a diverse community of soil organisms which

in turn influence plant growth through effects on

180
nutrient cycling, water uptake, pathogens, and soil

structure.

Energy flowing to soil organisms may have a specific

cost/return for participating host plants. This is likely to

vary in time and space and with changes in the

physical and biotic environment. Different organisms

do different jobs for different hosts in different

environments. Some organisms, like mycorrhizal fungi

and nitrogen fixers, benefit host plants in clearly

definable ways. Others may benefit hosts only during

periods of environmental extremes, at certain points in

plant development, following disturbance, or indirectly

by influencing important ecosystem properties such as

soil structure. Maintaining native populations of soil

organisms helps minimize site degradation by assuring

a diversity of plant responses to unpredictable or

varying environments.

IMPORTANT FUNCTIONS OF SOIL

ORGANISMS

Nutrient Cycling

Most woody host plants require soil organisms to

facilitate adequate nutrient uptake. Mycorrhizae

enhance such uptake not only by increasing the

absorbing surface area of roots but also through active

physiological mechanisms. Ectomycorrhizal fungi

release enzymes which increase the availability of

phosphorus to higher plants (Williamson and Alexander

1975, Ho and Zak 1979, Alexander and Hardy 1981).

This "extraction" process also extends to other

nutrients and enters them into the forest nutrient cycle.

Mycorrhizal fungi, common to members of the

Pinaceae, release compounds (chelating agents called

siderophores) that are especially important in iron

nutrition (Graustein and others 1977, Powell and others

1980). Perry and others (1982) found lower levels of

siderophores in soils from both burned and unburned

clearcuts than in undisturbed forest soil from 8 of 10

widely scattered sites in southwest, central, and eastcentral Oregon; the reduction in siderophores was

related to poor iron nutrition of Douglas-fir seedlings.

See Allen (1991) for a general overview on the ecology

of mycorrhizae.

Conservation of nutrient capital is necessary to prevent

site degradation. Few nutrients leach out when

populations of soil organisms are healthy and active.

This is particularly significant for soluble forms of

nitrogen (N) such as nitrate, which is mobile and

susceptible to leaching. Soil organisms form an

intricate web to capture and assimilate N into complex

organic compounds and then slowly release them into

the ecosystem. This capture, assimilation, and release

are critical to the nitrogen capital of forest soils.

Nitrogen Fixation

Inadequate nitrogen is a common limitation on

degraded sites. Loss of N due to timber harvest,

removing harvest residues, and site preparation

(especially intense burning) often exceeds natural

inputs to the forest (Perry and Rose 1987). Perhaps

the most practical and economical means of

replenishing this loss is through biological fixation of

nitrogen in the soil (Domingo 1983, Gordon and Avery

1985).

Symbiotic N fixation can add substantial amounts of

nitrogen continuously to Pacific Northwest forests

(Wollum and Youngberg 1964, Trappe and others

1968). Common nodulated plants such as lupine,

alder, and snowbrush (Lupinus, A/nus, and Ceanothus

species) form a mutually beneficial relationship with

certain bacteria and actinomycetes that convert ("fix")

atmospheric N2 into ammonium N. This fixed nitrogen

is released into the roots of host plants, thereby

increasing N concentrations in living tissue (Tarrant and

Trappe 1971). As N is returned to the soil by litterfall

and washing of leaves by rain, other species—including

tree species valued for timber-reap benefits.

Through asymbiotic N fixation, "free-living" organisms

contribute to the nitrogen pool. Nitrogen-fixing

bacteria have been found in association with woodrotting fungi in coniferous forest residues (Larsen and

others 1980). Azotobacter and Clostridium species fix

nitrogen under conditions characteristic of the

rhizosphere, where oxygen levels are low (Giller and

Day 1985). Fixation by these and other species, such

as the bacterium Azospirillum, may be enhanced by or

may totally depend on mycorrhizal fungi (Amaranthus

and others 1990). Use of these free-living organisms

to improve nitrogen levels is now receiving considerable

attention and deserves further research.

Protection Against Pathogens

Both symbiotic and free-living soil organisms can

protect trees against soil pathogens. Several

mycorrhizal fungi protect pine (Pinus) species and

Douglas-fir from pathogens such as Phytophthora

cinnamomi, Fusarium oxysporum, and Rhizoctonia

solani (Wingfield 1968, Marx 1973, Marx and Krupa

1978). The protective mechanisms include physical

barriers against root penetration by pathogens,

excretions of antibiotics, and stimulation of other

pathogen-inhibiting rhizosphere organisms (Zak 1964,

Marx 1972). Rose and others (1980) found that a freeliving Streptomyces species from the rhizosphere of

snowbrush (Ceanothus velutinus Dougl.) antagonized

three common root pathogens, Phellinus weirii, Fomes

annosus, and Phytophthora cinnamomi. Some

nonmycorrhizal fungi also may inhibit pathogens; for

181
example, the fungus species Trichoderma can reduce

the incidence of root rot in pine seedlings (Kelly 1976).

Soil Structure

Soil structure often declines on degraded sites

(Borchers and Perry 1992). Mycorrhizae and other

microbes affect soil structure by producing humic

compounds (Tan and others 1978), accelerating the

decomposition of primary minerals (Cromack and others

1979), and secreting organic "glues" (extracellular

polysaccharides) that bind soil particles into waterstable aggregates (Sutton and Sheppard 1976, Forster

1979, Tisdall and Oades 1979, 1982). The importance

of the first two processes to soil fertility and plant

nutrition is well documented. The third-aggregate

stability—also is important because the stability of

pores, essential for the movement of water and air

required by plant roots and soil organisms, is predicated

on the stability of aggregates formed when mineral

grains and homogeneous clays are linked by organic

glues (Lynch and Bragg 1985, Emerson and others

1986). When the flow of organic materials is

interrupted, soil aggregation-and ultimately forest sites

are degraded.

Examination of aggregates from granitic soils on a highelevation site in the Klamath Mountains of southwest

Oregon illustrates this decline. This site, logged in

1968, has not become reforested despite several

attempts (Amaranthus and Perry 1987). Now

dominated by grasses and herbs, it supports different

microflora than found associated with conifers.

Scanning electron micrographs show that soil structure

in the clearcut and adjacent, undisturbed forest differs

drastically: an abundance of relatively large (>50

microns) pores and fungal hyphae in the forest soil

contrast with a scarcity of large pores and fungal

hyphae in the clearcut soil. Degradation of soil

structure in clearcuts is likely to make reforestation

difficult, especially on potentially droughty sites. More

research is needed on how best to stabilize and

maintain soil structure after harvest.

Soil Moisture

Some soil organisms, especially mycorrhizal fungi,

seem to benefit associated plants when soil moisture is

limited (Mikola 1970). Mycorrhizae are important in

water uptake, storage, and transport (Reid 1979).

Mycorrhizal fungi substantially increase the surface

absorbing area of colonized roots, directly increasing

the total soil volume roots can explore for water. They

also prevent gaps between roots and soil, preserving

"liquid continuity" during drier periods. Differentiated

"vessel hyphae" can act as specialized channels for

water transport to the host plant (Read and others

1985).

FACTORS ADVERSELY AFFECTING

POPULATIONS OF BENEFICIAL

ORGANISMS

What reduces populations of beneficial organisms on

disturbed forest land? How do reduced populations

affect forest revegetation? Under what conditions are

forest sites most likely to be impaired? Final answers

are as yet unknown, but recent studies provide

sufficient evidence for reasonable speculation.

Much of the research on reductions in populations of

beneficial organisms after disturbance has focused on

ectomycorrhizae. Although information on other

organisms is scanty in the Pacific Northwest,

mycorrhizal fungus populations may serve as indicators

of the health and vigor of other associated beneficial

organisms. Mycorrhizae provide a biological substrate

for other microbial processes. For example, in a

Douglas-fir system, mycorrhizae provided 50% of the

annual biomass returned to the soil and 42% of the

annual nitrogen release (Fogel and Hunt 1983).

Microbial activity generally is stimulated in soil

surrounding ectomycorrhizae (Oswald and Ferchau

1963, Rambelli 1973)-an effect that likely extends

along the vast expanse of ectomycorrhizal fungal

mycelia in the soil. Numerous authors have reported

reductions in mycorrhiza populations due to forest

disturbance (Harvey and others 1980, Parke 1982,

Perry and others 1982, Amaranthus and others 1987 ).

However, the degree of reduction and its impact on

forest restoration vary widely and depend on many

factors.

Type and Severity of Disturbance

The most widespread activities which alter both the

aboveground and belowground environments and

which therefore potentially impact populations of soil

organisms are timber harvest and site preparation. In

the Pacific Northwest, clearcutting and prescribed

burning are the common harvesting and site-preparation

practices. Soil nutrient status, moisture, temperature,

pH, and organic matter content, litter inputs, and

species composition affect the growth and occurrence

of soil organisms (Harvey and others 1980)--and all of

these are influenced by harvesting and site preparation.

Clearing vegetation and disturbing the forest floor

remove nutrients and reallocate them within the

ecosystem. Harvesting host trees eliminates the

photosynthate source for dependent ectomycorrhizal

fungi and associated microbes. Converting a mature

forest to a clearcut typically increases soil temperatures

once the protective canopy is gone. Prescribed

broadcast burning increases soil pH, creates a nutrient

flush, and can reduce litter and duff levels (Amaranthus

and McNabb 1984). Soil organic matter, humified

material, and decaying wood are centers of microbial

182
activity and can substantially diminish as a result of

intense fire. Changes in aboveground community

composition alter the form of root exudates and litter

leachates.

Wright and Tarrant (1958) found fewer

ectomycorrhizae on Douglas-fir seedlings growing in

burned, compared to unbumed, clearcuts. The greatest

reductions were associated with the hottest burns.

Thus, not only the type of activity, but its severity, is

critical. Parke (1982) compared mycorrhiza formation

in soils from burned and unburned clearcuts of 36

"difficult to regenerate" sites in northwest California

and southwest Oregon. Douglas-fir and ponderosa pine

(P/'nus ponderosa Laws.) seedlings grown in soils from

the burned clearcuts formed 40% fewer

ectomycorrhizae, and seedlings grown in soils from the

unburned clearcuts 20% fewer ectomycorrhizae, than

seedlings grown in undisturbed forest soil. Yet it is

difficult to generalize about effects of burning on

microbial populations because they are highly

dependent on duration and intensity of fire as well as

soil and site conditions (Perry and Rose 1983).

Because ectomycorrhizae predominate in the organic

layers of the soil (Trappe and Fogel 1977, Harvey and

others 1979), the degree of organic matter lost from a

site can influence mycorrhiza populations. Harvey and

others (1979) found more than 87% of the active

ectomycorrhizal fungus types in humus and decaying

wood in a mature Douglas-fir/larch (Larix) forest.

Moreover, the physical consumption of humidified

material affects not only ectomycorrhizae but an array

of other beneficial organisms tied to site restoration.

Habitat for small mammals that are important in

distributing fungal spores of several belowground

mycorrhizal fungi (Maser and others 1978) also can be

lost. Thus, numbers, diversity, and activity of

beneficial soil organisms can be reduced by repeated

removal of organic matter from a site.

Minimizing soil compaction helps maintain healthy

populations of soil organisms by preserving soil

structure (Amaranthus and others 1996). Pore space is

essential for the movement of oxygen and water into

soil and the flushing of carbon dioxide out of it;

microbial activity is drastically altered when levels of

these basic elements become extreme. Undisturbed

forest soil is rarely saturated because large pores allow

for rapid downward percolation of water. However,

when soils are compacted, large pores are destroyed

and water movement through soil is reduced. Soil

compaction greatly influences the types and activities

of soil organisms sensitive to excess soil moisture.

Thus, the adverse response of plants to compaction

reflects both indirect effects on microbes and direct

effects on the plants themselves. Careful selection and

skillful implementation of harvest methods to minimize

compaction increase the potential for maintenance of

beneficial soil organisms.

Topsoil is a reservoir of both nutrients and fungal

spores and other propagules important for mycorrhiza

formation. Loss of this biological reservoir by erosion

will inevitably impair restoration. Reeves and others

(1979) found that the dominant species in a sagebrush

{Artemisia) community in Colorado all were

mycorrhizal. When topsoil was severely disturbed and

eroded, numbers of mycorrhizal propagules were

greatly decreased (Moorman and Reeves 1979), and

nonmycorrhizal weedy species could successfully

reestablish. Little is known about the effects of soil

erosion from deforested areas, but the density and

diversity of mycorrhizal inocula are reduced.

Climatic Conditions

Climate influences seedling growth and ectomycorrhiza

formation (Harvey and others 1980, Pilz and Perry

1984). The importance of early mycorrhiza formation

in dry areas has been emphasized (Parry 1953, Mikola

1970). Dry climates may limit the activity of

mycorrhizal fungi by decreasing the length of time for

spore production, germination, and optimal mycelial

growth, which in turn can decrease the chances for

planted seedlings to become colonized. Seedlings in

moist climates may be able to survive longer without

mycorrhizae than those in dry climates, increasing their

chances of becoming colonized. Fungal symbionts

differ in their capacity to function in periods of either

low or "excessive" soil moisture, and impaired function

will subsequently affect growth of associated plants

(Mexal and Reid 1973, Slankis 1974, Theodorou

1978). Moisture content also affects uptake of certain

nutrients by mycorrhizae (Gadgil 1972).

Seedlings growing in cold climates may also require

rapid, early mycorrhizal colonization to take advantage

of the short growing season and obtain the necessary

nutrients and water to survive the long cold season and

early frosts. In studies in the Klamath Mountains of

northwest California and southwest Oregon,

Amaranthus and Perry (1987) found that mycorrhiza

formation most strongly influences seedling survival

and growth on sites limited by both moisture and

temperature.

Biotic Conditions

What factors influence a seedling's potential to form

mycorrhizae on disturbed forest sites? Numerous

factors directly and indirectly influence the potential for

mycorrhiza formation, including seedling physiology,

environmental conditions, and abundance of soil

microorganisms and propagules (Perry and others

1987). Modification of these conditions may influence

a seedling's ability to form mycorrhizae. Mycorrhizal

183
potential following disturbance such as clearcutting and

slashburning depends primarily upon (1) the balance

between mortality and input of mycorrhiza-forming

propagules (such as spores and hyphae of mycorrhizal

fungi); (2) the recovery of host plants, which provides

the energy to stabilize populations of mycorrhizal fungi

and allows them to grow and colonize nonmycorrhizal

roots of surrounding plants; and (3) the diversity of

fungus species, which buffers the mycorrhizal

community against environmental changes following

disturbance.

Much remains to be learned about the importance of

aboveground species composition and arrangement to

belowground biological functioning. It is increasingly

apparent, however, that an ectomycorrhizal fungus can

link some plant species with fungal mycelia (Bjorkman

1970, Read and others 1985, Finlay and Read 1986).

In the natural forest environment, ectomycorrhizal fungi

supported by nonconiferous hosts can actively colonize

conifer seedlings. Root-chamber analysis of the

development of ectomycorrhizal mycelium has shown

that expanding hyphal fans not only act as nutrientabsorbing structures but also colonize nonmycorrhizal

feeder roots in host-plant combinations within and

among species. Using radioactive labeling, Finlay and

Read (1986) have demonstrated the free movement of

carbon among plants connected by mycorrhizal

mycelia. Clearly, the existence of "pipelines" for

distributing materials among plant species has

important implications for forest regeneration.

Little is known of the persistence and distribution of

ectomycorrhizae in the absence of living hosts. It has

been suggested (Hacskaylo 1973) that ectomycorrhizal

fungi do not persist long in the absence of hostsupplied substrates. In the Klamath Mountains of

northwest California and southwest Oregon, sites that

have been logged and burned are often rapidly invaded

by woody shrubs (Gratkowski 1961). Many of these

shrubs, members of the Ericaceae and Fagaceae, form

mycorrhizae with many of the same fungi as do

members of the Pinaceae (Molina and Trappe 1982).

This mechanism of natural redundancy preserves

mycorrhiza diversity during periods of rapidly changing

community structure.

Amaranthus and Perry (1989) transferred small

amounts of soil from a Pacific madrone {Arbutus

menziesii Pursh) stand to planting holes at three

locales-a site cleared of whiteleaf manzanita

(Arctostaphylos viscida Parry), a meadow cleared of

annual grasses, and under Oregon white oak (Quercus

garryana Hook) —on which Douglas-fir seedlings were

grown. In the first year, they found more rapid

mycorrhizal colonization of Douglas-fir feeder roots, and

dramatically improved seedling survival and growth, on

the site cleared of manzanita than in the meadow under

similar soil, moisture, and temperature conditions. In

the second year, Douglas-fir seedling performance

continued to increase dramatically on the site cleared

of manzanita compared to the other two locales. Rapid

mycorrhiza formation, especially on sites that are

difficult to regenerate, increases the chances that

succeeding species will become established.

Some woody shrub species may act as reservoirs not

only of mycorrhizal fungi but of other microflora as

well. Significantly higher rates of nitrogen fixation-and

increased seedling survival and growth-were found in

association with the roots of Douglas-fir seedlings in a

stand cleared of whiteleaf manzanita than in a meadow

cleared of annual grass (Amaranthus and others 1987).

Azospirillum, an N2-fixing bacterium, was isolated

within the mycorrhizae of Douglas-fir at the manzanita

site. Whiteleaf manzanita occupies particularly hot, dry

sites where fire is frequent. Because high N losses can

accompany intense fire, natural mechanisms by which

N is returned to the soil are important to site

restoration. Quite likely, the N fixed and gradually

accumulated in association with mycorrhizae is

biologically important for individual conifers over their

long life.

We suggest that conifers and some woody shrubs form

successional guilds, in which the rapidly sprouting

pioneer members of the guild maintain the guild's

shared soil microflora, forming a biological time bridge

between old and new conifer stands. In the long run,

maintaining diversity of soil microflora by pioneering

shrubs may optimize conifer performance during the

varied conditions likely over a rotation.

Effect of Mycorrhizal Hosts Shifts Over Time

Many grass species and shrub species such as

Ceanothus and Rubus (e.g., blackberry, salmonberry)

form vesicular-arbuscular mycorrhizae (VAM) (Rose and

Youngberg 1981, Trappe unpublished data). On sites

long dominated by VAM species, space availability and

the ectomycorrhizal fungi needed by members of the

Pinaceae may gradually disappear, and the soil

microbial complex associated with ectomycorrhizae will

shrivel. Occurrence of VAM in some members of the

Pinaceae has been reported, however, the ecological

significance remains undetermined (Ccizares and Trappe

1993, Ccizares and Smith 1996). Invasion of sites by

nonectomycorrhizal plants over years can seriously

affect reforestation (Amaranthus and Perry 1987) and

yield, particularly in the case of ectomycorrhizal tree

species growing on difficult sites where seedlings must

establish ectomycorrhizae early to survive. Changes in

soil microflora associated with nonhost species could

affect patterns of decomposition; nutrient and water

retention, availability, and uptake; and the balance

between pathogens and pathogen-inhibiting organisms.

184
How long soils retain their mycorrhizal colonization

potential in the absence of living hosts is unknown.

Ectomycorrhizal spores and hyphal fragments have

remained metabolically active after 2 years in

Scandinavian forests, though the number of active

fragments dropped dramatically over that period

(Persson 1982, Ferrier and Alexander 1985). In the

Pacific Northwest, mycorrhiza formation generally

decreases as the length of time between disturbance

and reforestation increases (Pilz and Perry 1984, Perry

and others 1987).

POTENTIAL FOR MANAGING SOIL

ORGANISMS ON DEGRADED SITES

Survival and growth of tree seedlings inoculated with

the ectomycorrhizal fungus Pisolithus tinctorius have

been enhanced in some areas (Marx 1975, Marx and

Krupa 1978). However, the generally disappointing

performance of P. tinctorius in the Pacific Northwest

(Alvarez 1982; M. Castellano and J.M. Trappe,

unpublished data, USDA Forest Service, Pacific

Northwest Research Station) suggests that there is no

"magic bullet." Because of the wide variety of soil,

climatic, and biotic conditions characterizing temperate

forests, it is improbable that a single fungus could

benefit all host species and adapt to all conditions. For

example, the types and activities of soil organisms

associated with young conifers may be quite different

from those associated with mature forests (Mason and

others 1983, Amaranthus and Perry 1987). On an old,

high-elevation, nonreforested clearcut, soil inoculum

transferred from a plantation significantly increased

survival and growth of Douglas-fir seedlings compared

to soil inoculum transferred from adjacent, mature

forest. Likewise, soil organisms needed to help

seedlings establish on difficult sites may differ from

those which sustain site productivity over a forest

rotation.

In the Pacific Northwest, Douglas-fir seedlings growing

in containers in nurseries have been successfully

inoculated with the mycorrhizal fungus Rhizopogon

vinicolor when its spores were added to irrigation water

(M. Castellano, unpublished data, 1986). Although not

all fungal symbionts can be inoculated in this manner,

success was attained with this promising species. R.

vinicolor is commonly found in clearcuts on young

Douglas-fir seedlings and can improve seedling

tolerance to drought (Parke and others 1983).

Preliminary results upon outplanting indicate that

seedlings inoculated with R. vinicolor also grow better

(Castellano and others 1985).

Inoculating seedlings with beneficial soil organisms in

bareroot nurseries has similar potential for improved

outplanting performance. However, mycorrhizae

adapted to outplanting sites are often lacking in modern

bareroot nurseries (Trappe 1977). High fertility levels

and routine fumigation, which are common in nurseries,

reduce mycorrhizal fungus populations. Methods to

increase mycorrhiza formation and diversity on

seedlings in the nursery need further attention.

Selecting the appropriate organisms is also critical.

Seedlings colonized by a diversity of soil organisms

adapted to conditions in the outplanting environment

may outperform seedlings colonized by less diverse

populations adapted to nursery conditions (Trappe

1977).

Much remains to be learned about the potential

benefits of reintroducing soil organisms that benefit

plants into degraded soils. Some, perhaps many,

restoration projects will be aided by this approach and

the evidence for this optimism is strong. However,

even where it does benefit restoration, reintroducing

beneficial organisms is not some kind of magic bullet

that will allow the rest of the ecosystem to be ignored.

The best advice for those working in reclamation is to

try. One does not have to be a scientist, or even have

a high school diploma, in order to experiment. All it

takes is common sense —perhaps backed up by a little

intuition and access to the proper soil organisms.

Forest managers should be aware that the restorationenhancing attributes of soil organisms are the product

of diverse, complex interactions within natural systems

that have co-evolved over millenia. Thus, the best

management approach is to minimize drastic or

cumulative impacts to the forest environment so as to

retain populations of beneficial organisms. To do this,

forest managers should:

1. Minimize disturbance severity-that is, intense

burns, extensive soil compaction or erosion.

2. Emphasize retention of site organic matter.

3. Emphasize rapid revegetation of sites by

indigenous host species and associated beneficial

soil organisms.

4. Recognize that sites with harsh environments such

as cold or drought are most susceptible to

restoration losses.

A new era is emerging in our understanding of forest

ecosystems. Incorporating the concept of the "living

soil" in our evaluation of site restoration is part of this

new, expanding view. Although much remains to be

learned, we are making progress at a rapid rate. By

understanding soil organisms, practicing foresters can

better achieve short-term management goals while

ensuring site restoration and recovery. The challenge

to forest managers is to assimilate the growing body of

185
information concerning soil organisms in forest

ecosystems and put that information to wise use.

LITERATURE CITED

Alexander, I.J.; Hardy, K. 1981. Surface phosphatase

activity of Sitka spruce mycorrhizas from a

serpentine site. Soil Biology and Biochemistry

13:301-305.

Allen, M.F. 1991. The ecology of mycorrhizae.

Cambridge Univ. Press, Cambridge.

Alvarez, I. 1982. Nursery, cold storage, and field

studies on western conifers inoculated with spores

of Pisoiithus tinctorius. Ph.D. thesis. Corvallis,

Ore.: Oregon State University.

Amaranthus, M.P.; McNabb, D.H. 1984. "Bare soil

exposure following logging and prescribed burning

in southwest Oregon." In New forests for a

changing world, 234-237. Proceedings, 1983

National Convention, Portland, Ore. Washington,

D.C.: Society of American Foresters.

Amaranthus, M.P.; Perry, D.A. 1987. Effect of soil

transfer on ectomycorrhiza formation and the

survival and growth of conifer seedlings on old,

nonreforested clear-cuts. Canadian Journal of

Forest Research 17(8):944-950.

Amaranthus, M.P.; Perry, D.A.; Borchers, S.L. 1987.

"Reduction of native mycorrhizae reduce growth of

Douglas-fir seedlings." In Proceedings of the North

American conference on mycorrhizae, Sylvia, D.M.;

Graham, J.H. (eds.). 79. Gainesville, Fla.:

University of Florida.

Amaranthus M.P.; Parrish, D; Perry, D.A. 1989.

Decaying logs as moisture reservoirs following

drought and wildfire. |n Stewardship of Soil, Air

and Water Resources. Alexander, E., (ed.).

Proceedings Watershed 89, Juneau, AK. USDA

Forest Service, Region 10. R10-MB-77. pp 191-

194.

Amaranthus, M.P.; Perry, D.A. 1989. Interaction

effects of vegetation type and Pacific madrone soil

inocula on survival, growth and mycorrhiza

formation of Douglas-fir. Canadian Journal of

Forest Resources 19:550-556.

Amaranthus, M.P.; Li, C.Y.; Perry, D.A. 1987.

Nitrogen fixation within mycorrhizae in the next

decade: practical applications and research

priorities. Sylvia, D.M.; Hung, L.L.; Graham, J.H.

(eds.). University of Florida, Gainesville, p. 79.

Amaranthus, M.P.; Li, C.Y.; Perry, D.A. 1990.

Influence of vegetation type and madrone soil on

associative nitrogen fixation in Douglas-fir

rhizospheres. Canadian Journal of Forest Research

20(3):368-371.

Amaranthus, M.P.; Page-Dumroese, D.; Harvey, A.;

Cazares, E.; Bednar, L.F. 1996. Soil compaction

and organic matter affect conifer seedling nonmycorrhizal and ectomycorrhizal root tip abundance

and diversity. Res. Pap. PNW-RP-494. Portland,

OR:U.S. Department of Agriculture, Forest Service,

PNW Research Station. 12 p.

Arnason, T.; Lambert, J.D.H.; Gale, J. 1982. Decline

of soil fertility due to intensification of land use by

shifting agriculturist in Belize, Central America.

Agroecosystems 8:27-37.

Borchers, J.G.; Perry, D.A. 1992. The influence of

soil texture and aggregation on carbon and

nitrogen dynamics in southwest Oregon forests

and clearcuts. Canadian Journal of Forest

Research 22:298-305.

Bjorkman, E. 1970. Forest tree mycorrhizae-the

conditions for formation and the significance for

growth and afforestation. Plant and Soil 32:589-

610.

Bollen, W.B. 1974. "Soil microbes." In Environmental

effects of forest residues management in the

Pacific Northwest, edited by O. Cramer, B1-B41.

General Technical Report PNW-24. Portland, Ore.:

Pacific Northwest Forest and Range Experiment

Station, USDA Forest Service.

Castellano, M.A.; Trappe, J.M.; Molina, R. 1985.

Inoculation of container-grown Douglas-fir

seedlings with basidiospores of Rhizopogon

vinicolor and R. colossus: effects of fertility and

spore application rate. Canadian Journal of Forest

Research 15:10-13.

Cazares, E.; Smith, J.E. 1996. Occurrence of

vesicular-arbuscular mycorrhizae in Pseudotsuga

menziesii and Tsuga heterophylla seedlings grown

in Oregon Coast Range soils. Mycorrhiza 6: 65-67

Cazares, E.; Trappe, J.M. 1993. Vesicular endophytes

in roots of the Pinaceae. Mycorrhiza 2: 153-156.

Cromack, K.; Sollins, P.; Graustein, W.C.; Seidel, K.;

Todd, A.W.; Spycher, G.; Ching, Y.L.; Todd, R.L.

1979. Calcium oxalate accumulation and soil

weathering in mats of the hypogeous fungus

Hysterangium crassum. Soil Biology and

Biochemistry 11:463-468.

186
Dindal, D.L. (editor). 1973. Proceedings of the first

soil micro-communities conference. A.E.C. Tech.

Inf. Ant. CONF-711076.

Domingo, I.L. 1983. "Nitrogen fixation in southeast

Asian forestry: research and practice." In

Biological nitrogen fixation in forest ecosystems:

foundations and applications, Gordon, J.C.;

Wheeler, C.T. (eds). 295-315. The Hague, The

Netherlands: Martinus Nijhoff/Dr W. Junk

Publishers.

Emerson, W.W.; Foster, R.C.; Oades, J.M. 1986.

"Organo-mineral complexes in relation to soil

aggregation and structure." In Interactions of soil

minerals with natural organics and microbes,

Huangm, P.M.; Schnitzer, M. (eds.). 521-548.

Special Publication 17. Madison, Wis.: Soil

Science Society of America.

Ferrier, R.C.; Alexander, I.J. 1985. "Persistence under

field conditions of excised fine roots and

mycorrhizas of spruce." In Ecological interactions

in soil. Fitter, A.H.; Atkinson, D.; Read, D.J.;

Usher, M.B. (eds.). 175-179. Oxford: Blackwell

Scientific Publications.

Finlay, R.D.; Read, D.J. 1986. Translocation of CO2-

labelled carbon between plants interconnected by a

common mycelium. New Phytologist 103:143-

156.

Fogel, R.; Hunt, G. 1983. Contribution of mycorrhizae

and soil fungi to nutrient cycling in a Douglas-fir

ecosystem. Canadian Journal of Forest Research

13:219-232.

Forster, S.M. 1979. Microbial aggregation of sand in

an embryo dune system. Soil Biology and

Biochemistry 11:537-541.

Foster, R.C.; Marks, G.C 1967. Observations of the

mycorrhizas of forest trees. II. The rhizosphere of

Pinus radiata D. Don. Australian Journal of

Biological Sciences 20:915-926.

Gadgil, P.D. 1972. Effect of waterlogging on

mycorrhizas of radiata pine and Douglas-fir. New

Zealand Journal of Forest Science 2:222-226.

Giller, K.E.; Day, J.M. 1985. "Nitrogen fixation in the

rhizosphere: significance in natural and agricultural

systems." In Ecological interactions in soil. Fitter,

A.H.; Atkinson, D.; Read, D.J.; Usher, M.B. (eds.).

127-147. Oxford: Blackwell Scientific

Publications.

Gordon, J.C.; Avery, M.E. 1985. "Improving tree

crops using microorganisms in designed systems."

In Attributes of trees as crop plants. Cannell,

M.G.R.; Jackson, J.E. (eds.). 316-326. Abbots

Ripton, Hunts, England: Institute of Terrestrial

Ecology.

Gratkowski, H. 1961. Brush problems in

southwestern Oregon. Portland, Ore.: Pacific

Northwest Forest and Range Experiment Station,

USDA Forest Service.

Graustein, W.; Cromack, K. Jr.; Sollins, P. 1977.

Calcium oxalate: occurrence in soils and effect on

nutrient and geochemical cycles. Science

198:1252-1254.

Hacskaylo, E. 1973. "Carbohydrate physiology of

ectomycorrhizae." In Ectomycorrhizae: their

ecology and physiology, edited by Marks, G.C;

Kozlowski, T.T., 207-230. London: Academic

Press.

Harley, J.L.; Smith, S.E. 1983. Mycorrhizal

symbioses. London: Academic Press.

Harvey, A.E.; Larsen, M.J.; Jurgensen, M.F. 1979.

Comparative distribution of ectomycorrhizae in

soils of three western Montana forest habitat

types. Forest Science 25:350-360.

Harvey, A.E.; Larsen, M.J.; Jurgensen, M.F. 1980.

Clearcut harvesting and ectomycorrhizae: survival

of activity on residual roots and influence on a

bordering forest stand in western Montana.

Canadian Journal of Forest Research 10:300-303.

Harvey, A.E.; Larsen, M.J.; Jurgensen, M.F.; Schlieter,

J.A. 1986. Distribution of active ectomycorrhizal

short roots in forest soils of the inland Northwest:

effects of site and disturbance. General Technical

Report INT-374. Ogden, Utah: Intermountain

Research Station, USDA Forest Service.

Ho, I.; Zak, B. 1979. Acid phosphatase activity of six

ectomycorrhizal fungi. Canadian Journal of Botany

57:1203-1205.

Kelly, W.D. 1976. Evaluation of Trichoderma

harzianum integrated clay granules as a biocontrol

for Phytophthora cinnamomi causing damping off

of pine seedlings. Phytopathology 66:1023-1027.

Larsen, M.J.; Harvey, A.E.; Jurgensen,M.F. 1980.

"Residue decay processes and associated

environmental functions in northern Rocky

Mountain forests." In Environmental consequences

of timber harvesting, 157-174. General Technical

Report INT-90. Ogden, Utah: Intermountain

Research Station, USDA Forest Service.

187
Lynch, J.M.; Bragg, E. 1985. Microorganisms and soil

aggregate stability. Advances in Soil Science

2:133-171.

Marx, D.H. 1972. Ectomycorrhizae as biological

deterrents to pathogenic root infections. Annual

Review of Phytopathology 10:429-454.

Marx, D.H. 1973. Growth of ectomycorrhizal and

nonmycorrhizal shortleaf pine seedlings in soil with

Phytophthora cinnamomi. Phytopathology 63:18-

23.

Marx, D.H. 1975. Mycorrhizae and establishment of

trees on strip-mined land. Ohio Journal of Science

75:288-297.

Marx, D.H.; Krupa, S.V. 1978. "Ectomycorrhizae." In

Interactions between non-pathogenic soil

microorganisms and plants. Dommergues, Y.R.;

Krupa, S.V. (eds.). 373-400. Amsterdam: Elsevier

Scientific Publishing Company.

Maser, C; Trappe, J.M.; Nussbaum, R. 1978. Fungalsmall mammal interrelationships with emphasis on

Oregon coniferous forests. Ecology 59:799-809.

Mason, P.A.; Wilson, J; Last, F.T.; Walker, C. 1983.

The concept of succession in relationship to the

spread of sheathing mycorrhizal fungi in inoculated

tree seedlings growing in unsterile soil. Plant and

Soil 71:247-256.

Mexal, J.; Reid, C.P.P. 1973. The growth of selected

mycorrhizal fungi in response to induced water

stress. Canadian Journal of Botany 51:1579-

1588.

Mikola, P. 1970. Mycorrhizal inoculation in

afforestation. International Review of Forestry

Research 3:123-196.

Molina, R.; Trappe, J.M. 1982. Lack of mycorrhizal

specificity by the ericaceous hosts Arbutus

menziesii and Arctostaphylos uva-ursi. New

Phytologist 90:495-509.

Moorman, T.; Reeves, F.B. 1979. The role of

endomycorrhizae in revegetation practices in the

semi-arid West. II. A bioassay to determine the

effect of land disturbance on endomycorrhizal

populations. American Journal of Botany 66:14-

18.

Parke, J.L. 1982. Factors affecting the inoculum

potential of VA and ectomycorrhizal fungi in forest

soils of southwest Oregon and northern California.

Ph.D. thesis. Corvallis, Ore.: Oregon State

University.

Parke, J.L.; Linderman, R.G.; Black, C.H. 1983. The

role of ectomycorrhizas in drought tolerance of

Douglas-fir seedlings. New Phytologist 95:83-95.

Parry, M.S. 1953. Tree planting in Tanganyika:

methods of planting. East African Agricultural

Journal 18:102-115.

Perry, D.A.; Meyer, M.M.; Egeland, D.; Rose, S.L.; Pilz,

D. 1982. Seedling growth and mycorrhizal

formation in clearcut and adjacent undisturbed soils

in Montana: a greenhouse bioassay. Forest

Ecology and Management 4:261-273.

Perry, D.A.; Molina, R.; Amaranthus, M.P. 1987.

Mycorrhizae, mycorrhizospheres, and reforestation:

current knowledge and research needs. Canadian

Journal of Forest Research 17(8):929-940.

Perry, D.A.; Rose, S.L. 1983. "Soil biology and forest

productivity: opportunities and constraints." In

IUFRO symposium on forest site and continuous

productivity. Ballard, R.; Gessel, S.P. (eds.). 229-

238. General Technical Report PNW-163.

Portland, Ore.: Pacific Northwest Forest and Range

Experiment Station, USDA Forest Service.

Perry, D.A.; Rose, S.L. 1987. "The influence of site

preparation on forest productivity." In

Proceedings, California conferences on forest

nutrition. Power, R.; Robson, T. (eds.). Berkeley,

Calif.: Pacific Southwest Forest and Range

Experiment Station, USDA Forest Service.

Perry, D.A.; Amaranthus, M.P.; Borchers, J.G.;

Borchers, S.L.; Brainerd, R.E. 1989.

Bootstrapping in ecosystems. BioSci. 39:230-237.

Persson, H. 1982. Changes in the tree and dwarf

shrub fine-roots after clearcutting in a mature

Scots pine stand. Swedish Coniferous Forest

Project, Technical Report 31.

Pilz, D.P.; Perry, D.A. 1984. Impact of clearcutting

and slashburning on ectomycorrhizal associations

of Douglas-fir. Canadian Journal of Forest

Research 14:94-100.

Oswald, E.T.; Ferchau, H.A. 1963. Bacterial

associations of coniferous mycorrhizae. Plant and

Soil 28:167.

Powell, P.E.; Cline, G.R.; Reid, C.P.P.; Szaniszlo, P.J.

1980. Occurrence of hydroxamate siderophore

iron chelators in soils. Nature 287:833-834.

188
Rambelli, A. 1973. "The rhizosphere of mycorrhizae."

In Ectomycorrhizae: their ecology and physiology.

Marks, G.C.; Kozlowski, T.T. (eds.). 229-249.

London: Academic Press.

Read, D.J.; Francis, R.; Finlay, R.D. 1985.

"Mycorrhizal mycelia and nutrient cycling in plant

communities." In Ecological interactions in soil.

Fitter, A.H.; Atkinson, D.; Read, D.J.; Usher, M.B.

(eds.). 193-218. Oxford: Blackwell Scientific

Publications.

Reeves, F.B.; Wagner, R.D.; Moorman, T.; Kiel, J.

1979. The role of endomycorrhizae in revegetation

practices in the semi-arid West. I. A comparison of

incidence of mycorrhizae in severely disturbed

natural environments. American Journal of Botany

66:6-13.

Reid, C.P.P. 1979. "Mycorrhizae and water stress."

In Root physiology and symbiosis, edited by A.

Riedacker and M.J. Gagnaire-Michard, 392-408.

Nancy, France: IUFRO Symposium.

Rose, S.L.; Li, C.Y.; Hutchins, A.S. 1980. A

streptomycete antagonist to Phellinus weirii, Fomes

annosus, and Phytophthora cinnamomi. Canadian

Journal of Microbiology 26:583-587.

Rose, S.L.; Youngberg, C. 1981. Tripartite

associations in snowbrush (Ceanothus velutinus):

effect of vesicular-arbuscular mycorrhizae on

growth, nodulation, and nitrogen fixation.

Canadian Journal of Botany 59:34-39.

Sharma, G.D. 1983. Influence of jhumming on the

structure and function of microorganisms in a

forested ecosystem. Hill Geography 2:1-11.

Slankis, V. 1974. Soil factors influencing formation of

mycorrhizae. Annual Review of Phytopathology

12:437-457.

Sutton, J.C.; Sheppard, B.R. 1976. Aggregation of

sand dune soil by endomycorrhizal fungi. Canadian

Journal of Botany 54:326-333.

Tan, K.H.; Sihanonth, P.; Todd, R.L. 1978. Formation

of humic acid like compounds by the

ectomycorrhizal fungus Pisolithus tinctorius. Soil

Science Society of America Journal 42:906-908.

Tarrant, R.F.; Trappe, J.M. 1971. The role of alder in

improving the forest environment. Plant and Soil,

Special Volume: 335-348.

Theodorou, C. 1978. Soil moisture and the

mycorrhizal association of Pinus radiata D. Don.

Soil Biology and Biochemistry 10:33-37.

Tisdall, J.M.; Oades, J.M. 1979. Stabilization of soil

aggregates by the root systems of ryegrass.

Australian Journal of Soil Research 17:429-441.

Tisdall, J.M.; Oades, J.M. 1982. Organic matter and

water stable aggregates in soils. Journal of Soil

Science 33:141-164.

Trappe, J.M. 1977. Selection of fungi for

ectomycorrhizal inoculation in nurseries. Annual

Review of Phytopathology 15:203-222.

Trappe, J.M.; Bollen, W.B. 1979. "Forest soils biology

of the Douglas-fir region." In Forest soils of the

Douglas-fir region, edited by Heilman, P.E.;

Anderson, H.W. Pullman, Wash.: Washington

State University.

Trappe, J.M.; Fogel, R.D. 1977. "Ecosystematic

functions of mycorrhizae." In The belowground

ecosystem, 205-214. Fort Collins, Colo.: Colorado

State University, Range Science Department

Scientific Series.

Trappe, J.M.; Franklin, J.F; Tarrant, R.F.; Hansen, G.M.

(editors). 1968. Biology of alder. Proceedings of

a symposium. Northwest Science Association.

Portland, Ore.: Pacific Northwest Forest and Range

Experiment Station, USDA Forest Service.

Williamson, B.; Alexander, I.J. 1975. Acid

phosphatase localized in the sheath of beech

mycorrhiza. Soil Biology and Biochemistry 7:195-

198.

Wingfield, E.B. 1968. Mycotrophy in loblolly pine.

Ph.D. thesis. Blacksburg, Va.: Virginia Polytechnic

Institute and State University.

Wollum, A.G., II; Youngberg, C.T. 1964. The

influence of nitrogen fixation by non-leguminous

woody plants on the growth of pine seedlings.

Journal of Forestry 62:316-321.

Wright, E.; Tarrant, R.F. 1958. Occurrence of

mycorrhizae after logging and slash burning in the

Douglas-fir forest type. Research Note PNW-160.

Portland, Ore.: Pacific Northwest Forest and Range

Experiment Station, USDA Forest Service.

Zak, B. 1964. Role of mycorrhizae in root disease.

Annual Review of Phytopathology 2:377-392.

189
SOIL ORGANISMS-APPLICATIONS FOR NATIONAL FOREST MANAGEMENT

Karen Bennett1

ABSTRACT

Maintaining health and vigor of ecosystems requires definition of thresholds. Soil quality standards, developed by the

US Forest Service Regions, focus on limiting area of impact, protecting physical soil properties, and maintaining some

level of organic matter. Soil compaction and puddling, displacement and erosion and severe burning are specific

indicators of detrimental conditions. Monitoring results show that many practices can meet soil quality standards, but

repeated entries, with ground based equipment, often don't meet standards. Soil microorganisms may be good early

warning indicators of declining conditions. Better understanding of below ground communities is important for

ecosystem sustainability. Collaboration between the scientific community and managers needs to improve.

Keywords: soil quality standards, compaction, displacement, severe burning, monitoring, ecosystem sustainability

The controversies we face in the management of the

National Forests is not new. Preservation versus

utilization of our nations resources was debated in the

halls of congress and throughout society since before

the authorization of the National Forest System in the

late 1800's. It will continue to be a topic of discussion

for centuries to come. Miners, ranchers, loggers,

environmentalists, scientists, economists and society

as a whole, all see a different value in and from the

forest. Everyone wants a piece of the resources, some

want it all, and all are motivated by differing needs and

desires.

We are no longer an isolated nation in these debates.

It is now recognized that our land management

decisions not only have national implications, but affect

global resources and global economies. There has not

been a greater declaration of concern for global forest

resources than at the Environmental Summit in Rio.

From that summit emerges these criteria:

1. Maintain the productive capacity of forest lands

2 Conserve biological diversity

3. Maintain the forest contribution to the global

carbon cycle

4. Conserve and maintain soil and water resources

5. Maintain ecosystem health and vigor.

The National Forest System and especially forests in

the Pacific Northwest, being among the most

productive ecosystems in the world, have a

responsibility to uphold these environmental criteria.

Recent forest planning efforts including the Northwest

Forest Plan and the Interior Columbia Basin Ecosystem

Management Project capture these concepts and the

future of forest conservation depends on their

implementation at a global scale.

No one who depends on the utilization of renewable

resources can deny the ultimate importance of these

guidelines. Ranchers want a productive forage base,

forest industries want a sustainable timber base,

everyone wants clean water and a healthy

environment. Can we have it all? Can we utilize

products from the forest while maintaining biodiversity

and protecting soil and water resources? At what level

do we maintain the health and vigor of ecosystems?

At what spatial and temporal scales? This is where the

debates begin This is where the agreement

erodes This is where the thresholds need to be

defined.

Our ability as humans to establish thresholds of

ecosystem sustainability has a rather poor track record.

Eric Ekholms book "Losing Ground" provides a critical

perspective on this issue. Remember the Cedars of

Lebanon? The great timbers that were utilized by

royalty for centuries for their castles and their ships?

Now in their place, a rocky shallow soil - eroded by

centuries of deforestation and overgrazing supports

only forb and shrub species. No one noticed the

decline!

The health and resilience of the forests in parts of

Europe today is drastically declining. Disease and poor

growth of trees is the norm after centuries of timber

harvest and the removal of forest litter for fuel and

farm compost. Nutrient cycling processes were

truncated. No one noticed the decline!

1

 Karen Bennett is Soil Scientist and Watershed Program Manager, USDA Forest Service, Siuslaw National Forest, Box 1148,

Corvallis, OR 97339.

190
Our counterparts in Australia, where a 30 year rotation

produces their commercial timber crop, have proven

that soil compaction and loss of organic materials

through silvicultural practices has resulted in a decline

of a whole site class in productivity in just one rotation.

No one noticed the decline!

How much degradation is acceptable to fulfill societal

needs? Where do we define our thresholds?

Many ecosystem processes depend upon the condition

of the soil resource. It was recognized shortly after

passage of the National Forest Management Act that

we as managers, needed appropriate, implementable

standards with which to protect soil resources. As a

result, we have soil quality standards in place.

Although each Region differs slightly in the specifics of

the standards, the concepts are generally the same:

- limit the amount of area that is impacted,

- protect soil physical properties, and

- maintain some level of organic matter on or in the

soil surface.

These standards were designed with at least two

specific intentions. They had to be appropriate and

they had to be implementable. Appropriate means they

have to be true indicators of detrimental soil conditions.

As such, they represent the best scientific

understanding of the factors that effect the

sustainability of the soil resource that we have today.

Implementable means that someone has to be able to

measure and detect the resource condition without

extensive and expensive procedures. This is also the

reason why we focus on changes to soil physical

properties, since they can be easily measured and infer

that these changes have implications for soil chemical

and biological properties as well. Not only do we have

a clear set of standards in place, we also have

standardized monitoring techniques that insure valid

conclusions are drawn from monitoring efforts.

Look at the indicators:

Severe burning means that the temperature

extremes and duration of fire on an area had to be

significant enough to literally oxidize the soil, result

in red soil colors and drive humic acids down into

the soil profile.

Compaction and puddling deal with changes in soil

structural properties. The standard is a 15 or 20%

increase in bulk density. There is currently debate

about the adequacy of this measurement versus soil

resistance to penetration. However, the point is

that the strength of the soil, or the amount of pore

space, is changed to a point where growth of above

ground plant species is negatively affected.

Displacement/Erosion allows the loss of 50% of the

A horizon of soils.

As defined, these indicators measure degraded soil

conditions. By the time these levels of soil

disturbance can be detected, there has truly been a

long-term degradation of the soil resource. These are

not conservative standards. It is the spatial limitation

to 15% of an activity area that make them acceptable

from an ecosystem sustainability view point.

These Soil Quality Standards establish our thresholds!

We are legally required to implement these standards to

fulfill the intent of long-term soil protection in the

National Forest Management Act. How are we doing?

The verdict is mixed. Where skyline logging or aerial

logging systems, small landings and a restricted road

networks are employed, our monitoring reports show

that we generally are able to meet the standards, the

most common soil disturbance has been associated

with hot slash burns and subsequent dry-ravel erosion.

But in areas where ground based logging systems are

employed, especially if this is the second or third entry

onto that same piece of ground, monitoring has shown

that we cannot consistently meet our soil standards.

Compaction and displacement of the surface soil

horizon are typically the dominant soil disturbing

factors. Recent Forest monitoring reports show that

close to 60% of the units logged by ground based

equipment exceeded Regional standards.

Why can't we implement our standards? The variety of

resource issues that get juggled in each environmental

analysis are numerous and complex. Most resource

managers focus on the signatures of forest or

rangeland health that they see expressed in the above

ground systems. As soil scientists, we are trained to

see the minute, to see small changes in landscapes, on

soil surfaces and in the soil profile. We look for these

changes because those changes help us to define soil

characteristics. We focus on these changes whether

it's a change in landscape position, in depth of the litter

layer, or in deformation of soil structure, to help us

interpret soil resource conditions. Others do not see

these subtle changes, they are not trained to look for

these changes. They can see dying trees, they can see

wildfire, they can see results from harvesting trees,

from piling and from burning slash. It is the condition

of the above ground environment that, in their minds,

either positively or negatively affects the ecosystem for

which they have management responsibility. They

cannot see the effects on the below ground ecosystem.

Therefore, they question the validity of the soil quality

standards when balanced against the management of

other resource conditions which are readily apparent.

As a result, there is no consistent assurance that soil

quality standards will be implemented. It is difficult to

be concerned about a system that you cannot see. It is

easier to believe that every soil is resilient and that it

will recover on its own. The need for us to be

191
educators is well defined or we will all be waiting for

the decline in ecosystem productivity.

We must become communicators, we must act as

educators. Our responsibilities as soil resource

managers, are to interpret the implications of degraded

soil systems on ecosystem productivity so that wise

management choices can be made. It is not an easy

task, it is not the role which the majority of us chose

as a career path. But unless we effectively

communicate and educate, the message will continue

to go unheard.

Being the educator implies many things, it implies that

we are educated ourselves, that we know the systems

in which we are working. We must be able to stratify

landscapes and define driving processes. We need to

know inherent constraints and how they differ for

different environments. We must discover important

ecological links. We must define and highlight sensitive

areas, separate out the resilient areas. We cannot do

all that alone, soil scientists understand certain parts of

the ecosystem, other professionals - geologists,

hydrologists, ecologists, foresters... understand

different parts. Pool the knowledge to develop this

understanding. Working together to understand a

landscape is an education in itself, take time to do it.!

The current revision of the soil quality standards adds

some guidelines which direct each Forest or ecoregion

to establish local standards of appropriate levels of

coarse and fine woody debris for various spatial and

temporal scales. Step forward, lead those efforts,

investigate the systems you work in and develop some

common understandings with your peers and

managers.

In addition, collaboration with the scientific community

must continue to improve. Together we can focus on

appropriate forest management questions and initiate

some sound management advice. Apart we may

develop separate conclusions. As resource specialists

we constantly have to disclose environmental effects

of our management activities. We've heard the experts

at this conference hesitate to answer the same

questions we are asked to answer every day. We can't

wait for results from a 200 year experiment to help us

understand the intricacies of the ecological components

we are manipulating everyday. As a community we

must move cautiously ahead, insuring that we keep the

critical parts of the ecosystem intact.

Scientists and forest managers need to work together

to assess our current soil quality standards which

speculate that if physical soil properties are

protected,the habitat for below ground communities

will thrive. If our mission is to sustain ecosystem

productivity, do we need to wait for physical indicators

to tell us our systems are degraded? Soil microbial

populations quickly alter based on current

environmental conditions. Microbes are adept at

manifesting "survival of the fittest"! - whichever

organisms are best adapted to a given environmental

condition flourish, others decline or are eliminated from

the system. Our ability to utilize microorganisms as

early indicators, as a warning system for environmental

degradation may help us to alter management practices

or employ some restoration techniques before we are

able to notice the decline through above or below

ground indicators. If soil microbial parameters are

sensitive biological markers that can be used to assess

changes in soil quality, then we should be using them.

We know that below ground communities are important

to ecosystem sustainability, they are vital. We do not

understand the intricacies of below ground systems,

we may never understand. That should inspire us to

ask questions, to explore, to investigate. Think of how

our management decisions would change if we could

identify subtle changes in microbial communities that

forewarn us of unacceptable changes in the

environment. The Forest Service and Bureau of Land

Management manage an incredible laboratory, quite a

diverse laboratory. A laboratory from which a

multitude of data could be assembled. The task

remains for us to build a library of information, collect

soil resource information and be able to interpret the

results for different ecological units.

Before we embark on such a venture, we need to

determine objectives for data collection and how to

meet those objectives. We need to determine the soil

physical, biological and chemical parameters that may

help us answer ecosystem sustainability questions.

Then we would need to design and organized a method

to fund and implement the data acquisition.

Embarking on such a venture would be a long-range

strategic process, possibly tied in with our ecological

unit inventories. Collaboration is the key, discovery is

the impetus. If we don't strive to get more insights

into these below-ground systems, we will we be asking

these same questions 70 years from now with no more

insight than we have today.

We need to start thinking about how below ground

communities are affected by our management

activities. We know that these complex and diverse

communities are susceptible to specific stimulations

and to specific stresses. We need to be able to

disclose this information in our activity assessments

and we need to be able to understand the implications

of changes in these communities as we look toward

restoring our ecosystems so that we will notice the

decline!

192
SOIL ORGANISMS-APPLICATIONS TO BUREAU OF LAND MANAGEMENT RANGELAND

MANAGEMENT

Mike Pellant1

ABSTRACT

Livestock grazing management on public lands administered by the Bureau of Land Management (BLM) has evolved

from a commodity oriented approach to an ecological approach that still includes sustainability of uses such as livestock

grazing. The role of the soil organisms in rangeland management has received little agency attention until recently. The

implementation of Standards for Rangeland Health in 1 997 has focused management attention on soil organisms

relative to their roles in nutrient cycling, energy flow, and the hydrologic cycle. The challenge for land managers is to

both understand the functions of soil organisms and to apply this knowledge to improving the health of public

rangelands managed by the BLM.

Keywords: soil organisms, soil biological crust, rangeland management, standards for rangeland health

INTRODUCTION

The management of public lands by the Bureau of Land

Management (BLM) is influenced by climatic, biological,

socioeconomic, and political forces. Over half of the

264 million acres managed by the BLM are considered

rangelands (USDI 1996b) and are managed for multiple

uses (livestock grazing, recreation, aesthetics,

watershed protection, and wilderness values) according

to the Federal Policy and Management Act (FLPMA) of

1976. Rangelands are non-forested lands dominated

by herbaceous species, succulents and shrubs. As

used herein, rangelands are a kind of land, not a use of

land by herbivores such as livestock.

Prior to FLPMA, the emphasis on much of the public

rangelands was commodity oriented with an emphasis

on sustained forage production to support a stable

livestock industry. Recently, the BLM's grazing

regulations were changed to reflect a more ecological

approach to livestock management on public lands.

These new regulations indirectly incorporate soil

organisms into the management of public lands since a

cornerstone of this initiative is maintaining or improving

the functioning of ecological processes (nutrient cycle,

energy flow, and hydrologic cycle).

The term soil organism includes a multitude of

organisms from bacteria to badgers. I will focus my

discussion on two components, 1) soil biological crusts

(surface dwelling lichens, mosses, cyanobacteria,

bacteria, algae and fungi) and, 2) subsurface

microorganisms (bacteria, actinomycetes, fungi and

algae). This emphasis is taken because these two

components are closely tied to soil stability (hydrologic

cycle) and the nutrient cycle and energy flow.

My focus on the role of soil organisms in the

management of public rangelands is directed toward

two topic areas. One is the implementation of the new

grazing regulations entitled, "Standards for Rangeland

Health" and the other is the effect of wildfires and

rehabilitation practices on soil organisms. This is not

an inclusive discussion list, but it includes two topics

with which I am most familiar with relative to the role

of soil organisms in BLM rangeland management.

STANDARDS FOR RANGELAND HEALTH

Before discussing the Standards for Rangeland Health, I

will provide a brief perspective on historical grazing

management policy and monitoring and inventory

procedures. This perspective is important in order to

understand the significance that the recently

implemented Standards for Rangeland Health are

having on public land management, especially in

regards to the role of soil organisms in rangeland

management.

Historical Perspective

Formal grazing management of western public lands

began with passage of the Taylor Grazing Act in 1946

which, in part, addressed the degradation of rangelands

by livestock. Prior to the passage of this Act, livestock

numbers, stocking rates and season of use was largely

unregulated resulting in overgrazed and degraded

rangelands. Accelerated loss of soil and reduced

1Mike Pellant is a rangeland ecologist, Bureau of Land Management, 1387 S. Vinnell Way, Idaho State Office, Boise ID, 83709.

193
rangeland productivity were also problems. The Taylor

Grazing Act set up the framework to resolve these

issues by a process of rangeland adjudication and

adjusting livestock numbers through a forage allocation

process. Ecological processes, including the role of soil

organisms, were not considered in forage allocations,

therefore management decisions, monitoring, and

inventory protocols were designed primarily to measure

and allocate forage for livestock.

Most early monitoring and inventory efforts focused on

the livestock carrying capacity of rangelands or the

degree of accelerated erosion on degraded rangelands.

In 1937, an Interagency Range Survey Committee,

composed of representatives of the major land

management agencies, agreed on a procedure to

conduct range surveys to determine livestock forage

production and watershed protection (Wagner 1989).

Rangeland inventories in the BLM are still focused on

biomass production of vascular plants with no

measures of production or abundance of soil organisms

taken.

Rangeland monitoring studies have traditionally focused

on the trend or direction of change in composition or

condition of "key" plant species (USDI 1996a), which

are generally plants preferred by livestock. For

example, the Parker 3-step method (in use in the

1950's) directed that plant vigor ratings should be

focused on those plants which carry most of the

grazing load (Wagner 1989). The objective of most

rangeland trend studies has been to determine whether

livestock forage species are vigorous and if they were

increasing or decreasing in abundance. Again, soil

organisms were not measured nor were they

considered in the interpretation or evaluations of

monitoring data.

The functional status of the ecological processes and

specifically the soil organisms that contribute to these

processes were seldom considered in designing

monitoring or inventory strategies prior to the

incorporation of rangeland health standards into BLM

rangeland management in 1995. A transition is

occurring at this time as management strategies and

inventory and monitoring procedures are being modified

to better incorporate rangeland health attributes into

protocols.

What Are the Standards for Rangeland Health?

New BLM grazing regulations implemented on February

22, 1995 entitled, "Fundamentals of Rangeland Health"

(43 Code of Federal Regulations Part 4, Subpart 4180)

changed the emphasis of livestock management and

monitoring from "livestock forage and erosion" to

looking at the overall "health" of rangelands.

Rangeland health is defined as "the degree to which

the integrity of the soil and ecological processes of

rangeland ecosystems are maintained" (National

Research Council 1994). The ecological processes

include the water cycle, energy flow and nutrient cycle

and are closely related to the integrity of the plant

community and site stability. These standards were

refined to fit local conditions by public "Resource

Advisory Councils" in each State in concert with BLM

technical specialists.

Soil organisms, especially the above ground and visible

soil biological crusts, are integrally linked to the

ecological processes, therefore they are included in

evaluations of rangeland health. The role of soil

organisms in the functioning of the ecological

processes is well established (Whitford and Herrick

1995). I will now present a few examples of these

roles specific to rangelands.

Nitrogen fixation is a key process in nutrient cycling.

Rychert and others (1978) stated that the biological

crust performed the majority of nitrogen fixation in

some desert ecosystems. Other soil organisms,

primarily bacteria and fungi, are important in the

decomposition process converting litter into organic

carbon. Ninety per cent of this mineralization of carbon

is carried out by soil organisms such as bacteria and

fungi (Swift and Anderson 1993). Many plants in arid

zones are benefitted nutritionally by mycorrhizal fungi

which absorb and translocate nutrients from the soil to

the host plant (Wicklow-Howard 1989). Whitford and

Herrick (1995) summarize the ecological benefits

provided by soil organisms as decomposition, nutrient

immobilization and mineralization, storage and release

of water and nutrients, nitrogen fixation, and

denitrification, production of plant growth factors, and

nutrient uptake by mycorrhizae. Thus, the importance

of soil organisms in nutrient cycling, energy flow and

the water cycle is well established. The dilemma for

land managers is to measure or describe the functional

status of the soil organisms contributing to the

functioning of the ecological processes over large

landscapes.

Incorporating Standards for Rangeland Health Into

Grazing Management Practices

The Standards for Rangeland Health, including the

associated "Guidelines for Grazing Management" (S&

G's) are now incorporated into the management of all

public lands. I will use Idaho as an example of how

one state is incorporating soil organisms into public

rangeland management to maintain or obtain "healthy

rangelands" via the S&G's. The Idaho S&G's were

developed by the three Resource Advisory Councils

(RAC's) with technical assistance from BLM. The

RAC's recognized that soil organisms, especially the

biological soil crusts, were important in maintaining

ecological processes in properly functioning condition.

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The final Idaho S&G's (USDI 1997) reflect this point in

that soil organisms were specifically mentioned in:

Standard 4 (Native Plant Communities)

includes the following indicators for a healthy,

productive, and diverse plant community

where the ecological processes are properly

functioning:

Native plant communities (flora and

microbiotic crust (emphasis added))

are maintained or improved to insure

the proper functioning of ecological

processes and continued productivity

and diversity of native plant species.

The Guidelines for Grazing Management are the

management prescriptions that should be considered if

the Standards for Rangeland Health are not met, not

making significant progress and if livestock

management is the reason for non attainment of the

Standard. Guideline 8 states:

Apply grazing management practices that maintain

or promote the interaction of the hydrologic cycle,

nutrient cycle, and energy flow that will support

the appropriate types and amounts of soil

organisms (emphasis added), plants and animals

appropriate to soil type, climate and land form.

The next step in implementing S&G's is to determine if

current grazing management practices are in

compliance with the S&G's. An interdisciplinary team

evaluates selected landscapes using both qualitative

assessments and available monitoring and inventory

data to make these determinations. Since information

on all aspects of the ecological processes is not

available, "indicators" must be used as surrogates for

the direct assessment of the status of ecological

processes including soil organisms. An indicator is an

element of an ecosystem that is used as an index of

those attributes that are too difficult or expensive to

measure. There is no one indicator of ecosystem

health; instead a suite of key indicators should be used

for an assessment (Karr 1992). A partial list of

indicators of the functional status the ecological

processes and therefore by inference the functional

status of soil organisms includes:

1) The presence of leguminous plants that fix nitrogen

in association with root nodules.

2) Lichens and cyanobacteria that fix nitrogen.

3) Adequate litter and roots (relative to site potential

and climate) available for decay by soil organisms.

4) Normal rates of soil and organic matter loss via

erosion as opposed to accelerated erosion that

removes organic matter and thus soil organism

substrate at a nonsustainable rate.

These indicators are evaluated by first visiting an

ecological reference area (ERA) and reviewing the

appropriate ecological site (range site) description for

the site. This ERA serves as the benchmark for

conducting assessments on other, similar sites being

evaluated. This allows the interdisciplinary team to get

familiar with the amount and distribution of biological

soil crusts, leguminous plants, litter, and natural rates

of erosion prior to conducting evaluations on

landscapes with similar potential to the ERA.

If allotment evaluations reveal that the Standards for

Rangeland Health are not being met, the authorized

officer (manager) must make changes in livestock

management using the Guidelines for Grazing

Management prior to the start of the next grazing

season. Relative to soil organisms, changes in grazing

management practices, if needed, will probably be

focused on soil biological crusts rather than the smaller,

subsurface obvious soil organisms for several reasons.

Soil biological crusts are visible and the impact on them

by livestock is evident. The effects or impacts of

livestock grazing on soil organisms below the soil

surface are much less obvious. Second, research on

the effects of livestock on the survival and recovery of

soil biological crusts, although minimal, is better and

more available than similar research on soil organisms

below the soil surface in the Great Basin Desert.

For example, winter grazing appears to be less

detrimental to soil biological crusts followed by spring

grazing when the soils are moist (Marble and Harper

1989; Memmott and others 1998). Late spring and

summer grazing when the soils are dry are the most

detrimental to crust health and recovery. Of course

grazing management changes must also consider the

vascular plant community along with soil organisms.

Herein lies the challenge of managing rangelands to

insure the maintenance or obtaining of properly

functioning ecological processes. In some Great Basin

plant communities, summer grazing (after seed ripe)

may be favorable for grass and forb species but not for

the biological soil crust and possibly for other soil

organisms. Good management strategies for livestock

grazing will require a better understanding of the

impacts of livestock grazing on soil organisms.

WILDFIRES AND REHABILITATION

PRACTICES

Wildfires have a major impact on resource values in the

Great Basin, particularly in southern Idaho where

wildfires are large and frequent. The BLM mitigates the

harmful impacts of wildfires on public lands by applying

rehabilitation practices that include, but are not limited

to, reseeding, protection from livestock grazing, and

installation of erosion control structures ("gully plugs,"

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contour furrows, and check dams). An interdisciplinary

team evaluates wildfire impacts on vascular plants to

determine if reseeding is needed or if the natural

recovery process can repopulate native plants on a

burned site.

The impacts of wildfires on soil organisms are rarely

evaluated on burned areas for several reasons. First,

the purpose of the BLM's Emergency Fire Rehabilitation

(EFR) program is to mitigate the adverse impacts of

wildfires on watershed function and to reduce risks to

private property in a timely and cost-effective manner.

Restoration of ecological processes, including soil

organisms, is not within the scope of the EFR Program.

Secondly, little is known about the impacts of wildfires

on soil organisms and the appropriate technology to

restore desired soil organisms on large burned areas is

not practical.

However, several cooperative studies sponsored by

BLM and the Biology Department, Boise State

University are providing some insight into the link

between soil organisms, wildfires, and post fire

recovery of vascular plants and biological soil crusts in

the Snake River Plain. Wicklow-Howard (1989)

determined that a high severity wildfire caused a

significant reduction in vesicular-arbuscular mycorrhizae

(VAM) concentration compared to a low severity

wildfire in a Wyoming big sagebrush community in

Southwestern Idaho. However, this effect lasted less

than two growing seasons. In a later study to further

elucidate the relationship between VAM and

reestablishment of sagebrush on burned rangelands,

Gurr (1993) hypothesized that a summer burn released

sufficient nutrients so that inoculating sagebrush

seedlings with VAM was unnecessary.

More recently Kaltenecker (1997) studied the effects of

wildfire and rehabilitation on biological soil crusts in

Wyoming big sagebrush (Artemisia tridentata) sites

with high potential for cheatgrass invasion. She

determined that reestablishment of a plant community

structure similar to that of the native plant community

(widely spaced perennial plants with open interspaces)

was necessary for recovery of the biological crust, and

that vascular plant and biological crust establishment

were synergistic. This was based on evidence of high

cover of biological crust in ungrazed, rehabilitated

communities, with diversity of lichens and mosses only

slightly lower than that found in unbumed sagebrush

communities. High plant densities, litter accumulation,

and the lack of open interspaces in communities

dominated by cheatgrass created conditions that were

not conducive to establishment of a fully developed

biological crust.

One area where soil organisms have proved to be

beneficial to the success of rehabilitation projects is the

addition of Rhizobium meliloti innoculant to leguminous

species, primarily dryland alfalfa (Medicago sativa),

included in rehabilitation seed mixtures. I have

observed that innoculated dryland alfalfa appears to

establish better than non innoculate legumes.

SUMMARY

Soil organisms, especially the biological soil crusts, are

receiving more attention on public lands administered

by the BLM with the implementation of State S&G's.

With this increased attention comes the need for a

better understanding of the effects of grazing and other

uses on soil organisms. Science must play an

important role in supporting management decisions

relative to changing livestock practices to support soil

organisms and therefore rangeland health. Existing

science must be made available to resource specialists

and managers now to insure that management

decisions are based on the best available science. In

addition, new studies are needed to fill information

gaps relative to the effects of management practices

on the role of soil organisms in rangeland health. This

research is needed quickly and must be distributed to

resource specialists, managers and the public in both

technical and nontechnical formats.

In addition to the need for an increased research in this

area, managers need to begin incorporating soil

biological crusts into rangeland monitoring and

inventory programs. Soil biological crusts can be easily

incorporated into rangeland monitoring and inventory

programs. In Idaho, cover and frequency of soil

biological crust components (mosses and lichens) are

recorded along with vascular plants in some rangeland

trend studies. Rangeland management decisions will

improve as this baseline information on soil biological

crusts is accumulated and incorporated into new

grazing management strategies.

The last issue that I would like to address in this

summary is the polarization that is occurring between

some "environmental" and "commodity" users of the

public lands over the importance of soil organisms in

the management of public rangelands. Again I will

focus on the soil biological crusts since most of the

current controversy surrounds this component of the

soil organism community. One extreme perspective of

the "pro biological crust" public was provided by Hogan

(1994). He states that loss of "cryptogams" has

greatly contributed to the loss of half of the topsoil on

arid lands and the solution is, "total removal of

livestock from western North American arid lands."

The media is promoting the controversy by printing

stories with titles of "West tramples key asset into

dust; rangeland 'skin'" (Twin Falls Idaho Times-News,

July 28, 1998). On the other side of the coin are

Savory and Parsons (1980) who promote grazing

196
practices designed to "break the crust" to increase

vascular plant establishment. This management

approach recognizes soil crusts as a sign of

deterioration on western rangelands that needs to be

"broken" by hooved animals in order to promote proper

functioning of rangeland ecosystems.

I feel that the middle road must be taken in issues that

concern the management of soil biological crusts in

particular and of soil organisms in general. Focusing on

just soil organisms for changing grazing management

practices neglects the functional links between vascular

plants, soil organisms and the ecological processes.

Litter and roots from vascular plants provide the

substrate for soil microorganisms to break down

organic material to carbon, while leguminous plants are

the host for Rhizobium species that provide nitrogen

used by the plants. There are many other examples of

the linkages between vascular plants and soil

organisms that could be cited to illustrate this point

(Whitford and Herrick,1996). Singling out one

component of the ecosystem for special management

is not the best approach for managing rangelands for

overall health. Management of our public rangelands

for health and productivity will be encouraged if

polarization is avoided and rangeland ecosystems are

managed as a whole and not divided into their parts to

forward different agendas.

LITERATURE CITED

Gurr, Jan E. 1993. The effect of fire on vesicular

arbuscular mycorrhizae, soil nutrients, and the reestablishment of Wyoming big sagebrush. Thesis

(M.S. Interdisciplinary Studies, Soil-Plant

Relationships). Boise State University, Boise, ID.

92 p.

Hogan, David. 1994. Cryptogamic doomsday. Wild

Earth: 26-28.

Kaltenecker, Julienne H. 1997. The recovery of

microbiotic crusts following post-fire rehabilitation

on rangelands of the western Snake River plain.

Thesis (M.S. Interdisciplinary Studies, Plant

Ecology and Soil Science). Boise State University,

Boise, ID. 99 p.

Karr, J.R. 1992. Ecological integrity: Protecting

earth's life support systems. In Costanza, R.;

Norton, B.G.; Haskell B.D. (eds.). Ecosystem

Health-New Goals for Environmental Management,

Island Press, Washington, D.C: 223-238.

Marble, J.R.; Harper, K.T. 1989. Effect of timing of

grazing on soil-surface cryptogamic communities in

a Great Basin low-shrub desert: a preliminary

report. Great Basin Naturalist 49: 104-107.

Memmott, K.L.; Anderson, V.M.; Monsen, S.B. 1998.

Seasonal grazing impact on cryptogamic crusts in a

cold desert ecosystem. Journal of Range

Management 51: 547-550.

National Research Council, Committee on Rangeland

Classification. 1994. Rangeland health: new

methods to classify, inventory, and monitor

rangelands. National Academy Press, Washington,

D.C. 180 p.

Rychert, R.; Skujins, J.; Sorensen, D.; Porcella, D.

1978. Nitrogen fixation by lichens and free-living

microorganisms in deserts. In: N.E. West and J.

Skujins, eds. Nitrogen in desert ecosystems.

Dowden, Hutchinson and Ross, Inc. Stroudsburg,

Penn: 20-30.

Savory, A.; Parsons, S.D. 1980. The Savory grazing

method. Rangelands. 2:234-237.

Swift, M.J.; Anderson, J.M. 1993. Biodiversity and

ecosystem function in agricultural systems. In:

Biodiversity and Ecosystem Function. Schluze,

E.D.; Mooney, H.H. Springer-Verlag. Berlin. 330-

345 pp.

U.S. Department of the Interior. 1996a. Utilization

studies and residual measurements. Interagency

Technical Reference BLM/RS/ST-96/004+ 1730.

USDI, Bureau of Land Management, National

Applied Resources Science Center, Denver, CO.

165 p.

U.S. Department of the Interior. 1996b. Public Land

Statistics. USDI, Bureau of Land Management,

Washington, D.C.:U.S. Department of the Interior,

Bureau of Land Management, Washington Office.

169 p.

U.S. Department of the Interior. 1997. Idaho

Standards for Rangeland Health and Guidelines for

Livestock Grazing Management. USDI, Bureau of

Land Management, Idaho State Office, Boise,

Idaho. 18 p.

Wagner, R.E. 1989. History and development of site

and condition criteria in the Bureau of Land

Management. In: Lauenroth, W.K.; Laycock, W.A.

(eds.). Secondary succession and the Evaluation of

Rangeland Condition, Westview, Boulder, CO: 35-

48.

Whitford, Walter G.; Herrick, Jeffrey E. 1996.

Maintaining soil processes for plant productivity

and community dynamics. Rangelands in a

sustainable biosphere: proceedings of the fifth

International Rangeland Congress. Salt Lake City,

Utah, July 23-28. 1995. Volume II Society for

Range Management, Denver, CO: p 33-37.

Wicklow-Howard, M. 1989. The occurrence of

vesicular-arbuscular mycorrhizae in burned areas of

the Snake River Birds of Prey Area, Idaho.

Mycotaxon 34(1): 253-257.

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NATURAL RESOURCES CONSERVATION SERVICE-DATABASES AND ANALYTICAL

SERVICES

Carol D. Franks1

ABSTRACT

This paper is a brief overview of NRCS databases, the National Soil Survey Laboratory (NSSL) soil analyses and

characterization, and the National Soil Survey Center's (NSSC) new program in Soil Biology and Soil Carbon. The NRCS

soils and plant community databases (NASIS and PLANTS) are discussed (Internet access: NASIS

http://www.statlab.iastate.edu/soils/, and PLANTS http://plants.usda.gov/). Current NSSL soil characterization analyses

as well as new projects and analyses for the Soil Biology and Soil Carbon program are presented. The Soil Biology and

Soil Carbon program augments our existing emphasis areas of Global Climate Change, Carbon Sequestration, Soil

Survey, and Technical Soil Services. Benchmark Site Selection is presented as a case study for database development

that can be adapted or modified to fit a variety of situations.

Keywords: soil databases, NASIS, plant databases, PLANTS, soil analysis, benchmark sites, organic carbon analyses,

microbial biomass, microbial activity

INTRODUCTION

The Natural Resources Conservation Service (NRCS)

estimates that our corporate databases are worth $8.5

billion. The NRCS has 6 major databases; soils, snow

survey, plants, field office, Natural Resources Inventory

(NRI) and others. These individual databases are

valued (billions of dollars); soils - $5.0, snow survey -

$0.3, plants - $0.2, field offices - $2.7, NRI - $0.2,

and other $0.1 (Harward, 1997). Our databases are

one of our most important assets. We use them

extensively, and we want other groups to have access

to them as well. The first section of this paper covers

key points about three of our databases, NASIS,

PLANTS and the NSSL Soil Characterization database

(a subset of NASIS).

The second section of this paper deals with the

National Soil Survey Centers new program in Soil

Biology and Soil Carbon. Finally, Benchmark Site

Selection is presented as a case study that ties

together all aspects of database development and can

be adapted to a variety of situations.

SOILS DATABASES

The National Soil Information System-NASIS

(http://www.statlab.iastate.edu/soils/)

NASIS is the new NRCS soils database. NASIS

replaces our State Soil Survey Database (3SD). NASIS

is designed to meet the following objectives:

1. Provide a flexible and dynamic management system

for soil data and information.

2. Improve the quality of soil data and information.

3. Improve automated map unit management.

With these objectives in mind NRCS focused on

customer data needs. The following areas of database

development were emphasized and represent the

components of an information management framework:

1) data collection, accumulation and aggregation (from

field observations to integrated soil survey maps and

interpretations); 2) data management and access; and

3) database application. Key functions and data that

NRCS typically uses were also identified. These key

functions include coordination, data acquisition and

data input. Types of data NRCS uses include soil,

water, air, plant, animal and others.

Soils data are currently used in the following

applications:

Water Quality Models

Water Balance/Budgets

Engineering Practices

RUSLE - Revised Universal Soil Loss Equation

Agricultural Waste Management

1Carol D. Franks is a research soil scientist, Natural Resources Conservation Service, National Soil Survey Center, Lincoln, NE

68508-3866.

198
Pesticide Management

Interpretive Maps

Conservation Practices

Conservation Practice Effects

Wind Erosion Equations

Grazing Lands Application

Grazing Lands Data System

Environmental Planning

National Resource Inventory

The following are NASIS priorities for new

development:

Lab Characterization data model

Convert pedon and lab data

Point/site data collection tool

Spatial analysis

Manuscript generation/management

Official series descriptions

Convert STATSGO Attribute data

NASIS interface enhancements

Generalized export

Improved report functionality

Improved query functionality

Natural resource data gateway

Data aggregation

Data comparison

In order to meet the objectives outlined above, NRCS

developed the National Soil Data Access Facility

(NSDAF), in conjunction with Iowa State University.

NASIS and NSDAF are designed for the following

applications; automated geographic tabular natural

resource databases supporting conservation planning

tools, resource inventory tools and technical

procedures, conservation effects tools, and practice

design tools. Future data access features will include;

a USDA wide area network, including all NRCS offices,

NSDAF, U.S. Forest Service, universities and clients.

The National Soil Data Access Facility-NSDAF

(http://www.statlab.iastate.edu/soils/)

The purpose of NSDAF is to serve as a clearinghouse

for map unit data, site data, soil geodata, and

taxonomic data for data producers and users such as;

universities, federal agencies, and the private sector.

The guiding principle for NSDAF is that anyone

requiring or producing soil geodata shall have the ability

to easily determine what exists, the ability to easily

access this data, and the ability to contribute data to

the National Repository. All NRCS standards for data

collection, input and quality will apply to data in this

facility.

NRCS National Soil Survey Center and National

Soil Survey Laboratory (NSSL) Database

This database includes standard soil characterization

analyses for more than 20,000 soils from around the

world and throughout the United States. There are

four types of standard soil characterization, acid soils,

alkaline soils, organic soils and andic soils. These four

types of analyses are based on characteristics of the

soil sampled. Acid soils typically have pHs less than 7,

lower base saturation and can be dominated by

aluminum chemistry. Alkaline soils typically have pHs

greater than 7, higher base saturation and can contain

appreciable amounts of chemical salts such as calcium

and magnesium carbonate and sodium chloride.

Organic soils typically contain more than 12 percent

organic carbon and are often called "peat" or "muck".

Andic soils typically formed in weakly weathered

volcanic materials dominated by glass (ex: volcanic

ash). Listed below are standard analyses NSSL

performs on the different soil types.

1. Acid Soils-bulk density, 1500 MPa water retention,

particle-size, total carbon, total nitrogen, calcium

carbonate equivalent (<2mm), organic carbon

(difference method), pH, cation exchange capacity,

extractable acidity, aluminum, iron, silica,

manganese, and phosphorous (Bray 1).

2. Alkaline Soils-bulk density, 1500 MPa water

retention, particle-size, total carbon, total nitrogen,

calcium carbonate equivalent (<2mm), organic

carbon (difference method), pH, cation exchange

capacity, electrical conductivity, and calcium,

magnesium, sodium and potassium (from the

saturation extract).

3. Organic Soils-bulk density, 1500 MPa water

retention, organic fiber content, mineral content,

total nitrogen, pH, cation exchange capacity,

extractable acidity, iron, silica, aluminum and

manganese.

4. Andic Soils-clay mineralogy, glass content, pH,

iron, aluminum, silica, manganese, and

phosphorous.

PLANT DATABASES

The National PLANTS Database was designed by NRCS

to provide a single source of standardized information

about plants. This database focuses primarily on plants

of the US and its territories. This database includes

checklists, distributional data, crop information, plant

symbols, plant growth data, references and other plant

information. PLANTS reduces costs by preventing the

duplication of efforts and by making information

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exchange possible across agencies and disciplines.

These Web pages allow you to query the PLANTS

database, download information, and view images of

plants.

NRCS PLANTS Database

(http://plants.usda.gov/)

The NRCS PLANTS database consists of six subsets:

1. Agronomy

2. Integrated Taxonomic Information System (ITIS)

3. National Plant Data Center (NPDC)

4. Ecological Site Information System (ESIS)

5. Vegetative Practice Design Application (VegSpec)

6. Plant Materials Program

The Agronomy Projects subset supports various

agronomy projects with data and data management

software. Supported agronomy projects include the

Revised Universal Soil Loss Equation (RUSLE) and Wind

Erosion Equation (WEQ) software applications.

The Integrated Taxonomic Information System (ITIS) is

a partnership of U.S., Canadian, and Mexican agencies,

other organizations, and taxonomic specialists

cooperating on the development of an on-line,

scientifically credible, list of biological names focusing

on the biota of North America. ITIS is also a

participating member of Species 2000, an international

project indexing the world's known species. ITIS data

may be accessed via the Web.

The National Plant Data Center (NPDC) focuses

resources on the acquisition, development, integration,

quality control, dissemination and access of plant

information; NPDC incorporates new plant technology

into the agency's automated tools. The NPDC is

responsible for the PLANTS database, which focuses

on the vascular and nonvascular plants of the United

States and its territories.

The Ecological Site Information System (ESIS) contains

the ecological site inventory (Range 417) database for

Rangeland Data, ecological site inventory (Wood 5)

database for Forest Soils, and the Ecological Site

Description (Range Sites) database. Currently, the

Ecological Site Inventory Database for Rangeland Data

is available. Later this summer the Ecological Site

Inventory for Forest Soils will be available. The

Ecological Site Description database will be available

sometime next year (pending data input).

The Vegetative Practice Design Application (VegSpec)

provides practice design technology for the preparation

of vegetative establishment specifications. The

vegetative establishment practices for which VegSpec

automates practice design include cover crop, critical

area planting, filter strip, pasture and hayland planting,

range seeding, tree planting, farmstead windbreak and

field windbreak. This application is interactive and fully

functional.

The Plant Materials Program develops plant materials

and plant technology for the conservation of the

Nation's natural resources. Many of today's

environmental problems can be addressed effectively

through the use of plants. Plant Materials centers and

specialists provide state-of-the-art technology to help

field offices and land users solve critical natural

resource issues. The Plant Materials program strives to

meet the change by (1) developing new technology for

current conservation needs, and (2) Transferring this

information to its customers.

SOIL ECOLOGY AND BIOLOGY

PROGRAM-NRCS NSSC

The purpose of this program is to branch out into new

areas of soil science in order to enhance our existing

Soil Survey Program, support our Global Climate

Change Program, support Soil Quality initiatives and

develop new soil interpretations and assess

management impacts on the soil. To accomplish this

task the National Soil Survey Center and the Soil

Survey Division are developing this new program in Soil

Biology and Soil Carbon. The following outlines areas

of emphasis for this new program.

Soil Biology Areas of Emphasis

1. Soil and Ecosystem Function

Soil Mapping

Technical Soil Services

2. Soil Quality

Soil Function

Microbes and Soil Climate

Microbes and Plant Communities

Microbial Biodiversity

Management impacts

Database Development

3. Wetland Identification and Function

4. Global Climate Change

Nutrient Cycling

Ecosystem Modeling

Threshold Identification

5. Resource Inventory and Assessment

National Resource Inventory (NRI)

Interagency Inventories

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6. Field Sampling and Observations

Near Surface Soil Properties

Field Sampling Method Adaptation and Development

7. Database Development

Modeling

Benchmark Sites-a Framework

8. Soil Characterization

Lab Analyses Methods

Adaptation and Development

Research and Development

Implementation

9. Soil Interpretations and Management Impacts

Rangeland

Cropland

Forestland

10. Bioremediation

The following section provides objectives for the eleven

soil biology and soil carbon projects NRCS is pursuing

at this time.

Benchmark sites-A framework to develop a pilot

database of microspatial and temporal near surface soil

properties and plant communities suitable for use in

modeling efforts, based on benchmark sites.

Microbial community fingerprinting-Test the Fatty

Acid Methyl Ester (FAME) microbial analysis for use on

archive benchmark soil samples dried at 45°C.

Resample benchmark sites and perform analyses on

fresh samples/ FAME, ergosterol (fungi), total biomass,

dehydrogenase activity and C and N mineralization.

Soil food web database-Cooperative effort between

NRCS Soil Quality Institute and National Soil Survey

Center (NSSC) and Oregon State University (OSU) Soil

Microbial Biomass Service. To link microbial data

collected to the NRCS interpretive soil database (soil

survey map units) through georeferencing.

Development of a Logical Data Model and

linkage-Completed,

Expanded organic carbon analyses-Test minimum

dataset for organic carbon for applicability to NRCS

programs and projects. Build a characterization

database based on benchmark sites that includes data

elements needed in carbon and nutrient modeling.

Microbial activity and biomass analyses-Adapt,

develop and implement NSSL capability to analyze

fresh soil samples for total microbial activity using

Chloroform Fumigation Incubation (CFI). In cooperation

with ARS test soil samples for the Conservation reserve

Program and International Projects.

Field sampling protocols-Document and implement

appropriate field sampling methods for all microbial

analyses.

Soil chemical, physical, and microbial characterization of benchmark sites-ln cooperation with OSU

Soil Microbial Biomass Service: analyze fresh soil

samples using extraction techniques to determine

various biomass measures; NSSL will characterize the

same benchmark site thus creating a point/site linkage

between the lab characterization and the microbial

biomass database under development by OSU; compare

NSSL incubation methods to OSU direct extraction

methods.

Oregon demonstration project-To test the potential

to combine federal inventories such as-NRI, FS-FIA

and FSIS, NASS enumerative survey, etc.; test the

various plant community inventory protocols in all of

the three major land uses; cropland, rangeland and

forestland; test repeatability of measurements between

field crews through double sampling; field test potential

Soil Quality Indicators in a national periodic inventory.

NRST/IRWET database-Use this WEPP dataset to

evaluate soil quality and nutrient cycling for the soils

sampled.

Wetland ecosystem project-Develop a laboratory

analysis to measure anaerobic N or C for use on

wetland soils.

Macronutrient cycling-Develop additional laboratory

analyses for evaluating C, N and P cycling in soils.

In order to complete these projects, we have

developed a variety of new soil analyses: microbial

biomass/activity, mineralizable nitrogen, particulate

organic matter (POM), organic matter residues (0

horizons and litter), root biomass and labile carbon.

These analyses are available on a limited basis for other

selected projects. In the future, we hope to offer these

standard analyses for the National Soil Survey

Laboratory.

Benchmark Site Selection

The following section is an outline for using Benchmark

Site Selection. This process can be used and modified

to focus research efforts, soil survey efforts, resource

inventory efforts or resource assessment efforts to

make them more efficient, cost effective and reduce

the number of sites needed for evaluation. This comes

from work I did for the Rangeland Component of the

interagency Water Erosion Prediction Project (WEPP) to

develop the new process based WEPP model for

erosion prediction (Franks and others 1993).

201
Why do we need a site selection process? With limited

time and resources we must effectively augment

existing erosion, soils and hydrology research on plant

community ecosystems in order to enhance the

modeling effort. This process is based on the concept

of a benchmark. Webster's New World Dictionary,

1988 defines a benchmark as a standard or point of

reference in measuring or judging quality, value, etc.

NRCS has used this standard to develop the concept of

a Benchmark Soil. A benchmark soil is one of large

extent, one that holds a key position in the soil

classification system, or of special significance to

farming, engineering, forestry, or other uses.

Knowledge of benchmark soils can be applied to other

soils with similar properties. The purpose of

benchmark soils lists is to focus research efforts on

soils that have the greatest potential for expansion of

data and interpretations. Lists of benchmark soils are

useful in planning for many kinds of studies.

Current benchmark soils lists have limitations because

they are based on political boundaries such as counties

and states. Typically, such lists emphasize agronomic

uses and are limited to 20 or 30 soils per state.

However, we can use many of the criteria used in

developing benchmark soils lists for other purposes. To

this end I have developed the Benchmark Site Selection

process (Franks and others 1993). The steps in this

process are:

1. Define what constitutes a benchmark for purposes

of the study.

2. Review benchmark soil criteria for applicability.

3. Develop new criteria essential to the study.

4. Determine which criteria occur in the national soils

database.

5. Access the database.

6. Develop new potential benchmark soils lists.

7. Select soils from this new list which have potential

sites where land use history is known and available;

has contrasting vegetation states (different plant

comminutes or cropping systems); easy access so

one can return and conduct further studies later.

8. Identify other limiting factors that effect the site

selection process.

The benchmark site selection process is an organized,

logical, criteria-based method to assist in the site

selection process for field studies. It can be used and

modified to enhance site selection for any soil-based

research for process-based modeling. All data gathered

from all sites must have the greatest potential possible

for extrapolation to larger areas, similar plant

community sites, and similar soils. The following maps

(fig. 1 and 2) illustrate how we used only twelve major

sites to collect data that can be extrapolated to similar

soils found in the region.

LITERATURE CITED

Franks, CD.; Brockmann, L.C.; Whited, M.; Waltman,

W.W. 1993. Using the Benchmark Site Selection

Process to Enhance Modeling. In: Pierson, F.B.;

Spaeth, K.E. (eds.). Predicting Range/and Hydrologic

and Erosion Processes. Proceedings of a

Symposium, February 16, 1993. The 46th Annual

Meeting of the Society for Range Management,

Albuquerque, New Mexico, February 14-19, 1993

202
National Range Study Team (NRST)

Major Land Resource Areas where NRST sites occur

Site

Designation

B

C

E

F

G

H

I

J

K

L

M

State

Nebraska

Texas

Kansas

Colorado

Wyoming

North Dakota

Wyoming

Idaho

Arizona

California

Utah

MLRA

M106

H77

H76

G67

G60A

F54

G58B

A13

D35

C15

D28A

Figure 1-This map shows the geographic distribution of the Major Land Resource Areas

(MLRAs) where NRST sites are found. NRST data can be extrapolated to apply to similar

soils and plant communities within the designated MLRA. Map source is Major Land

Resource Areas of the United States. NRCS, National Soil Survey Center, 1997. Table A

defines the respective MLRA for each state.

203
National Range Study Team (NRST)

Additional Major Land Resource Areas where the soil series found at

an NRST site also occurs

Site

Designation

B

C

G

H

I

J

K

L

M

State

Nebraska

Texas

Kansas

Colorado

Wyoming

North Dakota

Wyoming

Idaho

Arizona

California

Utah

MLRA

M106

H77

H76

G67

G60A

F54

G58B

A13

D35

C15

D28A

Additional MLRAs

107,75,102B

78,78A

76,106,112

69,72,49

32,33,58B,34,67

53A, 53B, 58B,58A

32,33,34,60A, 67

43

39,36

20

28B

Figure 2-This map shows the geographic distribution of additional MLRAs where the NRST

data may be extrapolated to similar soils and similar regions within these MLRAs. These

additional MLRAs were selected because the Soil Interpretation Records (NRCS, National Soils

Information System (NASIS), 1997) for the Soil Series found on the NRST sites indicate that

these Soil Series also occur within these additional MLRAs. Some MLRAs found on the Soil

Interpretation Records were excluded because of poor fit with the plant communities, range

cover regions, and range cover types (Shiflet 1994). Map source is Major Land Resource Areas

of the United States, NRCS, National Soil Survey Center, 1997. Table C provides the

descriptive names of the additional MLRAs.

204

C

C
SOIL ORGANISMS: FUNCTIONS AND PROCESSES-MANAGEMENT IMPLICATIONS.

A SYNOPSIS

Martin F. Jurgensen1

ABSTRACT

The objective of this symposium was to examine the various roles and functions of soil organisms, and how they relate

to forest and range management in the Pacific Northwest. This paper is a synopsis of my perspectives on the main

topics of the symposium. They were: 1) why are soil organisms important, 2) management impacts on soil organisms,

and 3) management of specific soil organisms. The consensus of the symposium speakers was that forest and range

management activities do have pronounced effects on soil organisms, but it was not clear if management-induced

changes in population numbers and diversity are important for long-term soil productivity and stability. Until more

information becomes available, forests and rangelands should be managed in ways which maintain soil organism

populations and diversity, and are consistent with overall management objectives. Management activities, such as

maintaining soil organic matter levels and reducing soil compaction, are usually beneficial to both soil organisms and

other components of forest and range ecosystems.

Keywords: soil organisms, organic matter decomposition, soil compaction, microbial diversity, timber harvesting

INTRODUCTION

Much has been written about the significance of soil

organisms to the functioning of forest and range

ecosystems, especially their role in forest health and

soil productivity/sustainability (Meurisse - this

symposium; Harvey 1994; Hessburg and others 1994).

However, many questions are being raised whether soil

organisms can be managed, and if they can, should

they become an integral part of management activities.

It was the objective of this symposium to examine the

various roles and functions of soil organisms, and how

they relate to forest and range management in the

Pacific Northwest (Ypsilantis - this symposium). The

papers presented at this meeting covered a wide range

of topics - from the very basic (genetic diversity of soil

microorganisms) to the very applied (landscape

considerations), many of which are published in this

proceedings. Rather than give detailed comments on

individual papers, I have decided to give my perspective

on what I feel were the main topics or focus of the

symposium. They were: 1) why are soil organisms

important, 2) management impacts on soil organisms,

and 3) management of specific soil organisms.

Unfortunately I was not able to read most of the papers

published in this symposium proceedings. As often

happens, papers published in a proceedings may be

quite different than the oral presentations. Therefore,

this paper is based on my notes and memory of

information given by the various authors at the

symposium. Consequently, if I have misinterpreted or

misquoted any of the symposium speakers, my

apologies.

IMPORTANCE OF SOIL ORGANISMS

Plants

While soil organisms affect many components of forest

and range ecosystems, the main focus of papers

presented was the impacts of soil organisms on plant

communities. The role of particular microbial and faunal

groups in organic matter decomposition and nutrient

release were high-lighted by Ingham, Norton, Molina,

Moore, and Seastedt (this symposium). Since nearly all

soil nitrogen is in organic matter and many western

forest soils are nitrogen limited (Binkley 1991; Moore

1994), the importance of soil organisms in the nitrogen

cycle, especially dinitrogen fixation and nitrogen

mineralization, was repeatedly stressed (Myrold; Stark;

Prescott - this symposium). Soil organisms also affect

plant water availability and aeration by incorporating

organic matter in mineral soil horizons (soil structure),

and forming large macropore (biopore) channels

(Herrick; Belnap - this symposium). The significance of

forest pathogens on ecosystem functions and forest

health was addressed by both Thies and Harvey (this

symposium). The comments by Thies that root disease

activity does not indicate an "unhealthy" forest, and

1Martin F. Jurgensen is a professor of forest soils, School of Forestry and Wood Products, Michigan Technological University,

Houghton, Ml 49931

205
that wood "lost" to pathogens is a social, not a

biological concept should be remembered when forest

health concerns and management are discussed. No

mention was made of growth factors and other plantstimulating compounds produced by soil

microorganisms (Colinas and others 1994), although

Ingham (this symposium) mentioned possible beneficial

effects of rhizosphere bacteria in protecting roots from

disease attack. Mycorrhizal fungi and other soil

microflora have been reported to produce antibiotics or

other metabolic products which inhibit various root

pathogens (Azcon-Aguilar and Barea 1992).

Water

The impacts of soil organisms on water quality was

indirectly addressed by the role of soil microorganisms

in the nitrification and denitrification processes (Myrold;

Stark - this symposium), which would affect the

amount of nitrate available for leaching into ground

water or surface streams and lakes. I was surprised

there was not a greater emphasis on this issue at this

meeting, considering the interest in managing riparian

zones of western forests. This is of major concern in

the southern U.S., where extensive research is being

conducted on the biological filtering processes in forest

buffer strips along streams and wetlands (Hubbard and

Lowrance 1994).

Animals

While animals would be indirectly affected by plant

community changes caused by soil organism activity,

the direct involvement of mycorrhizal fungi fruiting

bodies, as part of the food chain supporting spotted

owl populations, was discussed by both Molina and

Amaranthus (this symposium). Soil organisms also

seem to be an important food source for another

endangered species, the red cockaded woodpecker,

which is found in old-growth pine stands in the

southeastern U.S. After Hurricane Andrew in1989

destroyed many pine stands in the Coastal Plain of

South Carolina, it was projected that populations of

this woodpecker would plumate. Instead, numbers

increased to record numbers in some areas, as the

woodpeckers fed on insects associated with the large

amounts of coarse woody residue produced by the

hurricane2.

Biodiversity

The concept of soil organism biodiversity as being

importance to forest ecosystems and their management

is relatively new, and has generated considerable

2Personal communication. 1997. Carl Trettin, Center for

Forested Wetlands, Southern Research Station, Charleston,

SC.

interest and possible concern. While adequate

populations of soil organisms are recognized to be

necessary in carrying out the various processes needed

for plant, water and animal activity (discussed above),

maintaining the diversity of these soil organisms has

also become an issue (Molina; Spence; Moldenke-this

symposium). Soil microorganisms and microfauna are

the largest and most diverse of the biological groups,

and they are extremely difficult to identify and obtain

accurate estimates of population size. It has been

estimated that only 1 to 5% of soil microorganisms

have been characterized and named (Hawksworth and

Mound 1991). Consequently, much effort is going to

develop technologies for characterizing the taxonomic

and functional diversity of soil organism populations

(Busse; Franks - this symposium; Kennedy and Gewin

1997).

The reasons proposed for conserving soil organism

biodiversity are also diverse. One, which appeals to

the practical, utilitarian part of human nature, is

conservation of the large gene pool present in soil

organisms. These organisms, their genes, enzymes, or

metabolic bi-products have the potential to benefit

society, such as in the mitigation of polluted soil and

for antibiotic production. This trend is exemplified by

the recent proposal between a drug company and

Yellowstone National Park to search for unique

microorganisms in hot pools. At the other end of the

spectrum is the idea that soil organism diversity should

be maintained because "they are there", and have an

intrinsic worth in themselves beyond any practical

value. Spense (this symposium) noted that biodiversity

is a "value choice —a thing of the heart". This concept

is best illustrated by the Deep Ecology Platform of

Naess and Sessions (in Devall 1993).

Of more significance to forest and range ecosytems

and their management is the importance of soil

organism diversity to soil productivity and

sustainability. Increased plant diversity has been

positively correlated with increased ecosystem

productivity (Naeem and others 1996; Tilman 1996).

Losing one or several plant species from an ecosystem

can have pronounced effects on community

productivity and nutrient retention (Symstad and others

1998). These have been termed "keystones" species,

to signify the important role they have in ecosystem

functioning. While the loss of a plant species may

have a major impact on ecosystem productivity or

stability, this has not been clearly shown for soil

organisms. The difficulty in isolating and categorizing

the soil microflora into discrete taxonomic units has

made this concept seemingly unusable for soil bacteria

and fungi. Moldenke (this symposium) used keystone

species to describe certain soil fauna active in the

decomposition of forest floor litter.

206
Instead of keystone species Amaranthus (this

symposium) proposed using keystone microbial "groups

or functions", such as nitrogen-fixing bacteria and

mycorrhizal fungi, which are critical for maintaining

ecosystem productivity. This concept recognizes that

many different groups of soil organisms are active in

the same soil process, for example - nitrogen

mineralization. This overlap in function or

"redundancy" by soil organisms is an important factor

in determining what impact forest and range

management has on soil ecosystems (Moore -this

symposium). Unfortunately, the degree of microbial

redundancy for most soil processes is not known

(Busse - this symposium).

MANAGEMENT IMPACTS ON SOIL

ORGANISMS

Forest and range management can have profound

effects on numbers, activity and diversity of soil

organisms. These effects are either positive or

negative, depending on one's outlook or management

objectives. Boyle (this symposium) took an optimistic

view on soil resilience, and concluded that

management had very little negative impacts on longterm soil functions, except under extreme conditions.

Others (Belnap; Bennett; Ingham; Moldenke - this

symposium) tended to look more on "the dark side",

and concluded very little is known on long-term,

management-induced changes of soil organism

populations and processes.

Considerable information is available on the impacts of

timber harvesting and subsequent site preparation on

soil biological populations and activities (Edmonds;

Ingham; Moldenke; Molina; Moore; Norton; Visser - this

symposium). Similar information is available for grazing

on western rangelands (Belnap; Stark - this

symposium). The decrease in numbers and types of soil

microorganisms and microfauna usually found after

harvesting and site preparation is largely a response to

lower soil organic matter levels, and changes in soil

temperature and moisture conditions (Johnson 1992;

Jurgensen and others 1997). Lower C:N ratios and

lignin contents in litter from early successional plants

established after timber harvesting (Myrold; Prescott -

this symposium) could also affect the composition of

the microbial communities. Ingham (this symposium)

indicated that in western Oregon bacteria populations

increased in young successional stands, compared to

fungal-dominated soil populations in late seral and oldgrowth forests. She contends that by reducing soil

fungal populations and increasing soil bacteria,

clearcutting will decrease site productivity. In contrast,

Norton (this symposium) reported that bacteria biomass

decreased and the fungi/bacteria biomass ratio

increased after a clearcutting. Since bacteria have a

more rapid turnover than fungi, she concluded that

bacteria are more active than fungi in soil nutrient

availability.

Similar to timber harvesting, wildfire or prescribed

burning affects the types and numbers of soil

organisms by lowering soil organic matter levels and

changing plant litter quality (DeBano and others 1998).

However, a wildfire in an older, larger stand could

increase the amount of coarse woody debris (Edmonds

- this symposium). Fire favors the development of soil

bacteria by decreasing soil acidity (Norton - this

symposium). Of particular significance is an increase of

nitrifying bacteria, which increases soil nitrate levels

and possible nitrogen leaching losses (Myrold - this

symposium). Volatilization losses of soil nitrogen could

also increase the activity of symbiotic and

nonsymbiotic nitrogen-fixing microorganisms after fire

(Jurgensen and others 1997).

Soil compaction by logging equipment would directly

restrict the movement of various soil fauna in the

mineral soil by decreasing pore size, especially

macropores, and the burrowing-ability of larger soil

fauna, such as earthworms. Both the soil microflora

and fauna would be affected by changes in soil

moisture and oxygen levels at higher bulk densities

(Startsev and others 1998). All of these factors could

account for reported decreases in mycorrhizae numbers

and types in compacted soil (Amaranthus and others

1996). How long compaction would impact soil

organisms depends on how quickly soil bulk density

returns to preharvest conditions. If soil compaction

lowers plant growth, the reduction in the amount of

organic matter returned as litter and root turnover

would also affect soil organism populations and

activity. The impact of both soil compaction and soil

organic matter levels on long-term site productivity is

being investigated on various sites throughout the

United Sates and Canada (Powers 1991).

MANAGEMENT OF SOIL ORGANISMS

The main focus of most public and private forest

management has been on forest protection and

productivity. Consequently, the management of forest

soil organisms has dealt mostly with specific groups,

which had either detrimental or beneficial effects on

tree growth or stand development. Both Thies and

Harvey (this symposium) discussed the implications of

forest management operations on root diseases, and

how this relates to current questions on forest health

and productivity. The beneficial effects of mycorrhizal

fungi on plant growth is being used in forest nursery

management to try and improve the success of outplanted seedlings (Castellano 1996). There have been

various attempts to use nitrogen-fixing plants instead

207
of adding N fertilizers to Pacific Northwest forests

(Everett and others 1991; Trowbridge and Holl 1992).

However, the increased competition from nitrogenfixing plants when grown in mixed plantings with

conifers seems to have limited the widespread

application of this management option.

While efforts to manage specific soil organisms for the

maintenance or improvement of forest and range

productivity will continue, questions are being raised as

to whether management activities should also consider

soil organism diversity (Busse; Ingham; Moldenke - this

symposium). The answer is obvious if organisms are

listed as threatened or endangered (Molina - this

symposium), or are considered important in the

restoration of degraded ecosystems (Belnap; Bennett;

Pellant - this symposium). How important soil

organism diversity is to overall ecosystem diversity and

stability, and whether various management practices

(for example, clearcutting or prescribed fire) are

detrimental to soil biodiversity is not as clear.

As noted above, any management practice which

changes the amount and quality of soil organic matter,

raises soil pH, removes nitrogen, or compacts the soil

will affect soil organism populations and diversity.

Many studies have found reductions in numbers and

types of soil organisms immediately following or several

years after timber harvesting (Spence; Visser-this

symposium). While such short-term results are

important, managers also need to know: 1) how

different these management-related changes in soil

organism populations are from those occurring after

natural disturbances (fire, blowdown), 2) whether the

pattern of soil organism succession is changed as the

new stand develops, and if so, 3) do these changes

affect soil productivity or sustainability. This is

illustrated in Figure 1, where timber harvesting was

assumed to reduce soil organic matter content by

different amounts, which resulted in long-term

changes in soil organism diversity. Conventional

wisdom is that the soil with the greatest number of soil

organism types or species after harvest will result in a

more diverse soil ecosystem as the stand ages, and be

more productive and sustainable than the one with

lower initial soil diversity. However, the ability or

"redundancy" of many different soil organisms to carry

out important soil functions and processes makes this

assumption questionable. Does a reduction of 25 in

some measure or index of soil diversity mean that soil

respiration and nitrogen mineralization will be less, or

that soil organisms will be less able to adapt to possible

future climate change? How much can soil organism

diversity be reduced before such changes occur?

Unfortunately, there is very little information on the

impact initial changes in soil organisms have on longterm ecosystem diversity, and how important organism

redundancy is for soil processes. Spence (this

symposium) reported that arthropod populations in a

mixed aspen-conifer soil forest one and two years after

a wildfire were much different than in an adjacent

clearcut, but these differences were not evident after

30 years. In contrast, rhizosphere changes in

microflora and microfauna were still present ten years

after Douglas-fir trees were injected with chloropicrin or

methylisothiocyanate to control Phellinus weirii root rot

(Ingham and Thies 1997). Moldenke (this symposium)

stressed the importance of such long-term studies to

assess management effects on soil organism

populations and diversity.

Figure 1 —Possible changes in soil biodiversity as related to

changes in soil organic matter after timber harvesting: A.

Higher soil organic matter content, B. Lower soil organic

matter content.

BOTTOM LINE

The overwhelming consensus of this symposium was

that forest and range management activities do have

pronounced effects on soil organisms. However, it was

not clear if these management-induced changes in

population numbers and diversity are long-term, and if

so, how important are they for soil productivity and

stability. Until more information becomes available, it

208
would seem prudent to manage in ways which maintain

soil organism populations and diversity, and is

consistent with overall management objectives.

Fortunately, such management activities are usually

beneficial to both soil organisms and other components

of forest and range ecosystems. The keys to such

management is to maintain or increase the amount of

organic matter in or added to the soil, and reduce soil

compaction (Busse; Ingham - this symposium). Both of

these activities will favor soil organism populations and

diversity, and maintain or increase site productivity.

Consequently, managers should ask the questions: 1)

how will this management activity affect soil organic

matter levels and soil compaction now and in the

future, and 2) can organic matter losses and

compaction be reduced (for example - no dozer piling,

the use of forwarders, fencing streams) and still

accomplish management objectives. The current focus

on ecosystem management with it's emphasis on a

wide variety of successional stages, stand densities

and plant communities across a landscape should

provide various litter types and soil microsite conditions

to promote soil organism diversity.

LITERATURE CITED

Amaranthus, M.P.; Page-Dumroese, D.; Harvey, A.;

Cazares, E.; Bednar, L.F. 1996. Soil compaction

and organic matter affect conifer seedling

nonmycorrhizal and ectomycorrhizal root tip

abundance and diversity. USDA Forest Service,

Pacific Northwest Research Station, Portland, OR.

PNW-RP-494.

Azcon-Aguilar, C; Barea, J.M. 1992. Interactions

between mycorrhizal fungi and other rhizosphere

microorganisms. In: Allen, M.F. (ed.). Mycorrhizal

functioning: an integrative plant-fungal process.

Chapman and Hall, New York, NY: 163-198.

Binkley, D. 1991. Connecting soils with forest

productivity. In: Harvey, A.E.; Neuenschwander,

L.F. (compilers). Proceedings - Management and

productivity of Western-Montane forest soils.

General Technical Report INT-280. U.S.

Department of Agriculture, Forest Service,

Intermountain Research Station, Ogden, UT: 66-69.

Castellano, M.A. 1996. Outplanting performance of

mycorrhizal inoculated seedlings. In: Mukerji, K.G.

(ed.). Concepts in mycorrhizal research. Kluver

Academic Press, the Hague: 223 - 301.

Colinas, C; Perry, D.; Molina, R.; Amaranthus, M.

1 994. Survival and growth of Pseudotsuga

menziesii seedlings inoculated with biocide-treated

soils at planting in a degraded clearcut. Canadian

Journal of Forest Research 24:1741-1749.

DeBano, L.F.; Neary, D.G.; Ffolloitt, P.F. 1998. Fire's

effects on ecosystems. John Wiley and Sons, New

York. 333 p.

Devall, B. (editor). 1993. Clearcut: the tragedy of the

industrial forest. Sierra Club, Earth Island Press,

San Francisco, CA., 291 p.

Everett, R.; Zabowski; McColley, P. 1991. Vegetative

restoration of western-montane forest soils. In:

Harvey, A.E.; Neuenschwander, L.F. (compilers).

Proceedings - Management and productivity of

Western-Montane forest soils. General Technical

Report INT-280. U.S. Department of Agriculture,

Forest Service, Intermountain Research Station,

Ogden, UT: 161-166.

Harvey, A.E. 1994. Integrated roles for insects,

diseases and decomposers in fire-dominated forests

of the Inland Western United States: past present

and future forest health. Journal of Sustainable

Forestry 2:211-220.

Hawksworth, D.L.; L.A. Mound. 1991. Biodiversity

databases: The crucial significance of collections.

In: Hawksworth, D.L. (ed.). The biodiversity of

microorganisms and invertebrates: Its role in

sustainable agriculture. CAB International,

Redwood Press Ltd., Melksham, UK: 17-29.

Hessburg, P.F.; Mitchell, R.G.; Filip, G.M. 1994.

Historical and current roles of insects and pathogens

in eastern Oregon and Washington forested

landscapes. General Technical Report PNW-GTR327. U.S. Department of Agriculture, Forest

Service, Pacific Northwest Research Station,

Portland, OR. 72 p.

Hubbard, R.K.; Lowrance, R.R. 1994. Riparian forest

buffer system research at the Coastal Plain

Experiment Station, Tifton, GA. Water, Air and Soil

Pollution 77:409-432.

Ingham, E.R.; Thies, W.G. 1997. Changes in

rhizosphere microflora and microfauna 10 years

following Douglas-fir live tree injection with

chloropicrin or methylisocyanate. Canadian Journal

of Forest Research 27:724 - 731.

Johnson, D.W. 1992. Effects of forest management

on soil carbon storage. Water, Air and Soil Pollution

64:83-120.

Jurgensen, M..F.; Harvey, A.E.; Graham, R.T.; PageDumroese, D.S.; Tonn, J.R.; Larsen, M.J.; Jain, T.B.

1997. Impacts of timber harvesting on soil organic

209
matter, nitrogen, productivity and health of Inland

Northwest forests. Forest Science 43:234-251.

Kennedy, A.C; Gewin, V.L. 1997. Soil microbial

diversity: present and future considerations. Soil

Science 162:607-617.

Moore, J.A.; Mika, P.G.; Schwandt, J.W.; Shaw, T.

1994. Nutrition and forest health. In: Baumgartner,

D.M.; Lotan, J.E.; Tonn, J.R. (compilers). Interior

cedar-hemlock-white pine forests: ecology and

management. Cooperative Extension Publication,

Washington State University, Pullman: 173-176.

Naeem, S.; Hakansson, K.; Lawton, J.H.; Crawley.

M.J.; Thompson, L.J. 1996. Biodiversity and

plant productivity in a model assemblage of plant

species. Oikos 76:259-264.

Powers, R. 1991. Are we maintaining the productivity

of forest lands? Establishing guidelines through a

network of long-term studies. In: Harvey, A.E.;

Neuenschwander, L.F. (compilers). Proceedings -

Management and productivity of Western-Montane

forest soils. General Technical Report INT-280.

U.S. Department of Agriculture, Forest Service,

Intermountain Research Station, Ogden, UT: 70-81.

Startsev, N.A.; McNabb, D.H.; Startsev, A.D. 1998.

Soil biological activity in recent clearcuts in westcentral Alberta. Canadian Journal of Soil Science

78:69-76.

Symstad, A.J.; Tilman, D.; Wilson, J.; Knops, J.M.H.

1998. Species loss and ecosystem functioning:

effects of species identity and community

composition. Oikos 81:389-397.

Tilman, D. 1996. Biodiversity: population versus

ecosystem stability. Ecology 77:350-363.

Trowbridge, R.; Holl, F.B. 1992. Early growth of

lodgepole pine after establishment of alsike cloverRhizobium nitrogen-fixing symbiosis. Canadian

Journal of Forest Research 22:1089-1093.

210
ABSTRACTS OF PAPERS NOT SUBMITTED
SOIL FOODWEBS

Elaine Ingham1

ABSTRACT

The soil foodweb is important because plant productivity increases as soil foodweb complexity increases. An incredible

wealth of interactions, complex feedbacks, and less-than-obvious subtleties are hidden in this simple statement. In

order to reduce the cost of growing row crops, herbs, shrubs, orchards, and forests, we need to manage soil

organisms. We need to be aware that soil organisms exist, to know which ones help which plants grow, and which

ones harm those plants. In order to manage those organisms, we need to be able to measure them.

The soil foodweb includes several trophic levels. The first trophic level includes plants and other organisms that fix

carbon dioxide into biomass using sunlight for energy. The second trophic level are the bacteria, fungi, root-feeding

nematodes and arthropods, which convert plant material into their own biomass, metabolites, and respiratory products.

The third trophic level organisms, are the predators of bacteria and fungi, including protozoa, bacterial- and

fungal-feeding nematodes, and fungal-feeding microarthropods. Several higher trophic levels may exist (maybe as many

as fourth, fifth, sixth, and higher levels) which are predators of the predators, and include many forms of arthropods,

worms, insects, and aboveground birds, and mammals.

Bacteria and fungi immobilize N in their biomass, and along with soil organic matter, which these organisms produce,

contain the majority of the nutrients in a healthy soil. These nutrients need to be moved from the organisms into

plant-available forms at the speed and in the places the plants need them. If mineralization into plant-available forms is

more rapid than plants require, the nutrients can be leached or volatilized from the system. Too little mineralization, and

the plants will be nutrient limited. The organisms that eat bacteria and fungi turn the nutrients immobilized in bacterial

and fungal biomass into plant-available forms. Thus, protozoa, nematodes and microarthropods, are critical players in

plant production. But these predators can overeat bacteria and fungi, and so those predators that eat predators of

bacteria and fungi are important for keeping balance.

Along with nutrient cycling, a complex soil foodweb keeps disease-causing organisms in check. It also improves soil

aggregation, ease of root penetration and water infiltration, and increases decomposition of a greater variety of plant

materials, and anthropogenic compounds. As a greater diversity and a greater number of functional groups of

organisms occur in the soil, competition with disease-causing organisms is greater.

The soil foodweb greatly influences the production of soil aggregates, soil pores and soil channels. An analogy to

building a brick house is useful here. To build bricks, straw and sand have to stick together. Then the bricks are held

together with mortar to form walls. The house has structure when the walls are arranged in certain patterns. Different

organism groups in the soil foodweb do the same for soil structure. Bacteria glue the clays, silts and sands together

into microaggregates. Microaggregates are bound together by fungal hyphae, root hairs and roots. The structure of the

rooms are made by the arthropods, insects and earthworms. Only when all the organisms are present and active can

roots and water move into the soil with ease.

Soil compaction is a major disturbance of soil aggregation. When heavy machinery moves across the soil, the pores and

channels within the soil collapse. As the space in the soil is reduced, the organisms living in those spaces are killed.

The applied force and strength of the soil aggregates determines the extent of damage. The greater the force and the

stronger the aggregates, the less damage will occur to the soil. One way of looking at this is to say, if a highly complex

soil foodweb is present, better built soil aggregates will be present, and the bigger the truck is that can be driven on the

land. Alternatively, an assessment of the soil type and the soil foodweb should yield an understanding of how fragile

soil structure is, and thus what kinds of vehicles can be used without destroying critically important functions in that

soil.

It is critical to understand the benefitis a complex soil foodweb, can give to an ecosystem. It is important to

understand what disturbances do to this foodweb, and how to repair that damage if it proves necessary to allow those

disturbances to occur.

1Elaine Ingham is an Associate Professor in the Department of Botany and Plant Pathology at Oregon State University, Corvallis, OR

97331.

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NUTRIENT & SOIL ORGANISM POPULATION DYNAMICS-THE BASIS OF ECOSYSTEM

STABILITY

John C. Moore1

ABSTRACT

There are three objectives to this presentation. The first objective is to make explicit connections between the plants

and decomposers. Much of ecological theory is based on plant-herbivore-predator models. Differences between

terrestrial and aquatic ecosystems notwithstanding, most primary production is not consumed by herbivores, rather it is

decomposed by microbes. A thorough treatment of the detritus-decomposer-consumer pathway and linkage of this

pathway to the plant-herbivore pathway has yet to be conducted.

The second objective is to formalize an observed connection between changes in community structure, nutrient

dynamics and the mathematical stability of an ecosystem (local stability and resilience). My research has demonstrated

that communities are composed of stable assemblages of species that share common habitats, food, and life history

traits. Within soils, these assemblages are the pathways of trophic interactions and nutrient flux from plant roots,

bacteria and fungi. The resilience of these pathways depends on the physiologies of the species within them. The

bacterial pathway is more resilient and mineralizes nitrogen at a faster rate than the fungal pathway. Disturbance

induces a shift in activity from the fungal to the bacterial pathway resulting in increased decomposition and nitrogen

mineralization rates, and thus lower nutrient retention. These changes in nutrient dynamics have been shown to

destabilize mathematical models of the ecosystems.

The final objective is to explore the interplay between diversity and function. Linking function to biodiversity requires

studying both of them during community responses to disturbance. Traditional approaches identify a duality to diversity

and to the mechanisms that govern it, i.e., animal diversity is linked to plant diversity. Where detritus and detritivores

fit in is less clear, but will have to be resolved to answer how biodiversity affects ecosystem function. Within soils,

there appears to be a balance between taxonomic specificity and redundancy in function. Key taxa are needed for

decomposition to proceed, but, following major disturbance, functional diversity recovers before species diversity.

My overall theme is that the energetic organization of communities, as defined by nutrient and population dynamics,

forms the basis of ecosystem stability. Five propositions are presented.

PROPOSITION I

The utilization of resources by species is highly diverse and compartmentalized. The ability to shift from one food

source to another and between habitats are adaptations that have implications beyond the ability of an organism to

persist, but are important at the system and community levels as well.

PROPOSITION II

Ecosystems are organized as a series of quasi-independent interacting energy pathways. The pathways are made up of

species that share similar habitat requirements and complimentary life history traits. As a result of this, matter is

processed and energy is transferred within the pathways at distinct rates. In soils, these pathways originate from plant

roots and detritus. The detritus pathway is divided into a bacterial pathway and fungal pathway.

PROPOSITION III

The species composition of the plant community and decomposer communities are tightly coupled through plant litter

quality. Species within the fungal and bacterial pathways have adapted to the chemical composition of the substrates

produced by the producer (plant root) pathway.

The importance of litter quality to decomposition is well established. The relationship between plant and soil

communities is a function of litter quality. For example, pine forests possess a distinct decomposer community

compared to a sage meadow.

1

John C. Moore is Associate Professor, Department of Biological Sciences, University of Northern Colorado, Greeley, CO 80639.

212
PROPOSITION IV

Disturbances alter energy pathways and hence the pattern of energy flow within ecosystems. Nutrient availability to

plants and retention within the ecosystem are affected by changes in the activity of energy pathways. In soil,

disturbances have a disproportionately large adverse affect on the fungal pathway compared to the bacterial pathway.

Nutrient availability and retention were either tied to the activities of the fungal pathway, or is a function of the

combined activities of the bacterial and fungal pathways, and how synchronous these activities are with plant growth.

Links between nutrient dynamics and population dynamics are explored through the life histories of the organisms.

Evidence is presented from agricultural studies and sites under managed grazing.

PROPOSITION V

The distribution of biomass within a system (trophic structure), and how the system responds to a disturbance

(dynamics and resilience) is a function of the rate of energy input and the rate at which energy is transferred among

organisms within the system. The relationship between soil communities and how they respond to disturbance to

nutrient dynamics is explored.

A series of hypotheses is proposed as follows:

H1: Decomposition rates and nitrogen mineralization rates increase following physical disturbances of soil due to a shift

in function away from the fungal pathway to the bacterial pathway.

H2: Nutrient retention decreases in disturbed soils due to changes in the community from a stable to unstable pattern

of interactions.

H3: Functional diversity is more resilient than taxonomic diversity due to the high degree of redundancy in function

among taxa.

H4: Nutrient retention is enhanced by increased species diversity within the fungal rather than the bacterial pathway

due the greater resilience of the bacterial pathway and to the stronger coupling of fungi to resistant plant substrates.

213
FOREST MANAGEMENT INTERACTIONS WITH SOIL ORGANISMS

James R. Boyle1

ABSTRACT

Soil organisms, macro-, meso- and micro-, are involved in the dynamics of matter and energy in forest ecosystems,

influencing plant growth, soil properties and other ecosystem components. Critical ecosystem processes, including

transformations of organic matter to yield plant-available nutrient ions, and maintenance of soil porosity, are controlled

by soil organism activities.

Forest management activities (and "natural" disturbances) alter the physical and chemical environments of soil

organisms. Amount and composition of organic matter in forest floors and within soils are changed as plant cover is

altered and residues are added. Solar radiation inputs, temperature fluctuations, and water dynamics are influenced by

forest management. Changes in organic matter alter availability of energy, carbon and other elements on which soil

organism metabolism depends. Diversities and community compositions of soil organisms are therefor changed, as are

ecosystem processes.

Forest scientists and managers are well-advised to consider potential changes in critical or limiting ecosystem

processes, including those mediated by soil organisms, as they plan, execute and evaluate impacts of forest

management. Sustainability of forest ecosystems, simple or complex, intensively- or extensively-managed, is

dependent on basic ecosystem processes and components, not on flows of human- defined forest products and values.

Soil organisms are among the "hidden" ecosystem characteristics that must be maintained.

1

James R. Boyle is Professor of Forestry and Soil Ecology, Forest Resources Department, College of Forestry, Oregon State

University, Corvallis, OR 97331

214
Acknowledgments

Symposium Sponsors:

USDA Forest Service, Pacific Northwest Region

USDI Bureau of Land Management

USDA Natural Resources Conservation Service, Soil Quality Institute

Oregon State University

Oregon Society of Soil Scientists

Cover

Graphics Design by Janine Koselak,

USDI, Bureau of Land Management

National Applied Resources Services Center (NARSC)

Lakewood, Colorado

The Editors express special appreciation to Izella Stuivenga, Corvallis,

Oregon, for her excellent work in compilation of the manuscripts into a

single document. Her attention to details and her ability to incorporate the

many styles of tables and figures into a consistent format is especially

appreciated.

215
Meurisse, Robert T.;Ypsilantis, William, G.; Seybold, Cathy, tech. eds. 1999. Proceedings,

Pacific Northwest forest and rangeland soil organism symposium; 1998 March 17-19;

Corvallis, OR. Gen. Tech. Rep. PNW-GTR-461. Portland, OR: U.S. Department of

Agriculture, Forest Service, Pacific Northwest Research Station. 215 p.

To provide a review of the state-of-the-science about soil organisms for management, a symposium was convened to address soil organism functions and processes and management effects

on soil organisms and processes, and to describe some roles of soil organisms in restoration and

applications for land management. The focus of the symposium was on Pacific Northwest forest

and rangeland ecosystems.

Keywords: Soil organisms, soil quality, nutrient cycling, nitrogen transformation, foodwebs,

ecosystem, bacteria, fungi, anthropods, root diseases.

The Forest Service of the U.S. Department of Agriculture is dedicated to the

principle of multiple use management of the Nation's forest resources for sustained yields of wood, water, forage, wildlife, and recreation. Through forestry

research, cooperation with the States and private forest owners, and management of the National Forests and National Grasslands, it strives—as directed

by Congress—to provide increasingly greater service to a growing Nation.

The U.S. Department of Agriculture (USDA) prohibits discrimination in all its

programs and activities on the basis of race, color, national origin, gender,

religion, age, disability, political beliefs, sexual orientation, or marital or family

status. (Not all prohibited bases apply to all programs.) Persons with disabilities who require alternative means for communication of program information

(Braille, large print, audiotape, etc.) should contact USDA's TARGET Center

at (202) 720-2600 (voice and TDD).

To file a complaint of discrimination, write USDA, Director, Office of Civil

Rights, Room 326-W, Whitten Building, 14th and Independence Avenue, SW,

Washington, DC 20250-9410 or call (202) 720-5964 (voice and TDD). USDA

is an equal opportunity provider and employer.

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